x>EPA
United States
Environmental Protection
Agency
Environmental Research
Laboratory
Duluth MN 55804
EPA-600/3-78-076
Mugusi
Research and Development
Proceedings
of the First
and Second USA - USSR
Symposia on the Effects
of Pollutants
Upon Aquatic
Ecosystems
Volume
Duluth , Minnesota
USA Symposium
October 21 - 23 ,1975
Volume II
Borok , Jaroslavl Oblast
USSR Symposium
June 22 -26 ,1976
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RESEARCH REPORTING SERIES
Research reports of the Office of Research and Development, U.S. Environmental
Protection Agency, have been grouped into nine series. These nine broad cate-
gories were established to facilitate further development and application of en-
vironmental technology. Elimination of traditional grouping was consciously
planned to foster technology transfer and a maximum interface in related fields.
The nine series are:
1. Environmental Health Effects Research
2. Environmental Protection Technology
3. Ecological Research
4. Environmental Monitoring
5. Socioecon'omic Environmental Studies
6. Scientific and Technical Assessment Reports (STAR)
7 Interagency Energy-Environment Research and Development
8. "Special" Reports
9. Miscellaneous Reports
This report has been assigned to the ECOLOGICAL RESEARCH series. This series
describes research on the effects of pollution on humans, plant and animal spe-
cies, and materials. Problems are assessed for their long- and short-term influ-
ences. Investigations include formation, transport, and pathway studies to deter-
mine the fate of pollutants and their effects. This work provides the technical basis
for setting standards to minimize undesirable changes in living organisms in the
aquatic, terrestrial, and atmospheric environments.
This document is available to the public through the National Technical Informa-
tion Service, Springfield, Virginia 22161.
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EPA-600/3-78-076
August 1978
PROCEEDINGS OF THE FIRST AND SECOND USA-USSR
SYMPOSIA ON THE
EFFECTS OF POLLUTANTS UPON AQUATIC ECOSYSTEMS
Volume I: Duluth, Minnesota, USA Symposium
October 21-23, 1975
Volume II: Borok, Jaroslavl Oblast, USSR Symposium
June 22-26, 1976
ENVIRONMENTAL RESEARCH LABORATORY
OFFICE OF RESEARCH AND DEVELOPMENT
U.S. ENVIRONMENTAL PROTECTION AGENCY
DULUTH, MINNESOTA 55804
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VOLUME I
PROCEEDINGS OF THE FIRST USA-USSR
SYMPOSIUM ON THE
EFFECTS OF POLLUTANTS UPON AQUATIC ECOSYSTEMS
October 21-23, 1975
Duluth, Minnesota
USA
Edited by
Donald I. Mount
ENVIRONMENTAL RESEARCH LABORATORY-DULUTH
OFFICE OF RESEARCH AND DEVELOPMENT
U.S. ENVIRONMENTAL PROTECTION AGENCY
DULUTH, MINNESOTA 55804
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DISCLAIMER
This report has been reviewed by the Environmental Research Labora-
tory-Duluth, U.S. Environmental Protection Agency, and approved for publi-
cation. Mention of trade names or commercial products does not constitute
endorsement or recommendation for use.
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FOREWORD TO VOLUME I
These proceedings result from the first symposium held by Project
II-1.3 of the Joint US-USSR Committee on Cooperation in the field of
Environmental Protection, established in May 1972.
Broad review papers were included in the symposium in order to
acquaint scientists from each country with the water pollution perspective
upon which current programs are based. There are differences and therein
lies the value of meeting together.
IV
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PREFACE TO VOLUME I
This volume contains nineteen of twenty papers presented at the First
US-USSR Symposium on the Effects of Pollutants on Aquatic Ecosystems. All
papers, ten from each side, were given in English or Russian at Duluth,
Minnesota, USA between October 21 and 23, 1975, at the Environmental Re-
search Laboratory-Duluth of the U.S. Environmental Protection Agency.
The three-day symposium climaxed a two-week visit by the Soviets as
part of a working group on "Effects of Pollutants on Aquatic Ecosystems
and Allowable Levels of Pollution". This is one of 40 working groups
established in a five-year international "Agreement on Cooperation in the
Field of Environmental Protection Between the United States of America
and the Union of Soviet Socialist Republics", signed May 23, 1972 at the
Moscow Summit Meeting.
During their two-week visit, Dr. Donald I. Mount, Director of the
Duluth Laboratory and the U.S. Project Leader of the Working Group,
brought the six visiting Soviet scientists to water pollution research
laboratories in Cincinnati (Ohio), Columbia (Missouri), and Chicago
(Illinois), as well as Duluth (Minnesota), to observe the American faci-
lities and exchange technologies with U.S. researchers.
At the end of the symposium, Dr. Mount and Professor Nikolay V.
Butorin, Soviet Project Leader of the Working Group, signed an agreement
outlining future activities of the group, including a reciprocal visit by
American scientists to the USSR in June 1976. Both countries pledged
their continued commitment to cooperative environmental activities.
The publication of these proceedings is in accordance with that agree-
ment signed October 23, 1975 by Dr. Donald I. Mount and Professor Nikolay
V. Butorin.
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INTRODUCTION
The Joint US-USSR Agreement on Cooperation in the Field of Environ-
mental Protection was established in May 1972. These Proceedings result
from one of the projects, Project 02.02-1.3 "Effects of Pollutants Upon
Aquatic Ecosystems and Permissible Levels of Pollution".
The project derives its strength and value from the idea that it is
important for scientists who share a concern for the environment to take
a broad look at the subject, and to exchange views with their colleagues.
It is hoped by this process to help assure that the overall goals are not
lost in the clutter of minutia. These Proceedings cover Working Group
II's first meeting of specialists October 21-23, 1975 at Duluth,
Minnesota.
VI
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FIGURES
Section Page
1 Processes of mineralization in solutions with triethyl
stannic chloride 3
1 Processes of mineralization in solutions of pryror-70 . . 7
1 Processes of nitrification in solutions of pryror-70 ... 8
4 Toxicity curve demonstrating lethal threshold concentra-
tion (LTC) -. . 38
4 Tolerance of (A) fathead minnows (Pimephales promelas) and
(B) goldfish (Carassius auratus) to hydrogen sulfide at
different temperatures 41
4 Tolerance of bluegills (Lepomis macrochirus) to molecular
cyanide (HCN) at various temperatures 42
7 Platform from which oyster trays are suspended 70
7 Oyster tray 71
7 The structure of a natural diatom community /3
7 The structure of a diatom community under the effects of
pollution high in nutrients 75
7 The structure of a diatom community under the effects of
toxic conditions 76
7 Invertebrate sampler 78
9 The relationship between mean total nitrogen concentration
in streams and land use in the Eastern United States ... 93
9 The relationship between mean total phosphorus concentra-
tions in streams and land use in the Eastern United
States 93
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Section page
9 The relationship between total phosphorus loading, lake
morphometry, and lake trophic condition for selected phos-
phorus-limited lakes in the Northeastern United States . . 94
9 The relationship between total phosphorus loading, lake
morphometry, and lake trophic condition for selected nitro-
gen-limited lakes in the Northeastern United States ... 94
12 Map-diagram of Rybinskiy reservoir 122
12 Average values for phytoplankton production for 13 years
in the Rybinskiy reservoir 124
12 Average values of the number of bacteria according to data
for 2 years of standard observations 126
13 Change in component values in time—experiment 1 141
13 Change in component values in time—experiment 2 142
13 Change in component values in time—experiment 3 143
13 Change in component values in time—experiment 4 144
13 Change in component values in time—experiment 5 145
13 Change in component values in time—experiment 6 146
13 Change in component values in time—experiment 7 147
13 Change in component values in time—experiment 8 148
13 Change in component values in time—experiment 9 149
13 Change in component values in time—experiment 10 .... 150
13 Dependence of the specific growth rate of bacteria on
the nitrate nitrogen N concentration 151
14 Block diagram of basic functional links in the reser-
voir 154
14 Change in numbers of a species with exposure to a toxic
substance 155
16 Organization of investigational divisions at the Fish-
Pesticide Research Laboratory, Columbia, Mo 168
vm
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Sect i on Page
16 Multiple concentration, flow-through diluter with con-
trolled light and temperature used to determine sublethal
effects of toxaphene on growth and reproduction of channel
catfish 173
16 Effects of toxaphene on backbone structure of fathead
minnows 1/5
16 Effects of toxaphene on backbone structure of channel
catfish 1/7
16 Comparison of viability and hatch in eggs from brook
trout exposed continuously or by simulated usage pattern
to TFM 178
17 Sequence of changes in biological indices of Lebistes
reticulatus (P) in phenol concentrations 188
17 Effect of phenol on bioelectrical activity of peripheri-
cal, central nerve system and neuromuscular conjunction
of fish 189
17 Accumulation of phenol in model communities 191
18 The St. Lawrence Great Lakes with interstate and inter-
national boundaries 195
18 Average number of pounds and kilograms of fish produced
per acre and hectare by the commercial fishery of the
Great Lakes for 10-year intervals 200
18 The St. Lawrence Great Lakes showing canals between
central Atlantic Ocean and the lakes 202
19 Range of the sea lamprey 219
19 Life cycle of the sea lamprey 220
19 The mouth of the sea lamprey 221
19 Production of lake trout, 1930-66, and number of sea lamp-
reys caught in index streams in Lake Superior, 1953-69 . . 222
19 The Great Lakes 223
19 Sea lamprey catch from eight streams tributary to Lake
Superior 225
IX
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TABLES
Section Page
1 Permissible concentrations of toxic agents for bacteria
(saprophytic and nitrifying) 6
4 Application factors for various compounds 44
4 Application factors for H2S with juveniles of six fish
species 45
4 96-hour LC50 of eight fish species to pesticides 45
7 Summary of Catherwood diatometer readings at Station 1 . . /4
10 Characteristics of nitrogen exchange in this year's
group of perch 109
10 Content of Free amino acids in organs and tissues of IDE
yearlings 110
10 Effect of living blue-green algae on content of total
thiamine and thiaminase activity in liver and intestine
of fish Ill
10 Effect of blue-green algae on content of total thiamine
activity in liver and intestine of fish 112
12 Production of bacterial biomass in the Rybinskiy Reser-
voir 127
12 Destruction of organic matter in the water 127
12 Main elements of the balance of organic matter in the
Rybinskiy Reservoir 129
13 Values of components in dry weight at the steady state . . 134
13 Destruction of phenol when limited by nitrogen 137
14 Comparison of harmless concentrations of substances ... 158
17 Resistance of aquatic invertebrates of phenol 183
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Section Page
18 Dimensions of the Great Lakes 197
18 Estimated average concentration of dissolved chemical
constituents in the Great Lakes prior to 1900 198
18 Estimated population in Great Lakes Basin, 1900-1960 ... 203
18 Summary of fish species decline in the Great Lakes by
year, lake, and current commercial production 205
18 Estimated average concentrations of dissolved chemical
constituents in the Great Lakes in 1925 with percentage
change after 1900 in parentheses 206
18 Fish-species introductions in the Great Lakes by year
and lake, first year of commercial significance, and
current production 208
18 Fish species in the Great Lakes that have experienced
severe declines, lake affected, and suspected cause of
decline 209
18 Estimated average concentrations of dissolved chemical
constituents in the Great Lakes in 1950 with percentage
change since 1925 in parentheses 210
18 Estimated average concentrations of dissolved chemical
constituents in the Great Lakes in 1970 with percentage
change since 1950 in parentheses 214
XI
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ACKNOWLEDGEMENTS
The thorough and unselfish efforts of Ruby Johnson and Mildred Medlin
made possible the visit of the Soviet scientists to the U.S. and the
symposium. Nikki Vick's professional talents of editing and Elaine
Fitzback's assistance with the translations have made possible the publi-
cation of the Proceedings. All of the participants gave freely of their
time to prepare and present the papers.
As in any cooperative effort, many people share the success of this
effort.
Xll
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CONTENTS
Foreword iv
Preface v
Introduction vi
Figures vii
Tables x
Acknowledgements xii
1. Permissible Pollution Levels of Water Bodies
N.S. Stroganov 1
2. A Brief History of Water Pollution Research in the
United States
Clarence M. Tarzwell 11
3. Characteristics of the Moscow River Water Quality According
to Hydrobiological Indices
V.A. Abakumov and G.L. Margolina 32
4. Endpoints in Bioassay
Lloyd L. Smith, Jr 36
5. Physiological-Biochemical Aspects of Water Toxicology
V.I. Lukyanenko 47
6. Bioenergetic and Other Considerations Important in the
Study of Water Quality Influences on Fish Growth
Peter Doudoroff 55
7. Monitoring the Condition of Flowing Waters by Biological
Organisms
Ruth Patrick 68
8. The Role of Algae in the Pollution of Reservoirs and
Problems of Controlling their Numbers
V.G. Khobot'ev 82
9. Eutrophication in the United States: Past-Present-Future
A.F. Barsch, K.W. Malueg, C.F- Powers, and T.E. Maloney . . 87
10. Determining Threshold and Biologically Dangerous Concentra-
tions of Blue-Green Algae in Water Bodies
L.A. Sirenko, A. Ya. Malyarevskaya, and T.I. Birger .... 105
xm
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11. Toxic Organic Residues in Fish
Howard E. Johnson 115
12. Balance of Organic Matter in the Ecosystem of the Rybinskiy
Reservoir
V.I. Romanenko 121
13. The Importance of Trophic Bonds in the Bacterial Destruction
of Organic Matter
P.P. Umorin 132
14. Simulation of Potential Pollutant-Caused Changes in the Eco-
system, Resulting from the Sensitivity of Aquatic Organisms
to Toxicants
N.S. Stroganov 152
15. Fish-Population Studies in the Ohio River
William C. Klein 161
16. Registration of Pesticides: Considerations in Conducting
Aquatic Toxicity Tests
Richard A. Schoettger 166
17. Experimental Research on Phenol Intoxication of Aquatic
Organisms and Destruction of Phenol in Model Communities
M.M. Kamshilov and B.A. Flerov 181
18. History of Changes in Fish Species of the Great Lakes
John F. Carr 193
19. Sea Lamprey (Petromyzon marinus Linnaeus) in the Saint
Lawrence Great Lakes of North America: Effects, Control,
Results
Carlos M. Fetterolf, Jr 219
xiv
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SECTION 1
PERMISSIBLE POLLUTION LEVELS OF WATER BODIES
N.S. Stroganov*
The question of permissible levels of pollution for water bodies has be-
come more and more acute because the number of substances contaminating sur-
face waters has increased and treatment of discharges is expensive and corn-
lex. Among engineers and other specialists who are not biologists, the con-
cept is fairly widespread that it is possible to dump all pollutants into
water because aquatic organisms will degrade them. However, at the present
time such a concept is obviously erroneous and does not correspond to the
actual situation. Pollution levels can vary depending on the use of the
water body. Therefore, in order to talk about levels, it is necessary to
establish the requirements for which the water will be used, that is, the
requirements of the water users. The larger the body, the larger the num-
ber of water users will be. For complex and rational utilization of water,
one must take into account the requirements of many water users. If one
considers only the main requirements, meaning significance for the national
economy, then one should note the following: (1) drinking and household
water supply; (2) fishing industry; (3) agriculture (irrigation, livestock
farms, fur farms); (4) industry (food, chemical, pulp, metallurgical, pet-
roleum, chemical and others); (5) aesthetic and health purposes (sports,
tourism, recreation, etc.); (6) transportation and certain other water uses.
The quality of water can be very different for the uses mentioned. The
highest water quality is needed for drinking purposes and the fishing indus-
try, in special cases for industry (for example, the pulp industry), and
the lowest quality is adequate for water transportation.
Consequently, if one satisfies the requirements for the first two water
uses as to water quality, then all of the other uses will be protected. If
water quality is suitable only for water transportation, then the quality
of the water will be unsuitable for drinking water supplies or for the fish-
ing industry. Therefore, in order to establish the maximum permissible
level of pollution, beyond whose limits one cannot go without disrupting
the use of the water, one should: first, determine the chief water users
for multiple utilization of the water body, and, secondly, determine the
main water quality requirements. For fresh water, almost all of the more
or less large water bodies should support the interests of all water users
enumerated above, and for sea water—al I except for drinking purposes.
1-USSR, Biology Faculty of Moscow State Univ.
1
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As has already been noted, it is expedient to provide requirements for
the first of the two water uses for fresh water and only for the second for
sea water because in these cases these are the highest requirements for
water quality. We will examine the level of permissible pollution taking
into account these requirements and then the suitability or degree of pollu-
tion can be ascertained by comparing the existing water quality to the per-
missible concentrations.
Quality of water is shaped by aquatic organisms on the basis of hydro-
chemical and hydrological regimes. Toxicants in the water change the
hydrochemical composition of the surface water and have a definite effect,
depending on concentrations, on the processes determining its quality.
Pollutants discharged into surface waters gradually degrade or are
transformed to less active states. The degree and rate of breakdown de-
pend primarily on the nature of the pollutant, organisms involved in decom-
position, time and physical and chemical factors (pH, 02, salinity and
hardness).
Each of the factors enumerated can accelerate or retard decomposition.
Natural organic substances are fairly easily decomposed by bacteria, proto-
zoans, fungi and other aquatic organisms. Organochlorine pesticides and
detergents created by man and also heavy metals and long-lived radioactive
isotopes retain their toxicity for a long time and enter into the food
chain. Saprophytic and nitrifying bacteria from the Nitrosomonas and
Nitrobacter groups grow and multiply poorly in the presence of toxicants;
as a result, the process of recovery of the water is retarded. Figure 1
shows data on the effect of triethyl stannic chloride on biological oxida-
tion and nitrification. Delay in decomposition of organic substances due
to the effect of toxicants results in an increase (accumulation) of pollu-
tants. If the toxicants enter waterways (rivers, canals, etc.) in signifi-
cant concentrations (see Figure 1), then the water is polluted at great
distances from the emission source. The nature of the pollutant and, pri-
marily, its capability to be broken down by microorganisms will play a de-
cisive role in the degree of pollution.
The specific composition and number of aquatic organisms play an impor-
tant role in removing pollution of water. But they themselves are subject
to the effect of toxic agents and therefore their biological activity and
number depend on the quality and quantity of the toxicants. All of the
vital processes of aquatic organisms and, consequently, the rate of detoxi-
fication of the water medium, depend on time. In the final analysis,
aquatic organisms break down all toxicants or remove their toxicity, but in
what time period? For us it is important now that these processes occur
rapidly and completely but we can have little effect on the rate.
The physical and chemical medium has an essential meaning both for
vital activity of aquatic organisms that detoxify, and for the mode of de-
composition of toxicants (oxidation, ionization, hydrolysis, etc.).
-------
10 15 20 25
6
4
O
10 15 20
9
7
5
3
IO 15 20
Figure 1. Processes of mineralization in solutions with triethyl stan-
nic chloride. K—control; 1—0.1 mg/1; 2—1 mg/1; 3—10 mg/1;
4—50 mg/1. Along the abscissa—days of the experiment.
Key: a. Triethyl stannic chloride.
b. BPK [biokhimicheskoye potrebleniye kisloroda, BOD biochemical
oxygen demand].
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For an evaluation of a possible permissible pollution level we must
make the balance between pollutants coming into a water body and the
capability to degrade them completely or render them harmless. One can ex-
press these relationships in a diagram in the form of a balanced equation:
intake of pollutants (P) = decomposition (D) (self-purification)
+ deposition in bottom sediments (S), or P = D + S.
In order to prevent the bottom sediments from accumulating polluting
substances, we must permit only that quantity which can be decomposed
(P = D) in a unit of time. The capability of the water body to degrade
wastes plays a decisive role in this equilibrium. The higher the rate of
self-purification, the more pollutants that can be converted per unit of
time. Ideally, for maintaining water quality, it is necessary to achieve
an equality of P = D. In oligotrophic waters, such a balanced equality
essentially exists. In eutrophic waters, P > D and part of the organic sub-
stances are transferred to the bottom sediments. Cases where P < D are
not encountered in nature are difficult to find. The self-purification
capability of water can be increased several ways, primarily by increasing
the temperature and content of dissolved oxygen in the water (mixing, blow-
ing in air), and by selecting a complex of organisms. Of course, in-
creasing the assimilative capacity of a water body in a given time requires
an increase in the number of organisms which break down such toxicants
(phenols, hydrocarbons, etc.). An increase in the quantity of polluting
substances over the capability to break them down may result in an increase
in the number of saprophytic bacteria, fungi, protozoans and certain other
organisms, but assimilation is delayed and occurs after large community re-
organization. As a result, an accumulation of pollutants occurs in bottom
sediments and in the water mass which creates additional difficulties in
self-purification. Aquatic organisms play a decisive role in the equili-
brium of water. The more intensely they can convert the pollutants, the
cleaner the water will be and the larger the assimilative capacity of the
water body. Biological activity of aquatic organisms, in turn, depends on
living conditions. Toxicants introduced in any concentration will decrease
biological activity and with high concentrations completely suppress it.
Processes of growth, reproduction and effective conversion of the organism-
oxidizing agents will be affected by the toxic agents. The resulting
effect then will be determined by the nature of the toxicants and their
concentrations. Therefore, the permissible level of pollution (PLP) is
determined by the rate and degree of decomposition of the pollutant by the
aquatic organisms and they determine the maximum permissible emission (MPE)
of pollutants. A sequential connection of relationships reflecting a
balanced equilibrium is being established. Diagrammatically it can be ex-
pressed in the following way:
complex of organisms + their biological activity "^ PLP "^ MPE.
Man is primarily interested in maximum permissible emissions. Produc-
tion workers attempt to increase the size of MPE. This is economically
profitable and less troublesome. However, according to the feedback princi-
-------
pie, it destroys the structure and qualitative composition of complex
organism-decomposing agents which results in reorganization of the entire
community or its restructuring occurs simultaneously. Extinction, a de-
crease in numbers, or an increase of pollutant-tolerant organisms are the
limiting conditions for uncontrolled emission of polluting substances into
water bodies. While the quality of water for drinking purposes gets worse,
requests to discharge more pollutants increase.
The necessity has arisen to scientifically substantiate the maximum per-
missible emission (MPE) of pollutants into surface waters. It seems to me
that the scientific basis should be a balanced equality between the per-
missible level of pollution and limits on the amount of discharge.
The role of toxic agents in all of the processes of waste assimilation
is tremendous because toxicants have a great effect on life processes of
pollutant-decomposing organisms. Even saprophytic bacteria, as is seen in
Figure 1, cannot maintain necessary biological activity in the presence of
toxic agents and they themselves cannot provide initially the processes of
self-purification. One should keep in mind that the nitrifying organisms
are more sensitive to many toxicants than are the saprophytic bacteria.
They lose, or decrease their biological activity with concentrations of cer-
tain toxicants being 10—100 times less than those affecting saprophytic
bacteria (Table 1).
Substances in concentrations indicated are not completely harmless for
bacteria which mineralize organic substances. The BOD and the rate of N02
and N03 formation are somewhat smaller than in the control, but decreases
are less than 25% of the control. In order to achieve a control level, the
processes of mineralization are increased to 3—5 days and in the presence
of certain substances, a longer period is required.
Along with T.S. Balabanova we carried out tests on the breakdown of or-
ganic substances by microorganisms in a medium containing pyror-70.
A method of separate determination of BOD, N02 and N03 in closed con-
tainers was used in the first series of tests. The nutrient solution was
prepared from river water, adding glucose and peptone; (NHH^S04 of NaN02
were added for the nitrifying agents. They were incubated at 25 C.
In a second series of tests, open aquarium containers were used with 8
liters of solution (the same as in the first series). The quantity of or-
ganic substance was increased to a COM (chemical oxygen minimum) of 45--50
mg/liter of Oa- The addition of pyror-/0 somewhat increased the COM (with
300 mg/liter pyror per 10 mg 02/liter). Air was blown continuously through
the aquarium. The temperature was 18-20 C. After a certain time samples
were taken for BODs, N02 and NOs. The results obtained are shown in the
graphs in Figures 2 and 3.
Both in the closed containers (Figure 2) and in the open aquaria (Fig-
ure 3) the processes of decomposition of an organic substance were sup-
pressed by the toxic agent--pyror-70 (2-bromo-2-nitro-l,3-propanediol); the
degree of suppression was greater the higher the concentration of pyror.
5
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CT>
Table 1. PERMISSIBLE CONCENTRATIONS (tng/1) OF TOXIC AGENTS FOR BACTERIA (saprophytic
and nitrifying)
Indices
BOD
Formation of N02
Formation of N03
0
1
1
1
.1
.0
.0
2
10
1.0
1.0
3
1
0.1
0.01
4
0.1
0.01
0.1
5
10
0.01
0.1
6
0.1
0.1
0.1
0.
0.
0.
7
01
005
005
1
0
0
8
.0
.1
.1
1. Aminocolophony ch'ioroacetate
2. Pyror 400
3. Aminocolophony pentachlorophenolate (sodium salt)
4. Pentachlorophenol (precipitated A12(SO
-------
2 6 10 14 18 22 26 30 34
2.0
1.5
1.0
0.5
2.0
1.5
1.0
0.5
jl/K
6 10 14
10 14
18 22
Figure 2. Processes of mineralization in solutions of pyror-70, K—control, BOD:
1—10 mg/1; 2—1 mg/1; 3—50 mg/1; 4—100 mg/1; 5—300 mg/U 1—1 mg/1;
2—10 mg/1. Along the axis of the abscissa—days of the test.
Key: a. BOD.
-------
7
6
5
/x
/ \
/ \
/ \
- / \
/ \
; \
/ \
/ \
i \
i \
10
8
6
4
2
12 16 20 24 28 32 36 40 44 48 50
10
14 18
B 22 26 30 34 38 42 46 50
Figure 3. Processes of nitrification in solutions of pyror-70.
open aquariums with air blown in. K--control.
BOD: 1—100 mg/1; 2—10 mg/1; 3—1 mg/1;
N02: 1—10 mg/1; 2—100 mg/1; 3—1 mg/1;
N03: 1—10 mg/1; 1—1 mg/1; 3—100 mg/1.
Along the axis of the abscissa—days of the test.
Key: a. BOD.
Tests in
8
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One should give special attention to the following fact. The process
of decomposition of organic substances to complete mineralization occurs in
the presence of a toxic agent, but to accomplish this requires a great deal
of time. The shapes of the curves in Figures 2 and 3 indicates that sapro-
phytes and nitrifying agents were hardly suppressed in their activity under
the effect of pyror in the first test. The number of saprophytic organisms
is changed approximately along the same curve as the BOD, N02 and NO^. It
seems to me that this reflects the following phenomenon. Microorganisms
affected by a toxic agent die in a certain quantity. The more resistant
specimens remain and, after a certain length of time and a number of
generations, clones are produced which are resistant to pyror and which can
carry out biological oxidation and nitrification. But, for the formation
of a resistant clone, the greater the concentration of the toxic agent the
more time is required.
One should note that a delay in the process of biological oxidation or
nitrification (first and second phases) by 10-^0 days or more is in itself
an expression of pollution. If this occurs in a river in which the water
is flowing at an approximate rate of 3 km/hr, then the water is not purify-
ing (not breaking down organic substances) travels to a distance of 700-
1400 km from the emission source. This situation is intensified by the
fact that the processes of bacterial decomposition of organic substances
occurs in a strict sequence (BOD ->• formation of N02^ formation of NOs).
Therefore, according to the balanced equilibrium, the rate of pollutant
addition (P) must not be greater than the rate of decomposition (D).
In conclusion, one can formulate the following basic positions on per-
missible levels of pollution:
1. Different water uses permit different levels of water pollution.
The lowest levels are needed for drinking water supply and fisheries.
2. Organic substances are broken down by different microorganisms in a
specific sequence. Toxic substances having an injurious effect on these
microorganisms suppress the processes of mineralization more strongly, the
higher the concentration.
3. The maximum permissible emission (MPE) of pollutants into waters
must be limited by the permissible level of pollution (PLP) of a given
water at a given time. MPE, in turn, is limited by processes of self-
purification (D) in which many aquatic organisms, especially microorga-
nisms, participate. Their capability and the rate of decomposition of
pollutants (D) must be appropriate to the quality and quantity of pollu-
tants discharged.
4. Among all of the chains mentioned there must be an equilibrium of
MPE = PLP = D. If MPE > D, then the water body will be polluted.
5. One cannot make calculations of values for MPE and PLP without tak-
ing into account the peculiarities of the water body, the nature of the
pollutants and the season of the year.
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Aquatic organisms are the main active initiators of the processes of
assimilation and one must base all calculations of maximum permissible
levels of pollutants on their sensitivity and capacity.
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SECTION 2
A BRIEF HISTORY OF WATER POLLUTION RESEARCH IN UNITED STATES
Clarence M. Tarzwell
INTRODUCTION
Man has habitually discharged his wastes into streams. Because the
United States has many large streams, large amounts of waste could be
placed in them without much apparent effect. As populations in cities and
industries grew, however, some streams became open sewers and people in the
lower reaches began to complain. Typhoid fever became common as water
supplies deteriorated. By the middle of the last century conditions had
become quite bad in several areas. With little or no coordination, surveys
and studies were undertaken by people in many areas. Many different
approaches were used, and different studies were carried on concurrently,
so it is difficult to describe the research as an ongoing program. There-
fore the research for detection, evaluation, and abatement of water
pollution in the United States will be described briefly under five main
headings: (1) Water supply studies; (2) Pollution surveys and studies of
natural purification and biological indicators of pollution; (3) Treatment
of organic wastes; (4) Development, use, and standarization of bioassay
methods; and (5) Determination of water quality requirements for aquatic
life and development of water quality standards. In outlining these
activities prime consideration will be given to the most important agencies
and organizations. Early developments will be given in detail. Later work
will be summarized because in recent years research has attained such
diversity and magnitude that even a list of all the projects and their
sponsoring organizations would be too long in a review of this type. Des-
criptions of developments since 1948 will be largely confined to the
activities of the federal agency designated in Public Law 84-660 and subse-
quent federal laws dealing with water pollution.
WATER SUPPLY STUDIES
Treatment of water for domestic use may have originated in China or
India thousands of years ago. In the Bible lands alum was used for the
removal of turbidity as early as 900 B.C. In the fourth century before
Christ Hippocrates advocated the boiling and filtering of polluted water
before using it for drinking. London has been required by parliamentary
statute since 1855 to filter its water supplies through slow sand filters.
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Slow sand filters were first used in America in about 1870. The first
important modern rapid sand filtration plant was built in 1902 at Little
Falls, New Jersey.
For many years typhoid fever was a disease of prime importance.
Through circumstantial evidence it was concluded that typhoid fever was
usually associated with contaminated drinking water supplies. The
bacterium responsible for the disease was identified in 1880. During the
period 1890-1900 the incidence of typhoid fever was significantly reduced
by better sanitation and filtration of water supplies. After immunization
was developed in 1900, the occurrence of the disease decreased rapidly.
In the eighteen hundreds the aim was to make domestic water supplies
safe. When typhoid fever was conquered, some felt the push for pollution
abatement would be weakened. However, those dedicated to pollution control
pointed out that the objective was to make drinking water not only safe but
also palatable. Attention was then directed to tastes and odors,
turbidity, and color.
The first edition of the "Microscopy of Drinking Water," by George C.
Whipple, was published in 1899. This book dealt with the microscopic life
other than bacteria in fresh waters. It was a compilation of limnological
data and methods for the study of aquatic organisms. Although this book
was concerned primarily with drinking water, it did enter the field of the
natural self-purification of streams, a subject more closely associated
with sewage treatment but very significant in water supply. Many investi-
gators have studied microscopic water life, but outstanding among them are
Kent, Wolle, Stokes, Zacharias, Kofoid, West, Conn, Tilden, and Calkins.
In his book Professor Whipple assembled and integrated the findings of many
aquatic biologists—their methods, equipment, and data. In the preface to
the first edition, he mentioned especially W.T. Sedgwick of the
Massachusetts Institute of Technology. He further stated, "To Prof.
Sedgwick and Mr. Rafter water analysts are indebted for the most satis-
factory practical method for the microscopical examination of drinking
water yet devised."
It was not until the middle of the last century that the practical
aspects of the study of algae and other microscopic aquatic organisms were
recognized. At that time Hassall of London and Ferdinand Conn on the
Continent pointed out the correlation between microscopic aquatic life and
water purity. The water works departments of the cities in the north-
eastern portion of the United States were the first to make studies to
detect and identify filter-clogging algal blooms and growths of algae that
produce tastes and odors. To the Massachusetts State Board of Health
belongs the credit of having begun as early as 1887 a systematic examina-
tion of all the water supplies of the state to detect problems in their
early stages so effective control methods could be initiated. In 1889 the
State of Connecticut began a similar study, and city of Boston established
at Chestnut Hill Reservoir a laboratory for the systematic study of the
biological character of the various sources of their water supply. Algal
control methods and their use developed during the first quarter of this
century. In 1905 Moore and KeHerman used copper sulphate to eradicate
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unwanted growths of aquatic organisms. Just before the publication of
Whipple's book in 1899 and in the 28 years between the first and fourth
editions of this work, a great deal of effort was devoted to the study of
microscopic organisms in water. Outstanding among these studies were: "A
Biological Study of Lake St. Clair" in 1893 by J.E. Reighard; an examina-
tion of Lake Michigan by Henry B. Ward; and studies of the Crustacea of
Lake Mendota in Wisconsin by E.A. Birge. Biological stations were
established by a number of midwestern universities on or in the vicinity of
the Great Lakes and on the shores of smaller lakes in the Great Lakes
region.
The theological (stream) studies on the plankton of the Illinois River,
begun by Kofoid in 1894 and continued through the early years of the
present century as a part of the program of the Illinois Natural History
Survey, have been an outstanding source of information on the influence of
organic enrichment on plankton populations and the effects of these
increased growths on water supplies. The investigations of the U.S. Public
Health Service on the Potomac, Ohio, Illinois, Scioto, and upper
Mississippi Rivers have also supplied many valuable data on organic enrich-
ment, natural purification, and the growth of algae in streams receiving
sewage and other organic wastes.
The detection and elimination of pathogenic organisms are essential for
the provision of a safe drinking water supply. In their attempts to
accomplish this objective, the early bacteriologists found it very diffi-
cult to detect and quantify the pathogenic organisms in water supplies.
Because members of the coliform group are constantly present in alimentary
discharges, their presence usually indicates fecal pollution and the
possible presence of intestinal pathogens. The first test for detecting
and enumerating coliforms was developed at the New York State Department of
Health Laboratory in 1893 by Theobold Smith. After the further development
of culture methods and procedures for enumerating them and measuring the
effects of their activity, coliforms became the accepted indicator of fecal
pollution. This test became the criterion and standard method for deter-
mining the sanitary quality of a water. Workers in state health depart-
ments and water pollution laboratories improved on Smith's test and devised
better methods for sampling and culturing coliforms and evaluating and
reporting res.ults.
The U.S. Public Health Service also was prominent in these research
efforts. After the passage in 1912 of the law authorizing the service to
carry out water pollution investigations, a laboratory was established in
Cincinnati, Ohio, which was known as the Stream Pollution Investigation
Laboratory. In 1915 C.T. Butterfield joined the staff of this laboratory
as a bacteriologist. He pioneered in the development and use of coliform
tests as indicators of the sanitary quality of domestic water supplies.
These tests were accepted as the tool to be used for the estimation of
pollution and its natural purification, the evaluation of sewage treatment,
and the sanitary quality of drinking water supplies. Butterfield was also
actively engaged in the shellfish sanitation program and in the survey of
the performance of representative water-treatment plants in 31 cities along
the Ohio River and other rivers of the Midwest and the East. He and his
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small staff carried out a study of the germicidal properties of the
quaternary ammonium compounds and their use and value for sanitizing milk
and food utensils.
During the First World War methods were developed for the disinfection
of water at army posts and for military operations in the field. An inten-
sive and comprehensive study was made to evaluate the bactericidal
efficiency of free chlorine and chloramines at different residual levels.
The results of these studies were made available immediately to the army
and navy, and the results guided the military in obtaining the most
effective and economical use of chlorine for water disinfection. These
studies established a scientific basis for municipal water-chlorination
practice.
The trend of water-supply research from the 1920's into the early
1940's is indicated by the title of papers from the Cincinnati laboratory.
Representative of these are the following: "The Bacteriological Examina-
tion of Water"; "The Selection of a Dilution Water for Bacteriological
Examinations"; "Suggested Procedures for the Presumptive Test in the Deter-
mination of the Coli-aerogenes Group"; "Comparison of the Enumeration of
Bacteria by Means of Solid and Liquid Media"; "Determining the Bacteriologi-
cal Quality of Drinking Water"; "Notes on the Relation Between Coliforms
and Enteric Pathogens"; "Influence of pH and Temperature on Survival of
Coliforms and Enteric Pathogens When Exposed to Free Chlorine"; "Relative
Resistance of Escherichia coli and Eberthella typhi to Chlorine and
Chloramine"; Influence of pH and Temperature on Survival of Coliforms and
EnteHc Pathogens When Exposed to Chloramine"; "Bactericidal Properties of
Free and Combined Available Chlorine"; Bactericidal Properties of
Chloramines and Free Chlorine in Water"; and "Bactericidal Efficiency of
Quaternary Ammonium Compounds."
During the 1920's and 1930's the fecal coliform tests were used in
conjunction with the BOD (Biochemical oxygen demand) in all pollution sur-
veys. Methods for conducting these tests have been included in "Standard
Methods for the Examination of Water and Wastewater" by the American Public
Health Association, ejt aj_. for many years.
At the end of the Second World War, membrane filter techniques were
developed, compared with earlier procedures, and standardized. In this
period studies were made to develop methods for distinguishing human coli-
forms from those of other animals. In the early fifties viruses in water
supplies were studied at the Robert A. Taft Sanitary Engineering Center in
Cincinnati. These studies were directed toward the detection, enumeration,
and production of viruses in the laboratory, the determination of their
effects, and their control or removal by sewage treatment. A large number
of papers appeared in the 1960's describing the results of this research on
viruses in water supplies.
Studies of the toxicity of heavy metals in domestic water supplies have
been in progress for a number of years in several laboratories. This
activity was expanded because of the increase in metals and the need for
more definite data for the setting of drinking water standards. The explo-
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sive increase in the use of synthetic organic pesticides in the late 1940's
and 1950's led to research programs for their detection and measurement;
for the identification of their breakdown products; and for the determina-
tion of their accumulation in water, soil, and the bodies of organisms, and
their metabolic pathways and passage through the food chain. Extensive
studies have been and are being carried on to develop methods for
collecting these materials from water supplies and other portions of man's
environment and to determine their possible carcinogenic and other adverse
effects.
Following the passage by the U.S. Congress in 1948 of Public Law 845,
which was an important milestone in the struggle to abate pollution, the
Cincinnati laboratory of the U.S. Public Health Service was enlarged and
designated as the Environmental Health Center. Activities were reor-
ganized, and more emphasis was placed on research by the establishment of a
Research and Development Branch. An Aquatic Biology Section was set up
under my direction. The research effort was divided between two projects:
the biology of water supply and the determination of water quality criteria
for aquatic life.
The water-supply unit, under the leadership of C.M. Palmer, directed
its studies to the identification and control of organisms producing tastes
and odors, to the identification and removal of substances producing tastes
and odors, and to the identification of filter-clogging organisms and their
control. The usual method for the control of tastes and odors in water
supplies was to treat with chlorine or absorb the offending sub-
stances on activated carbon. The chlorine treatment was entirely experi-
mental and often was ineffective or resulted in the production of even more
odoriferous materials. The activated carbon treatment was usually success-
ful, but it often required tremendous amounts of carbon, which were costly
and presented a disposal problem. Something more exact than the cure-all
chlorine treatment was needed.
The first step in meeting the problem was to grow pure cultures of
those algae suspected of producing taste- and odor-causing substances to
determine which species actually produced such substances. The next step
was to collect and isolate those materials and determine their chemical
composition. It was believed that, if the chemical compositions of these
materials were known, methods could be developed for their removal from or
destruction in water-treatment plants. Over 100 species of algae were
grown in pure culture, but this line of research was not further supported,
and equipment and staff necessary for making the chemical analyses were not
secured. However, some Z5 years later this same research for the determina-
tion of the composition of taste and odor materials produced by living or-
ganisms was included as a research need in the National Academy of Science,
National Academy of Engineering report entitled, "Research Needs in Water
Quality Criteria 19/2."
Research for the development of culture methods for actinomycetes and
their pure culture was also carried out for the same objective. Cultures
were grown and odoriferous materials were isolated, but research for their
identification was not accomplished.
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The volume of sewage and other wastes discharged into our waters has
increased with growth in population and the construction of sewage
systems. This enrichment plus detergent carriers, certain industrial
wastes, and runoff from heavily fertilized agricultural lands has produced
large algal populations and the eutrophication of many lakes and reser-
voirs. These growths cause serious problems for water-treatment-plant
operators because of the clogging of filters. In some localities at
certain times backwashing requires one-fourth of the time of operation.
This procedure greatly increases costs and reduces the volume of finished
water produced. Known and suspected filter-clogging algae were cultured
for screening tests in an effort to determine the species causing the
trouble and to find a better and more selective algicide than copper
sulphate.
Series of screening tests were made with new materials that were
rapidly appearing on the market in the 1950's. We wanted to find algicides
that were specific for the target species and nontoxic to the others. Al-
though several good algicides were found, specific materials were not found
before the research work was discontinued when biological research was
transferred to the new national water quality laboratories.
In conjunction with the algicidal studies, research was carried out for
the development of biological controls. We found that several algae pro-
duced materials that inhibited the growth of other algae. We also found
that several algae produced antibiotics. In the course of these studies a
virus that destroys some bluegreen algae was discovered. Studies of this
virus have continued at the Cincinnati laboratory.
POLLUTION SURVEYS AND STUDIES OF NATURAL PURIFICATION AND BIOLOGICAL
INDICATORS OF POLLUTION
The establishment of the American Fisheries Society in 1870 and the
beginning of trout culture and the creation of the U.S. Commission of Fish
and Fisheries in the early 1870's indicated a national awakening of
interest in our fisheries and their protection. It had been noted that
fishing was greatly reduced or eliminated in many streams receiving sewage
or industrial wastes, or both. Fishermen began to complain and to point
out the need for pollution abatement. As a result of these complaints,
studies to determine the effects of pollution were undertaken.
In the 1870's Stephen A. Forbes of the Illinois State Laboratory of
Natural History began investigations of the Illinois River, which later
established a firm base for the comparison of stream conditions before and
after pollution. The study of the Illinois River by the Illinois Natural
History Survey is a classical study of the effects of stream pollution,
natural purification, and biological indicators of pollution. As sewage
from the city of Chicago was added to the river through the Chicago
Drainage Canal, the pollution moved progressively down the river as the
city and the waste load grew. This provided an excellent opportunity to
observe and study the progressive chemical, physical, and biological
effects of increasing pollution. Changes in color, turbidity, dissolved
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oxygen, carbon dioxide, bottom materials, plant growths, and aquatic animal
populations were observed and recorded. The findings of these
exceptionally pertinent investigations have been presented in a large
number of publications appearing over a period of half a century. They
describe changes in the aquatic biota as the pollution moved downstream and
also the natural purification brought about by the aquatic biota found in
the different areas. These studies dealt in detail with the plankton,
bottom organisms, and fishes, and changes in their populations over the
years as the organic load increased and the zones of pollution moved pro-
gressively downstream. Kofoid reported on "The Plankton of the Illinois
River 1894 to 1899" and "Microorganisms in Reservoirs and Their Relation to
Esthetic Qualities." Forbes and Richardson published a paper on "Studies
on the Biology of the Upper Illinois River" in 1913 and in 1919 another
paper entitled "Some Recent Changes in Illinois River Biology." In 1921
Richardson published a paper on "Changes in the Bottom and Shore Fauna of
the Middle Illinois River and Its Connecting Lakes Since 1913-1915 as a
Result of the Increase Southward of Sewage Pollution." In 1925 he
published "Changes in the Small Bottom Fauna of Peoria Lake 1920 to 1922."
These and many other reports on the effects of pollution in the Illinois
River furnish valuable data on the qualitative and quantitative composition
of aquatic populations in the different pollutional or life zones, their
value for characterizing the extent and severity of pollution, and their
role in natural purification.
After the turn of the century, many pollution surveys were made by
state conservation or fish and game departments and state health depart-
ments. After the passage of the federal law of 1912, the U.S. Public
Health Service made a survey of the Potomac River in 1913. At the urging
of W.T. Sedgwick of the Massachusetts Institute of Technology, W.C. Purdy
entered the pollution field and served as the plankton expert for the
survey. He studied the biology of the river and its flats and pointed out
the great value of the tidal flats for the digestion and natural purifica-
tion of the organic wastes from the city of Washington, D.C. His findings
were presented in a paper entitled "Investigation of the Pollution and
Sanitary Condition of the Potomac Watershed."
In 1914 Purdy was transferred to the U.S. Public Health Service Stream
Pollution Investigation Laboratory in Cincinnati. There he worked with the
bacteriologist, C.T. Butterfield, and later with the chemist, C.C.
Ruchhoft, who joined the laboratory staff in 1918. These three men and
their small staff made many valuable advances in the field of water
pollution research and pollution abatement. They participated in the Ohio
River surveys of 1914-1918 and 1937-1941, the Illinois and Scioto River
surveys, and the Lake Michigan survey. The results of their studies were
reported in a series of papers under two main headings, "Experimental
Studies of Natural Purification in Polluted Waters" and "Studies of Sewage
Purification."
In his natural purification studies to supplement field work Purdy set
up a small artificial stream using one-fourth mile of eave trough. It was
built on the laboratory grounds on a slope to ensure the desired current.
Water and a sewage waste were fed in at the upper end. Pollutional or life
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zones similar to those in sewage-polluted streams developed. This
artificial stream was observed year round, and it supplied valuable data on
natural purification; the role of different organisms in the purification
process; seasonal changes in the pollution zones and the purification
process; and the populations characteristic of the different pollutional
zones.
During the period from 1914 until the Second World War, pollution sur-
veys were made of many streams throughout the country by state and federal
agencies. Several universities conducted related biological studies or co-
operated with the state surveys. Birge and Juday of the University of
Wisconsin made limnological studies in the lakes of the state. The New
York Stream Survey under the direction of Emaline Moore (1926-1939)
supplied valuable data on aquatic populations living in organically
enriched streams and lakes. This survey was planned and carried out so
that one or more river basins were surveyed each year. Special attention
was given to polluted waters, the cause of pollution, and its effect.
In 1917 Weston and Turner at the Sanitary Research Laboratory and the
Sewage Experiment Station of the Massachusetts Institute of Technology
published a paper entitled, "Studies on the Digestion of a Sewage Filter
Effluent by a Small and Otherwise Unpolluted Stream." This investigation
was noteworthy because they studied in a natural stream the development of
pollutional zones, the natural purification process, and the development of
the aquatic biota characteristic of each of the zones of pollution.
As data on the extent of pollution and its effects became know, an
ever-increasing demand developed from fishermen, sportsmen's clubs, civic
groups, and fish and game departments for strong federal laws to control
pollution. Early in the 1920's the Isaac Walton League initiated a
national program for pollution abatement. This campaign was more effective
than the former attempts. The passage in 1948 of Public Law 845 was due in
part to the efforts of this group.
The U.S. Bureau of Fisheries and its successor, the U.S. Fish and
Wildlife Service, conducted surveys in several areas. In 1927 a biological
survey of the upper Mississippi River with special reference to pollution
was carried out under the direction of A.H. Wiebe. Several very productive
surveys were made by M.M. Ellis, Chief of the Fish and Wildlife Service
field station at the University of Missouri. He approached the problem
from the viewpoint of a physiologist and made many important contributions
on the environmental requirements of aquatic organisms.
Ruth Patrick of the Philadelphia Academy of Natural Sciences made exten-
sive studies of the role of plankton, especially diatoms, as indicators of
stream health or pollution. In connection with these studies the
"diatometer" was developed for the sampling of certain elements of the
plankton population.
In 1949 the Biology Section of the Environmental Health Center in
Cincinnati initiated the Lytle Creek study. This stream was selected for
special study after an extensive search for a stream with one source of
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pollution from a sewage treatment plant and in which all the zones of pollu-
tion from recent pollution to clean water were present. The stream was sur-
veyed, and a map was prepared showing pools, riffles and runs, and
different bottom types, as well as stream widths and depths. A sampling
program was established, and sampling stations were selected. A broad
crested weir and gauging station were built, and a weather station was
established. A trailer laboratory was equipped and placed near the
Wilmington, Ohio, primary sewage treatment plant and was used as a field
headquarters for chemical analyses and biological studies. Periodic
sampling studies over 24-hr periods with samples taken at each sampling
station every hour were conducted over a 2-year period. At least one such
continuous sampling was carried out in each of the seasons for each year.
During these studies hourly samples were taken for the determination of 02,
C02, pH, temperature, acidity, alkalinity, and turbidity. Samples were
also taken for BOD and COD (chemical oxygen demand) and periodic
bacteriological determinations. At selected times analyses were made for
PO^, NH3, N03, and H2S. Hourly plankton samples were taken during the
24-hr sampling periods to determine seasonal and die! fluctuations in the
populations. Periodic samples of the benthic macro- and microinvertebrates
were taken throughout the year under the direction of A.R. Gaufin. Monthly
seinings for fish were made at all stations throughout the stream.
These studies and samplings provided data on the various pollutional
zones: their biological, physical, and chemical characteristics; and sea-
sonal and die! changes in their characteristics and extent. This extensive
and intensive study produced many valuable data and several new concepts.
It was concluded that the quantitative and qualitative makeup of the biota
was characteristic of the so-called zones of pollution and was indicative
of environmental conditions or pollution. The mere presence or absence of
any single species could not be considered as an indicator of pollution.
In a polluted stream 02, C02, and pH could vary widely over a 24-hr period
at the same station. Such variations were especially noticeable in the
upper recovery zone where there were large growths of algae. These data
indicated that the sag curve, developed by nonbiologists who ignored the
effects of algal growths, could be very misleading, especially in the
smaller streams, because the samples for its determination were usually
taken after noon. Other important findings were the seasonal shift in
zones of pollution and changes in their character, the extension of
Sphaerotilus growth downstream in winter, the failure of fishes to enter in
winter the septic zone of summer even though 02 was abundant and the
inapplicability of the K factor developed for large rivers like the Ohio
River to small streams such as Lytle Creek, where it was 1.8 instead of
0.1. Data resulting from the Lytle Creek studies were reported in some 15
publications.
At the termination of the Lytle Creek studies in 1953, the laboratory
bioassay studies of the Biology Section were increased. Because the water
supply at the Sanitary Engineering Center was unsatisfactory for bioassay
investigations and water for such studies had to be brought from the
Newtown Fish Hatchery to the sixth floor of the center in glass containers,
a search was made for a water supply where meaningful studies could be
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made. In November 1953 a cooperative laboratory was set up with the
Department of Fish and Game Management of Oregon State University at
Corvallis, Oregon. This laboratory was under the direction of Peter
Doudoroff for the Biology Section and R.E. Dimick and C.E. Warren for
Oregon State University.
Biological studies at the State of Ohio Fish Hatchery at Newtown were
expanded over the years, and a temporary field laboratory was constructed
in the late 1950's. Dilution water for toxicity bioassays was secured from
the hatchery spring, and some of the hatchery ponds were used for field
studies. Bioassay studies and the facilities at the Newtown field station
were expanded under the immediate direction of Donald Mount, who joined the
staff of the Biology Section in 1960. Eventually, the building was
enlarged and the hatchery was secured for the toxicity bioassay studies.
TREATMENT OF ORGANIC WASTES
As communities increased in size, the need for sewage treatment and the
number of sewage treatment plants increased. Investigations for the
improvement of sewage treatment were carried out by state health depart-
ments and other state agencies. Rutgers University was one of the leaders
in this endeavor. Valuable work was conducted by William Rudolphs and H.
Heukelekian. Shortly after the establishment of the U.S. Public Health
Service Stream Pollution Investigation Laboratory in Cincinnati, a series
of investigations was undertaken that has continued to the present time
under different names. The results of basic studies conducted during the
1910's, 1920's and 1930's were published under the general title "Studies
of Sewage Purification," by Butterfield, Purdy, and Ruchhoft and their
small staff. Butterfield, in cooperation with Purdy, demonstrated the role
of certain protozoa in keeping bacterial populations active and efficient
in the utilization and breakdown of organic materials. Butterfield
investigated the die-away of coliforms in polluted waters and pioneered in
isolating zoogleal bacteria from activated sludges. He also demonstrated
that activated sludges consisting of pure cultures of zoogleal bacteria
were capable of rapid and efficient removal of BOD from both synthetic and
natural sewage.
Purdy published papers on the bulking of activated sludges as observed
at the Tenafly, New Jersey, sewage treatment plant and the use of chlorine
for the correction of sludge bulking in the activated sludge process.
James Lackey, who worked at the Cincinnati Laboratory from the 1920's to
the 1940's, published a series of papers under the general heading "Biology
of Sewage Disposal." He also published numerous papers on the role of
protozoa in waste treatment and water purification.
Chemical studies of the sewage-treatment and natural purification pro-
cesses were conducted by Ruchhoft and his staff. They developed analytical
methods for the detection and determination of waste materials and for
tracing waste streams to their sources in connection with stream surveys.
Considerable time was devoted to the development and improvement of the BOD
and COD tests and stream-survey methods and tests. In more recent years
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activated carbon was used for the collection and concentration of trace
materials from stream water and drinking water supplies so they could be
identified and quantified. Carcinogenic substances were found to be pre-
sent in small quantities in some water supplies. The carbon-filter techni-
que for the removal and concentration of trace materials from water used
for domestic supplies has in recent years brought to light the presence of
undesirable substances in water supplies hundreds of miles from their point
of discharge. It has also resulted in more emphasis on the detection and
control of toxic and harmful materials in drinking water supplies. With
the development of better analytical equipment and techniques, many
problems are being detected and solved that 30 years ago were impossible to
solve because of the lack of equipment and methods to detect and analyze
materials occurring in very small quantities in our waters.
DEVELOPMENT, USE, AND STANDARDIZATION OF BIOASSAY METHODS
Bioassays with fish as test organisms have been used for some time to
determine the acute toxicity of materials and wastes to selected test
species. In 1885 McDonald reported on his studies of the toxic effects
upon young shad of wastes from the Page ammoniacal works. In 1902 Knight
reported on his bioassay studies, and Moore and Kellermann reported results
of their bioassays in 1903 and 1905. In 1905 Marsh reported on studies of
the toxicity of some industrial wastes to fish. In the same year Levy
reported to the State Water Committee of Virginia on the investigation of
the effects of trade wastes (sulphite waste liquor) on the waters of the
James River at Richmond. In 1907 Marsh reported on the lethal dose of
copper sulphate in waters of different quality. Clark and Adams reported
results of their bioassay studies in Massachusetts in 1912. Wells
conducted extensive bioassys, and in 1913 he reported on reactions and
resistance of fishes to different concentrations of C02 and 02 and in 1915
on reactions and resistance of fishes to salts in their natural environment.
The use of bioassays increased between 1910 and 1920. In 1914 Adrian
Thomas conducted bioassays to test the toxicity of road tar. He used one
trout fingerling in each 1500-ml container and exposed the test fish to two
concentrations, 66 and 13 ppm by volume, for 3-19 days. Water in the test
chambers was changed once a week or more often. Aeration was very heavy,
and it may have removed some of the volatile components. In 1916 Shelford
and Wells reported on the use of sunfish to determine the toxicity of gas
house wastes. These were short-term acute toxicity bioassays of only 1-hr
duration. An important observation was that fish do not avoid this waste,
but swim into it. In 1917 Shelford reported on his continuing studies of
the effects of gas house wastes on fish. In 1917 Powers described his
bioassay studies in which he used the goldfish (Carassius carassius) as the
test animal. He reported additional work in 1920 on the influence of
temperature and concentration on the toxicity of salts to fishes. In 1919
Thomas of the Department of Game and Inland Fisheries of Virginia presented
a paper before the American Fisheries Society on the effects of certain
oils, tars, and creosotes on brook trout.
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During the 1920's the use of bioassays became more widespread. The
work of David Belding is noteworthy, as he had a good understanding of the
factors influencing the results of bioassays and the toxicity of wastes and
materials to aquatic life. In the 1924 paper by Belding, Merrill, and
Kilson, "Fisheries Investigations in Massachusetts," differences in the
sensitivity of different species to the same toxicant are pointed out.
They found that brook trout were seven times more sensitive than carp and
28 times more sensitive than goldfish to H2S. The authors stated, "There
is a marked difference in closely allied species such as the Salmonids."
They indicated that a species most sensitive to one material may be the
most resistant species in the group to another toxicant. They also pointed
out that fish vary seasonally in their resistance to toxicants; that the
quality of the receiving water affects toxicity; and that size or weight of
fish per volume of test solution, flow of water, 02 concentration, and
temperature are very important factors that influence results. Belding
developed these points further in the paper he presented to the American
Fisheries Society in 1927 entitled "Toxicity Experiments With Fish in
Reference to Trade Waste Pollution." In this report he discussed factors
that may be responsible for reported variations in toxicity of materials
and wastes such as: Individual variation in resistance among members of
the same species; differences in the sensitivity of a species to different
toxicants; differences in the sensitivity of different species to the same
toxicant; effects of age and size; differences in the dilution water, its
02 concentration, temperature, or dissolved materials; and differences in
type of test vessel used. Although he used only 24-hr tests, he recognized
that longer exposures at lower concentrations would produce kills. In his
bioassays he tested the toxicity of 20 materials to brook and rainbow
trout, Chinook salmon, carp, goldfish, and suckers.
Reports are available from several other investigators who carried out
bioassays in the 1920's. In 1928 Nightingale and Loosanoff used early life
stages of the Chinook salmon to test the toxicity of waste sulphite
liquor. Cole, Dilling, and Healey also conducted bioassays during this
period. In 1924 Thomas published a paper on the absorption of metal salts
by fishes. Wiebe conducted toxicity and pollution studies for a number of
years and reported on exposure of young fish to varying concentrations of
arsenic in 1930 and to sudden changes in pH in 1931; he also reported on
effects of dissolved phosphorus and organic nitrogen in the waters of the
Mississippi River in 1931.
During the 1930's bioassays were increasingly used for the evaluation
of problems by state and federal investigators. Studies were made of the
toxicity of cyanides, phenols, gas house wastes, pulp and paper mill
wastes, oil and petroleum products, and metals. Extensive studies were
also made on 02, C02, temperature, and pH requirements. Many bioassay in-
vestigations were carried out by the states and the U.S. Bureau of
Fisheries, which later in the decade became the U.S. Fish and Wildlife Ser-
vice. Among the latter, the research of Ellis was outstanding. In 1931
Surber and Meehan reported on lethal concentrations of arsenic for certain
aquatic organisms. Galtsoff made valuable contributions to knowledge of
the effects of oil on marine organisms, especially its effects on
shellfish. In the late 1930's and early 1940's Tennessee Valley Authority
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personnel engaged in extensive field investigations of the effects of
malaria control—oil and paris green larviciding--on aquatic life. During
the 1940's bioassays were performed at some universities. Among these the
work of Anderson with Daphnia warrants special mention. Of work carried
out in state laboratories that of Burdick in New York is outstanding.
With the introduction of the synthetic organic pesticides in the
1940's, there was a nation-wide surge of investigations of the toxicity of
these materials to aquatic life. The U.S. Public Health Service and the
U.S. Fish and Wildlife Service took dominant roles in these studies. At
the Technical Development Division, Communicable Disease Center of the U.S.
Public Health Service at Savannah, Georgia, I carried out and directed
extensive studies of the effects of ground and airplane spraying of DDT for
mosquito control on aquatic life. Effects of weekly applications of DDT
and other insecticides to water areas at 0.1, 0.05, and 0.025 pound per
acre were studied. Plankton, surface and benthic invertebrates,
terrestrial insects (especially bees), fishes, amphibians, reptiles, birds,
and mammals were studied. Applications of the insecticides were made by
hand dusters and sprayers and by airplane dusts, sprays, and thermal
aerosals. Results of this 3-year study were summarized in a series of
papers in the Public Health Reports of the U.S. Public Health Service.
As the number of investigators performing bioassays increased, many
different procedures, test organisms, dilution waters, and materials were
used. This diversity resulted in great difficulty in the comparison and
evaluation of the results reported by different investigators. Some
uniformity in testing procedures and in the reporting of results was
needed. In 1945 Hart, Doudoroff, and Greenbank published a book entitled
"The Evaluation of the Toxicity of Industrial Wastes, Chemicals, and Other
Substances to Fresh Water Fishes." In it they suggested procedures for
care and handling of test animals, preparation of dilution water, bioassay
procedures, and uniform methods for the reporting of results so that
results of different investigators could be compared and the tests could be
repeated. In 1949 Doudoroff, who was then on the staff of the Biology
Section of the Environmental Health Center at Cincinnati, invited prominent
workers in bioassay investigations to join him as members of a committee to
study the various bioassay procedures being used and to select or devise
and recommend procedures for bioassays which they considered best for
short-term toxicity tests with fishes. Members of this committee were: P.
Doudoroff, Chairman; B.G. Anderson; G.E. Burdick; P.S. Galtsoff; W.B. Hart;
R. Patrick; E.R. Strong; E.W. Surber; and W.M. Van Horn. The committee met
several times in Cincinnati and once in Woods Hole to draw up their
recommendations. These were published in 1951 under the title "Bioassay
Methods for the Evaluation of Acute Toxicity of Industrial Wastes to
Fish." This publication and the 1945 book by Hart, Doudoroff, and
Greenbank served as guides to those conducting bioassay studies and led to
more uniformity in the methods used.
Doudoroff and his associate, Max Katz, published a succession of papers
on bioassay studies and pertinent literature reviews while with the Biology
Section during the 1950's and early 1960's. These are listed in the list
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of publications of the Environmental Health Center and its successor, the
Robert A. Taft Sanitary Engineering Center.
During this period the U.S. Fish and Wildlife Service established a
special pesticide bioassay laboratory at Columbia, Missouri, the purpose of
which was to evaluate the toxicity of the new synthetic organic pesticides
to aquatic life. Another laboratory was set up at La Crosse, Wisconsin,
for the purpose of discovering materials or chemicals that were specific
for the control of undesirable aquatic species and that would act without
harm to desired organisms. This laboratory has a field station at Warm
Springs, Georgia.
After 1950 the growth in the use of bioassays was so rapid and so many
new workers entered the field in both the freshwater and marine environ-
ments that it is impossible to deal with all the developments and findings
in a review of limited size. It is proposed, therefore, to limit the
coverage of research activities after 1950 to those developments that in my
opinion have been most important in leading to the present pollution
abatement program of the U.S. Environmental Protection Agency.
DETERMINATION OF WATER QUALITY REQUIREMENTS FOR AQUATIC LIFE AND
DEVELOPMENT OF WATER QUALITY STANDARDS
In 1950 the pollution abatement program was not progressing as
expected. It appeared to me that, while some improvements had been made,
overall the situation was worsening. Several approaches had been tried,
but apparently a different approach was needed. Pollution abatement cases
in court were drawn-out and were often lost in long arguments over what
concentrations of wastes were really harmful and what really constituted
pollution. Local people and the courts were influenced by threats of
industry to move to another state. Some companies hired consultants to run
short-term bioassays to indicate that the concentrations of their waste in
the receiving water was not lethal. Hardship cases were pleaded on the
grounds that industries that were forced to treat their wastes, while
industries in other states were not, would be at an economic disadvantage.
Further, although chemical analyses had been made and the materials in
wastes identified, no firm data were available to indicate the maximum
concentration of waste that was not harmful under long-term or continuous
exposure. Courts were often not in sympathy with what they considered
drastic action in view of the supporting data, and they and many people
locally affected concluded that the only choice was fish or jobs, as
suggested by industrial and chambers-of-commerce spokesmen. In such a
situation they decided to take the jobs and let the environment take care
of itself. Suggestions had been made that government should tell
industries how to treat their wastes. Lack of such information was used by
some industries as an excuse for inaction, as no one had told them how to
treat their wastes at a profit. After reviewing the situation, the
ever-increasing number of new wastes and materials, and the present state
of knowledge as to what constituted pollution, I reached the conclusion
that the best way to attain pollution abatement was to set water quality
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standards for each water use based on a thorough knowledge of the water
quality requirements for each use. I reached this conclusion because:
1) Such standards would be uniform over large areas, everyone
would be required to meet them, and no economic advantage
could be acquired by anyone through exemption from treat-
ment;
2) requirements would apply to all sections of the country,
and thus there would be no incentive to move to escape
them;
3) standards would be based on carefully determined require-
ments, and no one would be required to treat more than the
essential amount; and
4) the standard, based on scientifically determined require-
ments, would provide a firm base for legal actions to
abate pollution.
Since standards should be based on water quality requirements, the
first task in a pollution abatement program is to determine water quality
requirements. Because a water that is favorable for aquatic life is
suitable for all other uses with recourse to available treatment methods,
with the possible exception of NOs in drinking water and bathing waters,
discussions of research in this review have been confined to those dealing
with the requirements for aquatic life and water supply. Essential
research that is to be considered is, therefore, that which is directed
toward the determination of water quality requirements for aquatic life.
Because the determination of water quality requirements for aquatic
life is largely a research problem in environmental requirements, ecology,
and toxicity, a well-trained, effective, and motivated scientific staff is
required along with money for the program, facilities, and equipment
essential for the research. Because many of the biologists working in the
U.S. Public Health Service regions felt isolated, a conference for all
aquatic biologists in the Service associated with any phase of water pollu-
tion research and investigations was held in Washington, D.C. in the fall
of 1950. This conference raised morale, fostered cooperation, promoted the
exchange of ideas and data, and improved the research effort.
Steps were taken to acquaint the leading conservation organizations
with the use and value of water quality standards in a pollution abatement
program. Groups contacted were the Sport Fishing Institute, the Isaac
Walton League, the National Wildlife Federation, the Wildlife Management
Institute, the Audubon Society, and the National Fisheries Institute.
Contacts with these groups were continued through the 1950's. I discussed
the need and value of water quality standards in six papers published
between 1957 and 1962.
In 1934 an annual literature review was begun by the Sewage Works
Journal, now the Journal of the Water Pollution Control Federation. Over
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the yaars its coverage was widened by the inclusion of papers in the fields
of stream surveys, chemistry, analytical methods, etc. A new section on
the biology of water supply and water pollution was included in the review
of the literature of 1953. A larger section was submitted for inclusion in
the 1954 review. This dealt with bioassays, studies of the toxicity of
chemicals and wastes to aquatic life, and biological indicators of pollu-
tion. The coverage was greatly expanded in the following years, and every
effort was made to supply summaries of papers dealing with environmental
requirements, the toxicity of wastes and other materials to aquatic life,
and water quality criteria and standards.
To promote further the objectives of the meeting held in Washington,
D.C., in the fall of 1950, the First Seminar on Biological Problems in
Water Pollution was held in Cincinnati in April 1956. This meeting was
attended by representatives from industrial concerns, academic institu-
tions, state conservation and health departments, and federal agencies.
Twenty-eight states and four provinces of Canada were represented. Biolog-
ical indicators of pollution, water quality criteria, and the use and value
of bioassays were discussed.
Because courts in some states were reluctant to accept as evidence the
results of bioassay tests in pollution cases, it was deemed advisable to
include a description of proposed bioassay methods in "Standard Methods for
the Examination of Water and Wastewater" by the American Public Health
Association ejt al_. Proposed standard bioassay methods, prepared by a
committee under my chairmanship and based largely on the 1951 report of the
Doudoroff committee, were included in the llth edition of this work which
was published in 1960. Their inclusion was instrumental in promoting more
uniform procedures, better and more comparable data, and greater use of
bioassays as a research and monitoring tool for the abatement of pollution.
From 1955 through 1966 research for the determination of water quality
requirements for aquatic life, the improvement of bioassay methods, and the
determination of the toxicity of pesticides was promoted to the fullest
extent possible by the Biology Section of the Cincinnati laboratory. The
research findings of the section during this period were described in 102
publications.
The Second Seminar on Biological Problems in Water Pollution was held
at Cincinnati in 1959. Attendance was much larger at this meeting than at
the first seminar. The seminar theme was the effects of pesticides on
aquatic life and allowable concentrations of various pesticides in the
aquatic environment. Other subjects discussed were the effects of the dis-
charge of radioactive materials, environmental requirements of aquatic
life, marine and estuarine problems, and the practical application of
biological findings in pollution abatement.
Contact was maintained with the private national conservation agencies,
and the leaders or staff members of a number of them attended the second
seminar. An advisory committee on water quality standards for aquatic life
made up of the leaders of these groups was established in 1960. Members of
this committee were Ira Gabrielson, director of the Wildlife Foundation
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Institute; Clarence C. Cottam, director of the Welder Wildlife Foundation
(formerly director and assistant director of the U.S. Fish and Wildlife
Service); Richard Stroud, executive vice president of the Sport Fishing
Institute; Thomas Kimball, executive director of the National Wildlife
Federation; Joseph Penfold, conservation director of the Isaac Walton
League; Charles Jackson of the National Fisheries Institute; and Charles H.
Callison, executive vice president of the Audubon Society.
These conservation organizations presented testimony before congres-
sional committees on various issues at frequent intervals. Some of their
testimony, especially that of Richard Stroud of the Sport Fishing
Institute, presented the need for and the value of national water quality
laboratories, one for fresh waters and one for marine waters, to carry out
research to determine water quality requirements for aquatic life. In
April 1962 the House of Representatives passed legislation authorizing two
water quality laboratories and appropriating money for their construction.
The Senate passed a similar bill in June. The conference committees came
to an agreement in August, and the bill authorizing the laboratories was
signed into law on August 12, 1962.
Planning for the Third Seminar on Biological Problems in Water
Pollution was under way for more than 2 years. The theme of this meeting
was water quality requirements for aquatic life. Every possible effort was
made to secure leading investigators to present papers and to assemble the
best possible program dealing with the chosen theme. The objective was to
produce a handbook summarizing available data on water quality requirements
for aquatic life. Representatives of 26 nations were in attendance.
Leaders of the national conservation groups took a prominent part in the
seminar, which was held August 13-17, 1962, just after the passage of the
legislation providing for the construction of the two water quality
laboratories.
Planning for the water quality laboratories was largely completed in
June 1963. Planning for the research program had been under way even
before the laboratories were authorized. The first staff member for the
National Water Quality Laboratory at Duluth, Minnesota, was housed in a
fish hatchery of the Minnesota Department of Conservation on the shore of
Lake Superior just northeast of Duluth in September 1964. The following
year several thousand square feet of space was made available by the
University of Minnesota at Duluth. The staff was enlarged and research
activities began. Initial activities of the National Marine Water Quality
Laboratory began on July 1, 1965, in office space provided by the
University of Rhode Island at Kingston. A search was made for laboratory
space on the coast, which could be used for research activities before the
construction of the new laboratory. Since none was available, the labora-
tory was set up in a former industrial laboratory at West Kingston, about 8
miles from Narragansett Bay. The assembled staff moved into this building
in September 1966. Laboratory furniture, equipment, and supplies, and a
laboratory staff were secured and assembled for both of the water quality
laboratories, and the research program for the determination of water
quality requirements for aquatic life was initiated under my direction.
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The bioassay laboratory that had been constructed on the grounds of the
Ohio Department of Conservation hatchery at Newtown, Ohio, was made a field
station of the Duluth laboratory. Construction of the National Marine
Water Quality Laboratory at Narragansett, Rhode Island, was delayed, but
construction of the National Water Quality Laboratory at Duluth proceeded,
and it was completed in the summer of 1967 and dedicated on August 12,
1967. Construction of the National Marine Water Quality Laboratory was not
initiated until August 12, 1975, 13 years after authorization.
The Federal Water Pollution Control Act, Public Law 84-660, as amended
by P.L. 87-88, P.L. 89-234, and by the Clean Water Restoration Act of 1966
(P.L. 89-753), required the states to establish water quality standards for
interstate waters by June 30, 1967. In case a state did not do this and
failed to call a pbulic hearing, the Secretary of the Interior was
authorized to set water quality standards for the interstate waters of that
state. On February 27, 1967, the Secretary appointed an advisory committee
to recommend water quality criteria for the following uses: Aesthetics and
recreation; public water supply; fish, other aquatic life, and wildlife;
and agricultural and industrial water supplies.
The committee on fish, other aquatic life, and wildlife was composed of
28 members of varied training and experience who collectively covered all
phases of the subject and represented a great deal of experience in
bioassay studies and water quality requirements for aquatic life. Their
task was to review available data on the water quality requirements of
aquatic life and then, on the basis of available data, their experience,
and judgement, to recommend water quality criteria. Their report was
completed by mid-June 1967. Their report on research needs was completed
in the spring of 1968, and both reports were published in April 1968 along
with the reports of the other committees. This report was updated and
expanded by a large committee of the National Academy of Sciences and the
National Academy of Engineering and was published in 1974 under the title
"Water Quality Criteria 1972."
The compilation of data for the 1968 report demonstrated that practi-
cally all the bioassay studies were of short duration and indicated only
the acute effects of toxicants on fishes. Methods had been suggested for
the use of application factors with data from acute toxicity studies to
predict long-term effects of toxicants, but few data were available to
indicate the maximum concentration of a toxicant in the aquatic environment
that was not harmful with continuous exposure. Studies of physical environ-
mental requirements, especially temperature and 02, has received the most
attention, and several field studies that extended over longer periods had
been made. Oxygen and temperature requirements of fishes were investigated
by a number of workers in the 1920's and 1930's. However, most of the
investigators reported on temperature and oxygen levels that were lethal,
and very few dealt with conditions that were favorable for the survival of
the species or that enabled them to compete successfully with their
competitors and predators. In the late 1920's Bel ding gave a good analysis
of the problem. There was no overall planning or coordination of the
investigations of environmental requirements, which were carried on by
investigators of diverse training, experience, and interests who were
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scattered throughout the country. The best work on environmental require-
ments in this period was that of M.M. Ellis and his staff. His paper on
"Detection and Measurement of Stream Pollution" has become a classic. His
recommendation of a minimum of 5 mg/liter of 02 for a well-rounded fish
population is still being used. It is good because it is based on field
studies in a large number of streams. Mention should be made of the
research carried on during the 1940's and 1950's at the University of
Toronto and its field station by F.E.J. Fry. His work and that of his
students had a great deal of influence on research on environmental
requirements. His leadership and foresight made his laboratory a world
leader in temperature and oxygen-requirement studies. In the mid-1950's a
long-term study of the oxygen requirements of fishes was initiated at the
cooperative water pollution laboratory of the Biology Section of the Robert
A. Taft Sanitary Engineering Center and Oregon State University under the
direction of Doudoroff and Warren.
The passage of the Water Quality Act of 1965 and the Clean Water Act of
1966 requiring the states to establish water quality standards stimulated
country-wide research on water quality requirements. This legislation and
increased public awareness of the ever-increasing detrimental effects of
water pollution due to population increases, industrial development, and
new highly toxic products caused an increase of research efforts in
freshwater, estuarine, and marine environments. The growth in size and
number of thermal electric generating plants and the construction of
nuclear plants caused a phenomenal increase in studies on the effects of
temperature on the aquatic environment. The great increase in sea
transport of oil and the Torrey Canyon spill brought a similar increase in
studies of the effects of petroleum products on the aquatic environment.
In addition, for about 30 years the toxicity of pesticides {insecticides,
herbicides, algicides, fungicides, etc.) and their effects on aquatic and
terrestrial non-target species including man had been an ongoing problem.
Water pollution hearings and enforcement actions requiring hard evidence
brought the water quality researchers to the front lines for the presenta-
tion of data, the collection of evidence, and the recommendation of
criteria and standards.
The increased and broadened research due to the above factors produced
a tremendous increase in the use of bioassays. With the expansion of the
investigations into the marine environment, there was a great increase in
the use of different groups and species as test organisms, creating a need
for additional bioassay methods. In 1966 the standard bioassay methods
committee began to prepare materials for the 13th edition of "Standard
Methods for the Examination of Water and Wastewater," Subcommittees were
set up in each of the water quality laboratories and prominent investiga-
tors in other federal agencies and the states were invited to serve on the
committee. Although I, as chairman, wished to include new methods for
marine organisms, the committee felt the tests were not yet well developed,
and the only new material in the 13th edition, printed in 1971, dealt in
the long-term tests for fishes.
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Although we have many marine laboratories in the United States, only a
few conducted research on water pollution problems until fairly recently.
Some of the early studies were carried out by Julius Nelson, who began his
investigations of oysters in the last quarter of the past century. His
son, Thurlow Nelson, continued the studies on oysters and presented his
findings in several reports from the Department of Biology of Rutgers
University (1917, 1921, 1923, 1927, 1936, and 1938). In the 1950's
bacteriological studies were made on sewage disposed of through ocean
outfalls by the determination of coliforms and other bacteria, their
abundance and rate of decrease, and the factors affecting each. Studies
were also made to evaluate bacterial contamination, bacteria in sediments,
bacteria in submerged outfalls, the bactericidal action of sea water, and
the survival of enteric organisms in sea water-
During the past 20 years biological oceanography and marine biology
have recieved more attention. Stations are now established at the mouths
of estuaries, and the emphasis has changed from observational, physical,
chemical, and taxonomic studies to biological, physiological, environ-
mental, and ecological approaches including bioassays, toxicological
studies, environmental requirements, and mariculture. Mariculture, or the
rearing of food organisms, is important especially in that it provides
techniques for the rearing, care, and handling of test organisms for
bioassays and the production of different life stages of organisms in
sufficient numbers for bioassay tests. The development of methods for
rearing mrine organisms is basic for long-term bioassays extending through
a portion or the entire life cycle of an organism.
From November 1972 through April 1974, I devoted a portion of my time
to a survey of the marine laboratories of the Atlantic and Gulf coasts.
This provided first-hand information on the various research projects under
way on the toxicity of pesticides, metals, and other materials; on the en-
vironmental requirements of estuarine organisms; and on the culture not
only of fishes, but also of many other marine organisms important
commercially and as food for other organisms. Some of the subjects that
are now being or have recently been studied are: (1) Microbial
decomposition of oil and pulp mill wastes; (2) bioaccumulation of heavy
metals by littoral and pelagic marine animals; (3) effects of toxicants on
the larval stages, juveniles, and adults of marine animals to ascertain the
most sensitive stages in the life cycle; (4) potential environmental
disturbances due to marine mining operations as a basis for developing
appropriate marine mining techniques; (5) decay of pesticides in marine
sediments, its rates and pathways, and identification of decomposition
products and their effects; (6) the distribution of radionuclides in the
marine environment; and (7) accumulation of persistent organic compounds in
phytoplankton and their effects. In addition, studies are being made of
the effects of ocean dumping of wastes; the collection of potential
toxicants in bottom deposits; the return of toxicants from benthic
sediments; the increase of dissolved toxicants in marine water; the
bioaccumulation of toxicants; the fate and effects of oil, pesticides, and
metals in the marine environment; and the development of bioassay methods
for use with marine invertebrates.
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Planning for the bioassay section of the 14th edition of "Standard
Methods for the Examination of Water and Wastewater" began in 1971. To
secure the broad coverage needed to meet the problems facing investiga-
tors and furnish methods for bioassays with the broad spectrum of organisms
being used or which should be used in the toxicological and water quality
studies, subcommittee chairmen were appointed for each of the main groups
of organisms commonly used in bioassays--i.e., phytoplankton, zooplankton
(protozoa, copepods, and Daphnia), corals, worms, crustaceans, aquatic
insects, mollusks, and fishes. Although several of the suggested methods
had not been extensively tested, it was considered imperative to make a
beginning so much-needed methods for these groups could be used and
improved. The 14th edition of this work was published in 1975. It is
hoped that these suggested methods will provide the base upon which
adequate bioassay methods for all the different groups of organisms can be
developed to meet the needs for the determination of water quality
requirements, the detection and evaluation of pollution, and the supplying
of data necessary for setting effluent requirements, granting discharge
permits, and enforcement actions.
The passage of Public Law 92-500 in 1972 added another great need and
impetus for research to supply the required information for pollution
abatement. Research for the determination of water quality requirements on
which water quality standards must be based is now developing and expanding
rapidly, and more of it is now in the right direction. I have had the
privilege of knowing the men who pioneered water pollution research from
1910 to 1930 and discussing problems with them. These contacts and my
field and laboratory investigations in aquatic biology, ecology, environ-
mental improvement, fisheries management, pollution abatement, and water
quality requirements since 1928 have been a wonderful experience. Although
much remains to be done and there have been mistakes and defeats, good
progress is now being made. With qualified and motivated leaders, well
trained and experienced in the work they are supervising, and competent
dedicated workers, pollution can be abated.
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SECTION 3
CHARACTERISTICS OF THE MOSCOW RIVER WATER QUALITY
ACCORDING TO HYDROBIOLOGICAL INDICES
V.A. Abakumov and G.L. Margolina
The Moscow River water quality is characterized by a combination of hy-
drobiological indices including indices of microbiological activity', pri-
mary production, saprobility, biotic potential, species diversity, and toxi-
cological indices.
In order to obtain hydrobiological indices characterizing the water
quality of the Moscow River, the qualitative and quantitative composition
of bacterioplankton, phytoplankton, zooplankton, zoobenthos, and higher
water plants were studied in a region of the Moscow River extending from
Zvenigorod to Kolomna at the following five locations:
1. above Zvenigorod;
2. below Moscow;
3. at 107 km from the Moscow River mouth;
4. at 85 km from the Moscow River mouth;
5. in the Kolomna region (at 7 km from the mouth).
On the basis of the investigations carried out, the Moscow River water
above Zvenigorod is estimated as pure with some indications of a slight pol-
lution according to microbiological indices. The highest indices of group
and species variety of plankton and benthos organisms with rapid develop-
ment of oligosaprobic organisms and high biotic potential were noted here.
In this region, such oligosaprobic species of phytoplankton as Fragill aria
virescens, f_. capucina, Coelastrum microporum, and Gomphosphaeria lacustris
are abundant. Along with them, one should distinguish Melosira granulata,
Scenedesmus and Chlorella. This region is characterized by the richness of
the species composition of high water plants, number of individuals, and
high protective covering. This region near the bank is characterized by an
emergent plant community and a community of plants with floating leaves.
Here there is an abundance of Scirpus lacustris, Glyceria maxkma, Nuphar
Pathogenic microorganisms are not considered here.
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lutea, Hydrocharis morsus-ranae, Polygonum amfibium, Myriophyllum
verticillatum, Ceratophyllum submersum, Potamogeton lucens, and Potamogeton
perfoliatus. These qualitative characteristics are supported by the follow-
ing quantitative indices: Number of plankton alga species - 33; number of
higher water plant species - 22; number of groups of benthic organisms - 7;
number of phytoplankton organisms - 369 thousand cells/liter; number of
zooplankton organisms - 7 thousand specimen/m3; number of Oligochaeta -
2280 specimen/m2; relative number of Oligochaeta (per cent of the total
number of zoobenthos organisms) - 21%; mean protective covering of higher
water plants - 90%; biotic potential - 6 points; total number of bacteria -
3 million/ml; number of saprophitic bacteria - 3400/ml; number of spore-
forming bacteria in 1 ml - 10; ratio of the total number of bacteria to the
number of saprophitic bacteria - 900; and mg/liter of 02 demand for a 24-hr
period - 0.32. The toxicology investigations completed showed results
which did not differ from the control.
Among the investigated sites, the greatest change in water quality is
observed directly below Moscow. In this region of the river, the negative
effect of water quality on all components of the ecological system is ob-
served; this effect is displayed in the decrease of species variety, in the
predominance of a-mesosaprobic organisms, low biotic potential, low P/R
ratios, and in the increase in the number of saprophitic bacteria. Higher
water plants are numerous here and Potamogeton pectinatus is the predomi-
nant species among them. Zoobenthos is composed of Oligochaeta, chirono-
mids, and leeches, and chironomids are represented by one species only.
Among plankton algae, Melosira granulata, Scenedesmus quadricauda, etc.
were encountered. Toxicological investigations allow us to ascertain
slight water toxicity. The following quantitative indices may illustrate
the above mentioned qualitative features: Number of plankton alga species -
10; number of species of higher water plants - 4; number of groups of zoo-
benthos organisms - 3; number of phytoplankton organisms - 178 thousand
cells/liter; number of zooplankton organisms - 31 thousand specimen/m3;
number of Oligochaeta - 56,000 specimen/m2; relative number of Oligochaeta -
93%; mean protective covering of higher water plants - 20%; biotic potential
- 2 points; total quantity of bacteria - 3.7 million/ml; number of sapro-
phitic bacteria - 30 thousand/ml; number of spore-forming bacteria in 1 ml
- 200; ratio of the total number of bacteria to the number of saprophitic
bacteria - 120; and mg/liter of 02 demand for a 24-hr period - 0.96.
The pollution level produced by such a large city as Moscow proved to
be so small due to the high effectiveness of treatment installations, so
that even a small river, such as the Moscow River, copes with pollution
rapidly. The rapid increase in water quality at sites mentioned below
verifies this fact. At the site 107 km from the mouth of the Moscow River,
significant indications of an increase in water quality in comparison with
a previous site are observed. This is also verified by the appearance of
g-mesosaprobic indicator organisms, an increase in both the number of or-
ganisms and their species variety. Coelastrum microporum and Closteium
lunula appeared among the phytoplankton. This site is characterized by the
following quantitative indices: Primary production - 0.52 mg 02/liter for
a 24-hr period; demand - 1.71 mg 02/liter for a 24-hr period; P/R ratio -
33
-------
0.3; number of plankton alga species - 13; number of higher water plant
species - 7; number of groups of zoobenthos organisms - 3; number of phyto-
plankton organisms - 2,170 thousand cells/liter; number of zooplankton or-
ganisms - 35 thousand specimen/m3; number of Oligochaeta - 33,780 specimen/
m2; mean protective covering of higher water plants - 50%; biotic potential
- 2 points; total quantity of bacteria - 3.6 million/ml; number of spore-
forming bacteria in 1 ml - 200.
In the region 85 km from the mouth of the Moscow River, the water qual-
ity has considerably increased due to the development of self-purification
processes. Production processes are increased with the relative decrease
in destruction processes. The number of plankton organisms increases.
Species variety is increased. The greatest number of species of algae and
higher water plants in the lower Moscow River fow are located here. Benthic
organisms are represented by five systematic animal groups, and the number
of chironomid species increased to four. The indication according to sapro-
bility of plankton forms demonstrated that they are completely devoid of
polysaprobic species and that the a-mesosaprobic species approaches the
minimum. Bacterial pollution is also considerably decreased. Thus, the
water of the Moscow River in the 85 km region may be characterized on the
whole as slightly polluted. The following data do not contradict this con-
clusion: Primary gross production - 2.56 mg Oa/liter for a 24-hr period;
demand - 1.41 mg 02/liter for a 24-hr period; P/R ratio - 1.8; number of
plankton alga species - 20; number of higher macrophytes - 12; number of
phytoplankton organisms - 9,300 thousand cells/liter; number of zooplankton
organisms - 20,000 thousand specimen/m3; number of Oligochaeta - 16,130
specimen/m2; relative number of Oligochaeta - 80%; mean protective covering
of higher macrophytes - 70%; biotic potential - 4 points; total number of
bacteria - 2.5 million mg/liter; number of saprophitic bacteria - 10 thou-
sand/ml; the ratio of the total number of bacteria to the number of sapro-
phitic bacteria - 250.
In the Kolomna region, due to the fact that self-purification processes
are developed here to the greatest extent, a further increase in water qual-
ity is taking place and indices of oligosaprobility appear. The number of
organisms increased: The number of phytoplankton organisms approaches 13
million cells/liter, and zooplankton - 206 thousand specimen/m3. The
species variety of zooplankton is sharply increased mainly due to Rotifera.
Production processes reach their highest intensity - 6.12 mg 02/liter for a
24-hr period. According to all the indices of benthic organisms develop-
ment, the river is also characterized by a very slight pollution; the number
of zoobenthos groups - 6; the number of Oligochaeta - 1920 specimen/m2
(lower than in the vicinity of Zvenigorod); relative number of Oligochaeta
- 11%, etc.
However, the water quality in the vicinity of Kolomna does not reach
the level which is observed in the vicinity of Zvenigorod. Actually,
according to a number of indices, higher water quality near Kolomna is not
expressed so clearly: The decrease in bacterial pollution is not so rapid;
destruction processes are increased; in comparison with the 85 km outlet,
the decrease in species variety for plant organisms (phytoplankton and
34
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macrophytes) is observed. Among plankton algae, for example, one species
of diatoms - Melosira granulata - dominates. It comprises 80% of the total
amount of algae. Macrophytes are represented by a community of water-marsh
plants, which verifies the fact that the river mouth is silting in. Thus,
the water in the mouth of the Moscow River in the vicinity of Kolomna
according to hydrobiological indices is characterized mainly as slightly
polluted.
In conclusion, it should be noted that the research carried out proves
the necessity to apply a combination of water quality control methods
according to hydrobiological indices which provide information not only on
water pollution, but also on the state of aquatic organisms in water bodies
which are valuable from the point of view of fisheries.
35
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SECTION 4
ENDPOINTS IN BIOASSAY
Lloyd L. Smith, Jr.
INTRODUCTION
Bioassay with fish or invertebrates has long been used to determine the
suitability of water for aquatic life and the toxicity or deleterious
effects of industrial and domestic sewage effluents on aquatic habitat.
Various techniques, exposure times, and definitive endpoints of tests to
describe effects have been employed. An endpoint in a bioassay is defined
for our discussion as a physiological or behavioral response to a specific
concentration of a toxicant after a definite period of exposure. In order
that data provided on a worldwide basis can be compared and similar conclu-
sions derived from similar values, it is essential that a clear understand-
ing of the usefulness of various endpoints or indicators of adverse effects
be developed.
The purpose of defining specific endpoints for bioassays is to secure
values in terms of milligrams per liter or degrees C which can be trans-
lated into an adequate assessment of the effects of a toxicant or effluent
on fish or other aquatic species. In order that the value designated for
such an endpoint have broad usefulness, it must first have a predictive
capability for either acute or long-term effects of a toxicant or effluent
on fish and invertebrate populations; second, it must permit comparisons of
effects between different species and between different toxicants or other
deleterious materials; third, it must be a practical endpoint to observe
without unnecessary laboratory sophistication or excessively time-consuming
analytical procedures; and fourth, it must be reported in a form that will
permit comparison between laboratory and field data. Finally, an endpoint
must be selected which is applicable to the particular problem. Unfortun-
ately, much bioassay information in the world literature has not been based
on careful attention to the points enumerated above. The purpose of this
discussion is to elaborate on means of defining meaningful endpoints and
interpreting the resultant findings.
TYPES OF ENDPOINTS
Four broad types of endpoints have been employed by various investiga-
tors at various times. They are: (1) endpoints indicating acute toxicity
and resulting in death of the test organism in short-term; (2) endpoints
36
-------
defined by reduced fecundity, growth, or changed behavior in long-term
tests at subacute concentrations; (3) endpoints defined by chemical changes
in the body or changed physiological rates; and (4) endpoints defined by
behavioral responses.
Acute Tests
Acute tests resulting in death are performed for short periods of time
ranging from several hours to 30 days. Interpretation of results has been
accomplished by calculation of a tolerance limit or concentration of the
test solutions which will cause death to some proportion of the test
organisms within a specified time. The median tolerance limit (LC50, TLm)
or concentration killing 50% of test organisms has been employed most fre-
quently. Two broad uses have been made of acute tests. One use has been
to identify potentially toxic materials or deleterious effluents through a
"screen test." This test is usually a static test of short duration, nor-
mally not exceeding 24 or 48 hr, which provides a gross estimate of
probable toxicity. A second form of acute test is designated to determine
a short-term response which may be used as a base to estimate effects of
long-term exposures or to provide a criterion for monitoring and enforce-
ment of water quality standards. This type of test has conventionally been
of 96-hr duration, especially in the United States. More recently, longer
time periods have been employed to secure a better base for predictions
that will accommodate the differing speeds with which toxicants may act.
The screen test is of short-duration and must permit calculation of a
median tolerance limit after 24-48 hr. The 96-hr test similarly depends on
calculation of median tolerance limits of concentration over this time
period, which presumably gives both a standard value and is long enough to
permit acute lethal effects to develop. Because many toxicants or
effluents do not effectively demonstrate their toxicity within the 96-hr
period, the asymptotic or time-independent test for acute response is being
used with increasing frequency. This test may run for up to 30 days and is
interpreted by the flattening of the toxicity curve. It is stopped when no
death occurs in test chambers for a period of 24-72 hr. This could be
described as a lethal threshold concentration (LTC) (Figure 1).
Long-Term Test
A second set of endpoints can be employed in tests with long-term expo-
sure of organisms to toxicants or effluents. The test concentrations will
be lower than the median tolerance concentrations (LC50), and in conse-
quence, the significant endpoints will usually be demonstrated by physio-
logical inhibitions such as reductions of fecundity, growth rate, or
ability to do work. Increased deaths in treatments over losses in the
control after long exposure may also be used as an endpoint. Interpreta-
tion of these long-term inhibitions at low concentrations cannot be arbi-
trarily assessed. Usually the toxicant level which does not result in
reduced growth rate, lessened fecundity, or lowered fertility is considered
to be a "safe level" and consequently an acceptable concentration of poten-
tially toxic materials. A satisfactory endpoint of a long-term test,
37
-------
00
a:
UJ
o
O
LETHAL THRESHOLD CONCENTRATION
(LTC)
456
TIME - DAYS
8
10
Figure 1. Toxicity curve demonstrating lethal threshold concentration (LTC).
-------
however, may not be developed until the successful reproduction and growth
rate of second-generation individuals has been shown to be unaffected.
Biochemical and Physiological Endpoints
A third group of endpoints has been explored by various investigators
and has shown promise. These endpoints are related to the changes in blood
characteristics and enzyme activity, changes in metabolic rates, or other
physiological changes. A deficiency of this type of endpoint in long-term
tests is that it usually requires sacrificing of the organisms or submit-
ting the organism to unnatural conditions which may change metabolic res-
ponses caused by restraint of movement, excitement, or lethargy. Further,
the ecological significance of these endpoints and the actual detrimental
effects on the organism are difficult to interpret. Change in these para-
meters does not, a priori, indicate a disadvantage to survival characteris-
tics or growth characteristics of the organisms. These endpoints have the
further disadvantage that they must be described and interpreted by
scientists. Tests of this kind cannot be delegated to field investigators
or relatively unskilled laboratory personnel who must be employed in large-
scale monitoring or surveillance programs. They have substantial value for
sophisticated research and identification and classification of potential
or real toxicants where results of these physiological criteria can be
equilibrated with results easily interpreted in terms of ecological conse-
quences that have practical value.
Behavioral Endpoints
A fourth type of endpoint may be described as a behavioral variation
which is considered to be inhibitory to growth and reproduction or long-
term survival. In various experiments it has been demonstrated that be-
havioral aberration will prevent spawning at toxicant levels where other
measurable parameters appear to be unaffected. Changes in degree of
mobility, increase or decrease of opercular ventilation, and avoidance
reactions have all been used to indicate adverse physical conditions for
the individual organism. As in the biochemical endpoints, interpretation
of the observed results in terms of ecological and field significance is
difficult unless comparative tests with other ecologically significant end-
points are made.
FACTORS AFFECTING VALIDITY OF ENDPOINT
After a significant endpoint for bioassay has been selected with care-
ful attention to the objective sought, a series of potential limitations of
the test procedure itself will determine the validity of the toxicant
values produced when the endpoint is reached.
The first consideration is the health of the test species. The test
fish must be in good physical health, either having been laboratory-reared
under controlled and disease-free conditions or captured from a known wild
stock where fish have not been stressed by pollutants or other physical
factors. Disease, starvation, or careless handling before tests will
39
-------
seriously affect results of acute or long-term tests, regardless of which
bioassay endpoint has been selected. Usually holding fish for an acclima-
tion period in the laboratory before testing will insure a reliable
response if fish remain disease free and accept food normally. It also may
be desirable to subject samples of the test fish to the effects of a
reference toxicant for which response has been well documented with the
species.
Another factor that is important in verifying the reliability of
endpoints is the degree of crowding of test organisms in the test chamber.
When small fish are used, 1-2 g of fish per liter of water in the test
chamber will usually allow sufficient space to permit free movement and pre-
vent secondary effects from too many test specimens. If large fish are
used, especially adults of some species, antogonism between individuals can
seriously affect final results by causing fish to reach the designated end-
point at lower toxicant concentrations than fish not stressed by behavioral
patterns. Test conditions which overstimulate the fish to activity or de-
press activity unnaturally will affect the validity of results at the
selected endpoint.
Three factors influencing the validity of a selected endpoint are
temperature, oxygen concentration, and pH. Temperature has a marked effect
on the sensitivity of test organisms and consequently on the calculated
LC50 or other endpoints selected. The effect of temperature cannot be pre-
dicted with certainty. For example, the tolerance of fathead minnows and
goldfish to hydrogen sulfide is greatly increased by a 15 C lowering of
temperature (Figure 2). In contrast, tolerance of bluegills is decreased
by a lowering of temperature (Figure 3). It is therefore important that
when an endpoint for a bioassay is selected, the test temperature is
related to the objective of the test. A standard test temperature of 25 C
does not necessarily relate to the ambient ecological condition in nature
where the test results are expected to be applied.
Oxygen concentrations below 4 or 5 mg/liter will increase the sensitiv-
ity of test species to most toxicants. At low dissolved 02 concentrations
(below 4 mg/liter) a new stress is added that increases the adverse
response of the organism to the toxicant. Similarly, extreme variations in
the hydrogen ion concentration (pH) of test water can alter response and
affect the validity of the endpoint chosen. This influence may be exer-
cised through the effect of pH on ionization of the material being tested
or on the physiological conditions imposed upon the fish which make absorp-
tion or blood changes more or less responsive to changes in concentration
of toxicant.
A factor frequently overlooked in choosing bioassay endpoints for
various fish species has been the difference in tolerance of eggs, larvae,
juveniles, and adult fish. Frequently fish in the early life-history
stages are much more sensitive than older fish and in consequence a satis-
factory endpoint for one life-history stage will not necessarily
demonstrate the sensitivity of the species through its entire life cycle.
Examples of differences can be drawn from H2S studies where frey or larvae
are the most sensitive form and may vary markedly from juveniles. In con-
40
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800
600
^400
CVJ
I
^ 200
9
8 100
CJ 80
60
i <°
20
B
i i i i i i
i i t
j i
6 8 10 B 20 30 40
TEMPERATURE C
Figure 2. Tolerance of (A) fathead minnows (Pimephales promelas) and (B)
goldfish (Carassius auratus) to hydrogen sirtfide at different
temperatures.
41
-------
140
120
100
80
o
b
60
40
20
10
15
20
25
TEMPERATURE (C)
Figure 3. Tolerance of bluegills (Legomis macrochirus) to molecular cyanide
(HCN) at various temperatures. (Dotted portion of curve is
extrapolated.)
42
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trast, hydrogen cyanide (HCN) studies show that juvenile fish may be 75-80%
more sensitive than the egg in some species. Further, the behavioral
inhibitions on spawnig of some species like bluegill and brook trout
(Salvelinus fontinalis) in respose to H2S may be apparent at levels far
below those which cause acute mortality in 96-hr or which reduce growth
over long periods. These factors make it important that an appropriate
life-history stage be selected for testing and that an endpoint be chosen
which reflects the true sensitivity of that stage to the toxicant.
Selection of test fish stocks from widely separated geographical areas
or different cultured stock may introduce wide variations in results. In
fathead minnows acute sensitivity to cyanide (HCN) and H2S may vary as much
as 30-40% between stocks. It is also important that the influence of test
conditions (temperature, pH, 02, fish numbers) be taken into consideration
when selecting an appropriate endpoint for bioassay and in evaluating the
results when they are obtained.
PREDICTION OF LONG-TERM ADVERSE EFFECTS FROM SHORT-TERM TESTS
Ideally a toxicant should be administered to a fish or an invertebrate
throughout its entire life history, beginning with the egg through repro-
duction and into the second generation, to determine concentrations which
will not adversely affect the population. However, the large number of
known toxicants and unknown mixed toxicants which must be tested will pre-
vent definition of safe levels for many materials by long-term tests. It
is therefore common practice to make an acute test (96 hr) defining some
median tolerance limit (LC50 or TLm) and then to apply a mathematical
factor which will reduce the value of this concentration to that considered
safe for completion of all life-history stages. This factor is usually
called an application factor and is calculated by dividing the safe concen-
tration by the 96-hr LC50 (TLm) of the toxicant. Historically, first
approximations of this factor were made by comparison of acute median
tolerance limit tests and long-term or chronic tests of a limited number of
toxicants. The tests were conducted on the same species and under the same
conditions. Results of acute tests were also compared to concentration of
toxicants in streams which contained normal fish populations. While many
materials appear to have application factors of similar magnitude, certain
families of materials have factors different by orders of magnitude from
established means and factors (Table 1). With a single toxicant the appli-
cation factor may vary substantially between species. An example can be
drawn from our studies of H2S in which the application factor varies widely
between species and between life-history stages of the same species (Table
2). Also fish species react very differently to materials in similar
chemical families (Table 3). These examples indicate that no single appli-
cation factor can be used to predict "no-effect" toxicant concentrations
from short-term tests. In consequence, uses of application factors with
short-term tests have tended to be more restrictive in their results than
necessary in some cases and much less restrictive than required in others.
43
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TABLE 1. COMPARISON OF APPLICATION FACTORS FOR VARIOUS COMPOUNDS
Species
96-hr LC50 MATCd Application
(yg/liter) (yg/liter) factor
Fathead minnow
Bluegill
(Mount and Stephan, 1967)
Malathion
(Organophosphorus pesticide)
9,000 200-580
89 3.6-7.1
Guthion
(Organophosphorus pesticide)
Fathead minnow
(Adelman, Smith, and
Siesennop, 1976)
Fathead minnow
(Smith, unpublished)
9,000
Cyanide
107
Cadmium
Fathead minnow 7,200
(Pickering and Gast, 1972)
Bluegill
(Eaton, 1974)
24,000
Fathead minnow
Soft water
(Mount and Stephan, 1969)
Hard water
(Mount, 1968)
Brook trout
Soft water
(McKim and Benoit, 1971)
Copper
84
430
100
37-57
31-80
0.02-0.06
0.04-0.09
0.33-0.55 0.00017-0.00027
5.3-17.3 0.049-0.162
0.005-0.008
0.0015-0.0039
10.6-18.4 0.13-0.22
14.5-33.0 0.03-0.08
9.5-17.4 0.10-0.17
aMaximum allowable toxicant concentration.
44
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TABLE 2. APPLICATION FACTORS FOR H2S WITH JUVENILES OF SIX FISH
SPECIES (EXPRESSED AS yg/LITER)
Species
Brook trout
Rainbow trout
Fathead minnow
Goldfish
Walleye
Bluegill
LTC
16
9
16
71
19
32
Chronic
no-effect
6
3
3
5
3
1
Application
factor
0.375
0.353
0.188
0.070
0.158
0.030
TABLE 3. 96-HR LC50 OF EIGHT FISH SPECIES TO PESTICIDES
(EXPRESSED AS yg/LITER)a
Species
Ictaluridae
Catfish
Cyprinidae
Goldfish
Carp
Centrarchidae
Sunfish
Black bass (largemouth)
Salmonidae
Rainbow trout
Coho salmon
DDT
16
21
10
5
7
4
Parathion
5,710
9,000
7,130
5,170
5,220
2,750
5,300
Guthion
3,290
4,270
695
52
5
14
17
Percidae
Yellow perch 9 3,060 13
aMacek and McAllister, 1970.
45
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SUMMARY
On the basis of the foregoing considerations, it is recommended that
(1) the most sensitive stage be used where possible as the basis for acute
median tolerance limit tests; (2) temperature used for tests should
approximate the natural conditions to which the fish will be subjected
during critical periods in the outdoor environment; (3) median tolerance
limit value should be based on the time at which an asymptote is reached in
the decline of the toxicity curve; (4) where possible, reproductive
behavior and success be used as the final criterion of the proper endpoint
used to determine safe concentration of toxicants or effluents; and (5)
uniform application factors not be used over a broad spectrum of species or
toxicants without definitive chronic tests.
REFERENCES
Adelman, I.R., L.L. Smith, Jr., and G.D. Siesennop. 1976. Chronic toxicity
of guthion to the fathead minnow (Pimephales promelas Refinesque).
Bull. Environ. Contam. Toxicol. 15: 726-733.
Eaton, J.G. 1974. Chronic cadmium toxicity to the bluegill (Lepomis
macrochirus Rafinesque). Trans. Am. Fish. Soc. 103: 729-735.
Macek, K.J., and W.A. McAllister. 1970. Insecticide susceptibility of some
common fish family representatives. Trans. Am. Fish. Soc. 99: 20-27.
McKim, O.M., and D.A. Benoit. 1971. Effects of long-term exposures to
copper on survival, growth, and reproduction of brook trout (Salvelinus
fontinalis). J. Fish. Res. Board Can. 28: 655-662.
Mount, D.I. 1968. Chronic toxicity of copper to fathead minnows (Pime-
phales promelas Rafinesque). Water Res. 2: 215-223.
Mount, D.I., and C.E. Stephan. 1967. A method for establishing acceptable
toxicant limits for fish - malathion and the butoxyethanol ester of
2,4-D. Trans. Am. Fish. Soc. 96: 185-193.
1969. Chronic toxicity of copper to the fathead minnow
(Pimephales promelas) in soft water. J. Fish. Res. Board Can. 26:
2449-2457.
Pickering, Q.H., and M.H. Gast. 1972. Acute and chronic toxicity of cad-
mium to the fathead minnow (Pimephales promelas). J. Fish. Res. Board
Can. 29: 1099-1106.
46
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SECTION 5
PHYSIOLOGICAL-BIOCHEMICAL ASPECTS OF WATER TOXICOLOGY
V.I. Lukyanenko'
The most important biological and economic problem for an industrialized
society today is the problem of "clean water". This problem stems from the
increasing rate of water consumption and the rising pollution of inland
water bodies by wastewaters, oil and oil products, and pesticides. Cur-
rently the volume of polluted wastewater discharged into our water bodies is
approaching 700 cubic kilometers. There are thousands of substances in this
wastewater which are toxic to organisms in some way or another. During the
last decade, the problem of preventing pollution in inland water bodies be-
came still more complicated due to the wide-spread use of agricultural
chemicals for pest control and plant protection. The list of pesticides
used in agriculture grows continuously larger. World production of these
toxicants in now approaching 1.5 million tons/year.
During the twenties and thirties of this century, a general concept of
protecting a water body from pollution was formulated both in our country
and abroad, so as to restrain the progressing chemical pollution of water
bodies and guarantees the multiple use of water resources. In accordance
with this concept, state regulation of wastewater discharge and control of
pollution by establishing maximum permissible concentrations (MPC) of harm-
ful substances discharged into water bodies was introduced. For quite
understandable reasons, the dominant role in solving the problem of pro-
tecting water bodies from pollution and establishing the MPC in water
bodies in our country belonged to the field of health specialists.
Medical specialists have done a considerable amount of work on this pro-
blem. They have developed ideas on hygienic criteria for the harmful
effects of wastewater; strengthened the physiological-biochemical direction
in studies of water hygiene and sanitary protection of water bodies, and at
present have experimentally proven the MPC of about 300 harmful substances
introduced into water bodies that exclude unfavorable effects of these sub-
stances on people's health (Cerkinskiy, 1949; Cherkinskiy, Krasowskiy,
1967). However, after a short period of time, it was learned that the MPC
of many substances (salts of heavy metals, insecticides) which fully satisfy
the health specialists do not guarantee the purity of water bodies from a
general biological standpoint or from the fishing industry's point of view.
HsNIORH, Astrakhan, USSR.
47
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The MPCs for many substances were found to be significantly lower for
fish and other organic organisms than for man. What is considered harmless
for man proves to be fatal for fish, especially under chronic exposure.
The lower tolerance of fish and other aquatic organisms to pollutants dis-
charged into water bodies is quite understandable because the polluted
water is their habitat.
All these facts drew our attention to the need for biological standardi-
zation of harmful substances in tests on fish and other aquatic organisms in
order to retain the normal course of biological processes and a high biolog-
ical productivity in them. But the solution to this central water toxi-
cology problem is senseless without detailed studies of character and means
of influence of harmful substances on the vital activity of aquatic
organisms on the whole, and fish in particular, accordi-ng to differences in
their organization, and taking into account their high sensitivity to many
toxicants of inorganic and especially nature. Therefore, understanding the
effect of toxicants on fish and other aquatic organisms is a necessary pre-
requisite for developing scientific bases and methods for determining MPCs
which are applicable to the problems of biological standards. We are
talking about a physiological-biochemical approach to solving the basic
problems of water toxicology, in research to develop biological standards
for protecting water bodies used by the fishing industry.
The first stage of research was to show the similarity and reaction pro-
perties of fish as cold-blooded vertebrates to various toxicants in compari-
son with warm-blooded vertebrates. As our studies on the phenol intoxica-
tion fish model, at the USSR Academy of Sciences, Institute of Biology of
Inland Waters in 1961-1964 and continued at the Central Sturgeon Research
Institute, showed, there are various pathological changes in fish organs
which affect the many physiological systems and which precede the death of
poisoned fish. Examples are a disturbance in the behavior reflex activity,
a disturbance in respiration, changes in the electrocardiogram, changes in
the activity of tissue cholines-terase and ammonia content in the brain, de-
crease in hemoglobin concentration and increase in blood sugar, changes in
albuminous content of blood serum, changes in ESR, and a series of other
hematological changes (3-8).
In the last few years, other laboratories have begun to follow this di-
rection, have added new data (9-18) on a variety of fish reactions and ways
toxicants influence aquatic organisms. All the experimental data available
today permits us to confirm that fish reactions to various groups of toxi-
cants, by direction as well as by content, are mainly similar to those re-
actions known for warm-blooded vertebrates. Basic principles and methods
for assessing toxicity developed by general and sanitary toxicology results
could, therefore, be used for aquatic toxicology. This is why we think
that the MPC of a harmful substance for fish, just as for higher terres-
trial vertebrates, should not exert a toxic effect on any of the numerous
facets of its vital activity.
In other words, the MPC should not exert a toxic effect on any of the
numerous functions of the organism since the disturbance of any of the func-
48
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tions might lead to the disturbance of the physiological normal level of
the organism and its biological well-being (3). Such an attitude is quite
natural for a physiologist, since from the physiological point of view, any
one of the numerous functional systems is equivalent and indispensable for
the normal activity of the whole organism, and the persistant disturbance
of each of them will inevitably destroy the activity of the others, in-
cluding the reproductive function.
According to this point of view, we have formulated ideas on toxicity
criteria in water toxicology. We must consider as toxic (threshold) a con-
centration which evokes some expressed pathological change in any of the
functional systems of an individual organism, since persistant disturbance
of the activity of any physiological system, whether it be blood circulation
or hemopiesis, respiration or nutrition, behavior or reproduction, sooner
or later leads to irreversible homeostatic disturbances and finally to the
destruction of organisms.
It appears from the above that the threshold concentration value deter-
mined from fish toxicological investigations to develop criteria, depend
largely on how correctly we identify the function affected, i .e., the ade-
quacy of methods for estimating pathological changes in the activity of the
functional system. Therefore, the search for more sensitive, specific
methods with high resolution which permit us in a very short time to obtain
scientifically substantiated MFCs for harmful substances, is of primary
importance. Already MPC values for almost a thousand various substances
discharged into water used for fisheries are needed.
The solution of this colossal problem in a very short time is possible
only with the aid of more sensitive contemporary physiological and biochemi-
cal methods for monitoring the functional state of test fish and other
aquatic organisms. These methods surpass 10-100 times the "fish trial-and-
error method" due to their resolution capabilities. The selection of the
specific method to determine the MPC value of any harmful substance must be
based on the knowledge of toxicodynamics of the substance under investiga-
tion and the understanding of the effect of various toxin groups, i.e., a
clear identification of the most sensitive target function. Here is an
example to explain this thesis.
For example, a large group of substances of an organic nature (toxins
of the phenol series, many pesticides, dyes, etc.) cause a complicated
symptom-complex of intoxication in fish. This permits us to assume the in-
fluence of these substances is on the central nervous system. However,
this required direct experimental proof and it was obtained (4, 3, 19-21)
using the model of phenol intoxication in fish. In a number of experiments
beginning in 1962-63 at the USSR Academy of Sciences Institute for Biology
of Inland Waters, the dominant role of the central nervous system in the
development of the complicated symptom-complex of intoxication in fish with
toxins of the phenol series has been proven. For example, we succeeded in
completely cutting off the first phase of phenol intoxication—rapid motor
activity—in anesthetized crucians (novocain, urethane). In other words,
generalized inhibition of the central nervous system, resulting from narco-
49
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sis, prevents the development of the most typical symptoms of phenol intoxi-
cation in fish. A further pharmacological analysis of phenol effects on
fish was carried out using curariform preparations (succinilcholine,
phlaxedil, paramyon) possessing pronounced blocking action on neuromuscular
conductivity in the myoneural synapse. The experiments indicated that
phenol does not exert a direct stimulating effect on the muscles of the
fish body, and that the nerve impulses arising from the central nervous
system are the basis for motor reaction of fish under the influence of
phenol. This is indicated by the inhibition of motor reaction in phenol-
exposed fish by means of a pharmacological disturbance in the neuromuscular
nerve impulse transmission within the region of the myoneural synapse. In
this respect succinilcholine, from the group of preparations producing
stable depolarization of the terminal motor plate, was the most effective.
K. Kuba's (22) electrophysiological work completely confirmed the results
of our experiments. He also concluded that the myoneural synapse was one
of the points in which phenol action plays a dominant role.
We obtained new evidence indicating the dominant role of the central
nervous system in the reaction of fish to phenol stimulation, and in parti-
cular, of stimulation of the brain in fishes with several spinal cords
(operative disconnection) and on an isolated head preparation (4, 3). The
brain proved to be prominent in the development of more characteristic com-
ponents of the reaction of fish to the toxic effect of organic toxins.
Motor activity occurred at the beginning and spasms later.
After the complete removal of the brain, not one of these reactions
developed. The spinal cord is the most important link in the reflex arch,
the conductor of the impulses which are caused by phenol stimulation, from
various branches of the brain to peripheral neuromuscular systems in fish
(3).
Naturally, the question of the specific phenol effect on the central
nervous system in fish was been raised. A partial answer to this question
was found during experiments with anticholinesterase preparations (phospha-
col, neostigmine, physostigmine). The preliminary injection of these pre-
parations into test crucians fully inhibited the external symptom-complex
of phenol intoxication in fish. Experiments with anticholinesterase pre-
parations led to the conclusion that the dynamics of an acetylcholine meta-
bolism and, first of all, of a system of acetylcholine-cholinesterase in
cholinergic synapses of the central nervous system and neuromuscular
synapses, plays a dominant role in the development of the complex of re-
actions caused by phenol (4). Confirmation of this point of view came from
biochemical data on changes in the muscular cholinesterase activity under
the influence of phenol (5), as well as from electrophysiological data.
According to the latter, phenol increases the amplitude of the stimulating
synaptic potential and causes the appearance of tiny potentials on the
terminal plate, i.e., it facilitates neuromuscular transmission. It is
fascinating to explain the biphase course of phenol intoxication in fish,
namely, the intitial highest motor stimulation with subsequent spasms and
paralysis, in light of the dynamics of change in the acetylcholine concen-
tration in cholinergic structures of the central nervous system and in
50
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myoneural synapses. The initial highest central stimulation caused by
phenol appears to be connected with the stabilization of "physiological"
acetylcholine and its accumulation in cholinergic synapses. Central paraly-
sis, coming after the stimulation and caused by phenol, might be understood
as the result of the accumulation of acetylcholine in brain synapses in
ordinate pessimal concentrations which cause the inhibition. Certainly,
the acetylcholine accumulation in synapses might be caused by two methods--
either due to cholinesterase inactivation or due to an increase in acetyl-
choline quanta isolation from nerve endings, but most likely both processes
take place. However, it is not of particular concern, since both methods
lead to acetylcholine accumulation in synapses in pessimal concentrations.
Such are the basic results of the experimental studies into the mechan-
isms of phenol effect on fish, which we attempted in'the early sixties.
They permitted us (4) to substantiate prospective use of the behavior-
reflex method during ichthyotoxicological experiments as the most sensitive
tests for determining the chronic effect of trace concentrations of various
homologues of the phenol series and for determining the MPC value of this
group of substances. Experimental data obtained by B.A. Flyorov and V.E.
Matey in their experiments on gold crucians and Lebistes groups (23-26)
fully supported this point of view. As should be expected, pathological
changes in behavior reflex activity in fish occur long before the appearance
of expressed phenol intoxication symptoms and are observed in concentrations
that are 5-8 times lower than acute toxic concentrations of this substance.
There are good reasons to believe that the behavior reflex method will hold
a fitting place among other methods for determining the MPC of various
groups of organic toxins with expressed activity on the central nervous
system which is characteristic for them.
The growth of international cooperation in the field of water toxicology
and ichthyotoxicology brings our attention to the question of unifying
methods for estimating the degree of toxicity for various groups of harmful
substances, and standardizing experimental conditions and principles for
interpreting test data. At present, this is quite possible, thanks to the
accumulation of data concerning the dependence of the results from ichthyo-
toxicological experiments on many variables. In this case, two groups of
factors characterizing both fish habitat (chemical water state, oxygen con-
tent, pH value, water temperature, etc.) and the test object itself
(species, age, and sexual properties of fish sensitivity and stability as
well as initial functional state) play a dominant role. Therefore, in
order to develop an actual and potential toxicity for a harmful substance,
it is necessary to carry out experiments which allow for fluctuations in
physical-chemical parameters of the water medium (27), i.e., experiments
carried out at a relatively high temperature and a moderately low oxygen
content. Water hardness and pH value are selected in such a way as to
develop the highest possible toxicity for the substance under investigation.
The most sensitive species of fish of the ichthyofauna under investigation
(28) should be used as the test object. In this case, it is important to
consider the sensitivity of the test species at various stages of ontogene-
sis, and choose the most sensitive one (29, 3). Together with the physical-
chemical factors of the water environment and the biological properties of
51
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the test object, the exposure time plays a dominant role in determining an
ichthyotoxicological experiment. Depending on the specific problem before
an investigator and on the degree to which the substance of interest is
studied, the duration of an experiment must be limited to 24-96 hours
(acute experiment), to 10-30 days (subacute experiment), or 1-3 months
(chronic experiment). The duration of an experiment is determined by the
resolving powers of the research method used by an investigator. The
higher the sensitivity and resolving powers of a method, the shorter the
time for determining the toxicity of a substance and the MPC value. Just
as the physiological-biochemical methods for toxicity determinations are
different, the mechanisms of the effect of various toxin groups are not the
same (30). There is no doubt that unifying methods for estimating the de-
gree of toxicity of substances and determining the MPC for these substances
will bear fruits in the near future and facilitate the solution of the
basic problem of water toxicology which faces all water toxicologist--of
limitation biologically harmful substances discharged into water bodies—by
determining the MPC. The solution of this problem is possible only on the
basis of a synthesis of the ideas and general and sanitary toxicology
methods with the achievements of modern physiology and biochemistry of fish
and other aquatic organisms (31).
REFERENCES
1. Cherkinskiy, S.N. In: Sanitary Protection of Water Bodies From Indus-
trial Waste Water Pollution. Moscow, 1949, 52-81.
2. Cherkinskiy, S.N., and Krasovskiy, 6.N. In: Industrial Pollution of
Water Bodies. Issue 8, Moscow, 1967, 5-19.
3. Lukyanenko, V.I. Toxicology of Fish. Moscow, 1967.
4. Lukyanenko, V.I. In: Problems of Hydrobiology. Moscow, 1965,
261-263.
5. Lukyanenko, V.I. and Petukhova, L.A. In: Biology of Fish in the Volga
Water Reservoirs. Moscow-Leningrad, 1966, 311-318.
5. Sorokin, Yu. I., and Lukyanenko, V.I. Pharmacology and Toxicology.
1966, No. 1, 109-110.
7. Lukyanenko, V.I., Geraskin, P.P., Sedov, S.I. and Kokoza, A.A. In:
Some Problems of the Sturgeon Industry in the Caspian Basin. Mos-
cow, 1966, 68-74.
8. Lukyanenko, V.I. Problems of Ichthyology. Issue 3, 1967, 7, 547-562.
9. Metelyov, V.V. In: Experimental Water Toxicology. Issue 2, Riga,
1971, 104-121.
10. Zvirgzds, Yu.K., Latse, Z.M., Grundule, M.V., and Zuyka, A.A. In: Ex-
perimental Water Toxicology. Issue 2, Riga, 1971, 12-25.
52
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11. Komarovskiy, F.Ya. In: Symposium on Water Toxicology. Leningrad,
1969, 68-69.
12. Komarovskiy, F.Ya., and Popova, G.V. In: Symposium on Water Toxi-
cology. Leningrad, 1969, 68-69.
13. Popova, G.A. The Character of the Effect of Some Herbicides on Carp.
Petrozavodsk, 1973, Autoref. cand. dis.
14. Volkov, I.V. Experimental Investigation of Blood Physiology Under
the Influence of Unfavorable Factors of the External Environment.
ref. cand. diss.
15. Samylin, A.F. Influence on Some Toxicants on Fish During the Early
Period of Ontogenesis. Petrozavodsk, 1974, Autoref. cand. diss.
16. Osetrov, V.S. Studies of Toxic Action of 5,4-dichlorsalycilalanilide
On Carp. Moscow, 1972. Autoref. cand. diss.
17. Cernyshev, V.I. Some Physical-Chemical Aspects of Toxicology of
Fishes. 1969, Autoref. cand. diss.
18. Slava, E.E. Biophysical Aspects of Water Toxicology. Riga, Autoref.
cand. diss.
19. Lukyanenko, V.I. Summaries of Reports From the All-Union Conference
on Problems of Water Toxicology. Moscow, 1968, 16-17.
20. Lukyanenko, V.I. Summaries of Reports From the All-Union Conference
on Problems of Water Toxicology. Moscow, 1968, 49-51.
21. Lukyanenko, V.I. In: Problems of Water Toxicology. Moscow, 1970,
154-162.
22. Kuba, K. Japan J. Physio!., 1969, 19.
23. Flyorov, B.A. Hydrobiol. 0., 1965, 3, 49-53.
24. Matey, V.E. Summaries of Reports From the All-Union Scientific Con-
ference on Problems of Water Toxicology. Moscow, 1968, 51-52.
25. Matey, V.E. Hydrobiol. J., 1970, n. 3.
26. Matey, .V.E. and Flyorov, B.A. In: Problems of Water Toxicology. Mos-
cow, 1970, 175-181.
27. Lukyanenko, V.I. In: Biophysical Aspects of Biospheric Pollution.
Moscow, 1973, 88-89.
28. Lukyanenko, V.I. and Flyorov, B.A. Hydrobiol. J., 1965, 2, 48-53.
53
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29. lukyanenko, V.I. and Flyorov, B.A. Pharmacology and Toxicology. 1963,
5, 625-629.
30. Lukyanenko, V.I. In: Symposium on Water Toxicology. Leningrad, 1969,
78-79.
31. Lukyanenko, V.I. In: Experimental Water Toxicology. Riga, Issue 4,
1973, 9-29.
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SECTION 6
BIOENERGETIC AND OTHER CONSIDERATIONS IMPORTANT IN THE STUDY OF
WATER QUALITY INFLUENCES ON FISH GROWTH
Peter Doudoroff
The determination of maximum permissible concentrations of water pollu-
tants in waters in which fish must be adequately protected has long been a
major objective of physiological, toxicological, and ecological research in
the field of water pollution in the United States and in the Soviet Union.
We have fully recognized that prevention of fish mortalities is not alone a
sufficient goal of pollution control directed toward the protection of
fisheries, for unimpaired fish production clearly depends upon adequate re-
production and normal growth of fish, and not only on their survival.
However, research into the effects of water pollutants on fish growth has
been limited, and much of it is rather crude or superficial and not very
helpful in the prediction of effects under natural environmental condi-
tions. Effects on reproduction, which certainly is essential to fish pro-
duction, have received more attention. But some interference with reproduc-
tion of fish may sometimes have relatively little or no effect on their
production, because numbers of young produced remain sufficient for nearly
full utilization of the available habitat and food resources. Production
(elaboration of new tissues) may even increase under some circumstances,
because reduced competition for food permits a larger portion of the food
materials and energy to be utilized for accelerated growth and less for
maintenance metabolism. On the other hand, any interference by pollution
with the growth of the young must necessarily result in impairment of
production.
To understand effects of environmental factors on the growth of
animals, it is essential to consider to what extent and to what ends the
energy and materials of food consumed by the animals are utilized and how
they are distributed, under the different conditions, among portions having
different fates. It is not my purpose to expound in depth here on the
principles of the science of bioenergetics and their application in the
study of the growth of fish, a field of research to which Soviet
scientists, notably the late V.S. Ivlev and also G.G. Vinberg, have made
outstanding contributions. Nor can I treat fully the influence of water
pollutants on fish growth. A more detailed exposition of the principles
involved and their applications in pollution research can be found in the
chapter on "Bioenergetics and Growth" of Warren's (1971) "Biology and Water
Pollution Control" (p. 135-167). My purpose here is, mainly, to- propose
and explain a general scheme of procedure, based on important bioenergetic
55
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and ecological considerations, for efficient experimental investigation of
undesirable effects on fish growth of water pollutants, especially toxic
ones. The proposed studies are directed toward reasonably reliable
estimation of limits of water quality alterations having virtually no such
harmful effects in nature. Because of the unavoidable complexity and
difficulty of sufficiently thorough investigations directed toward the
stated goal, it is clearly desirable to minimize the number of the most
difficult and costly experiments to be performed. In the plan of
investigation suggested here, the sequence of different experiments will
provide, in my opinion, for maximal efficiency of the studies of a
practical nature, which should not be expected to lead to complete
understanding of observed effects of pollutants.
A reduction of growth rates of fish can be a consequence of any one or
more of the following effects of degradation of water quality: (1) reduc-
tion of the available food supply; (2) impairment of the appetite of the
fish for food; (3) reduction of the feeding activity of the fish and their
ability to find and capture their prey; and (4) impairment of the effi-
ciency of metabolic utilization of food by the fish and its conversion into
body tissues. We must look for each of these possible effects.
Maximum concentrations of toxic water pollutants having no material
effect on over-all food resources are very difficult to determine experi-
mentally, because most fish depend, directly or indirectly, on a large
variety of aquatic organisms for their food supply. Demonstrable adverse
effects on the reproduction and growth of some food organisms may not be
assumed to bring about a lower availability of food, since the production
of other, more tolerant species that can be utilized by fish may increase
as competition for space and food by the less tolerant species declines.
In waters polluted with energy-rich organic materials, such as pulp mill
wastes, the abundance of fish foods in the aggregate sometimes even
increases, while some disappear, because the growth of some microorganisms
that serve as a primary food resource for invertebrates is stimulated.
Often, therefore, reliable prediction of effects of pollutants on the
availability of fish foods in natural habitats can be achieved only through
experiments performed under nearly natural conditions, such as experiments
with artificial or modified natural streams in which complex plant and
animal communities can be maintained. The results of such costly experi-
ments often are difficult to interpret if there are no good reasons to
believe that any observed impairment of the growth of fish could have been
due only to a reduction of availability of foods and not to one of the
other possible causes mentioned above. Obviously, it is not with difficult
experiments of this kind that one should begin in seeking to reduce the
number or range of concentrations of a pollutant that need to be tested in
other additional experiments.
On the other hand, the maximum concentration of a toxic substance
having no pronounced adverse effect on the appetite of juvenile fish, or on
the highest rate of food consumption of which they are capable, or on their
efficiency of utilization of food resources if their activity is not
materially depressed by the poison at that level, can be quite easily deter-
mined through laboratory experiments. To find the level that does not
56
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impair the appetite, it is necessary only to expose groups of fish for
sufficiently long periods to different concentrations in small aquaria, to
supply them continuously or frequently with as much attractive, palatable,
and nutritious food (preferably live food) as they will consume, and to
measure the food consumed. The fish, held preferably in individual aquaria
or compartments, should be uniform in initial size and carefully weighed.
By weighing accurately the food offered and the uneaten food removed from
the aquaria at daily or other suitable intervals, the mean daily consump-
tion (in grams per gram of fish) at each tested concentration of toxicant
can be determined and compared with that of controls. The gross efficiency
of food conversion can be determined by dividing the gain in weight of the
fish during an experiment or a suitable time interval between weighings by
the weight of food consumed during that period. This efficiency can be
expected to be reduced, as compared with that of controls, when the food-
consumption rate is markedly reduced. If it is found to be reduced also at
a concentration of a poison at which the food consumption is not reduced,
impairment of metabolic processes is indicated, and lower concentrations
must be tested to determine the highest one at which no such effect is de-
monstrable. The duration of the food-consumption and growth tests need not
exceed a month and can be much shorter when growth is rapid, but it is ad-
visable to expose fish for a fairly long time before final measurements of
their growth and food intake are made, especially when substances known to
be accumulative poisons are tested.
Any concentration of a poison at which the food intake is found de-
finitely to be reduced can be taken to be level at which the growth of the
fish probably would be impaired under natural conditions whenever the avail-
ability of food is not a limiting factor. This statement, or proposal, can
be reasonably countered, however, with the objection that the food consump-
tion and growth of fish in their natural environment generally are limited
by the availability of food and not be the appetite of the fish. One may
well argue that, for this reason, the concentration of a poison at which
food consumption begins to decline in aquaria where food is so abundant
that the fish can obtain as much as they can eat with little or no effort
is essentially meaningless. In addition, it is doubtless true that the
annual food consumption of fish in nature is, as a general rule, if not
always, far less than their annual assimilative capacity. At natural tem-
peratures growth rates of well-fed fish in laboratory can greatly exceed
those usually found in nature, where losses of weight during periods of
food shortage are not unusual. Perhaps few biologists who have studied the
growth of fish believe that the availability of food ever is not a limiting
factor for any considerable periods of time. But it seems to me not un-
reasonable to assume that in some very productive natural environments the
rates of food consumption are not limited by food availability during some
fairly extended periods of maximal or nearly maximal abundance of food in
the season or seasons in which most of the growth of fish takes place. An
inability of fish to take full advantage of a temporary abundance of food
in such a situation could have a considerable effect on their annual weight
gains.
Whenever there is sufficient reason to reject the proposition that the
growth of fish in a given environment is not food-limited at certain times
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and for periods long enough to be of consequence, as when that growth is
known to be always much slower than that of controls in the proposed labora-
tory tests, a different alternative approach can be used. Each fish in
the aquaria then can receive daily a uniformly restricted food supply that
is believed or assumed to be not much less than the maximum amount of food
consumed per day by individuals of the same size in the natural environ-
ment, excepting rare occasions. The maximum concentration of a poison at
which this restricted food ration is fully consumed by the experimental
fish and their growth is not demonstrably impaired, as compared with that
of controls, than can be determined. In all such experiments the food
should be as much as possible like natural foods; for comparative purposes,
amounts of these foods should be expressed not in grams but in caloric
equivalents. Water temperatures normal for the season of maximum food con-
sumption in nature should be maintained during these tests. Various
methods for the estimation of rates of food consumption by fish in their
natural environments have been described (Da\n's and Warren, 1968). Esti-
mates of amounts of food consumed during short intervals of time probably
cannot be very reliable. Average rates of growth of fish in nature during
longer periods (seasons of the year) can be evaluated through appropriate
observations, and their average daily food consumption during these periods
can be estimated through laboratory experiments by which rates of growth in
aquaria are related to food-consumption rates. In the absence of contrary
information, the assumption then can be made that a food-consumption rate
twice the average seasonal rate is a rate that should be attainable but
does not have to be exceeded at any time during corresponding season if
growth in the natural environment is not to be materially impaired.
Accordingly, a daily feeding rate twice the estimated, average, natural
food-consumption rate during the season of maximum food intake can be taken
to be appropriate for the proposed toxicity tests. Admittedly, this recom-
mendation is somewhat arbitrary, for frequent fluctuations of natural food
availability and consumption may be sometimes much smaller and sometimes
greater than those that it implies, but I believe that it cannot lead to
serious error in the estimation of maximum harmless levels of water pollu-
tants.
When an unrestricted food supply is provided to the experimental and
control fish, as first suggested, it will sometimes be found that fish ex-
posed to toxic substances consume more food, and not less, than the con-
trols. In this way they may compensate partly or wholly for a reduction of
the efficiency with which the consumed food can be utilized, so that growth
in the aquaria may be reduced little or not at all. But such compensation
is possible only when food is extremely abundant and available. It is cer-
tainly true that the growth of most fish in nature is limited, at least
during a large part of each year, by the limited availability of food,
which renders the compensation impossible. In their natural habitat
poisoned fish are likely to find and capture less of their prey than normal
fish would when there is a shortage of food. Therefore, any reduction of
the efficiency of food conversion in fish affected by poisons must result
in impairment of their growth whenever the density of their prey is low
enough to be a limiting factor.
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Increases of food consumption by fish together with impairment of the
efficiency of their utilization of food in aquaria in which the food was
abundant have been observed in experiments on the effects of sublethal con-
centrations of sodium cyanide (Leduc, 1966) and of potassium pentachlor-
phenate (Chapman, 1965; Warren, 1971, p. 163-164). Cyanide levels not far
below lethal levels depressed the food conversion efficiency (gross) of ci-
chlids, Cichlasoma bimaculatum, by as much as 35% on the average, but
caused increases of the amount of food consumed during the 36-day tests
that averaged as much as 30%. Potassium pentachlorphenol (0.2 mg/liter) at
first depressed the food-consumption rate of these fish, but later in the
42-day tests the consumption rate of the poisoned fish was much greater
than that of controls. Growth rates were initially depressed by both the
cyanide at high concentrations and the pentachlorphenate, but before the
tests were concluded the fish exposed to the poisons were growing even
faster than the controls. Averaged weight gains of the experimental and
control fish during the entire experiments therefore differed little or not
at all. When food rations were uniformly restricted, however, the growth
of fish in the pentachlorphenate solutions was decidedly less than that of
controls because of the deranged metabolism and reduced efficiency of food
conversion.
Whenever the food consumption of fish that are given all the food that
they can eat is found to increase in the presence of the toxicant studied,
the food supply should be uniformly restricted so that all of the fish,
including the controls, will consume all of the food offered. The re-
stricted daily ration in one series of tests should not be much less,
however, then the maximum ration found to be consumed by all the fish. The
highest concentration of the toxicant that does not demonstrably impair the
growth of the fish receiving this fixed ration should be determined. The
impairment of food-conversion efficiency by poisons can be much more pro-
nounced when the levels of food availability and consumption are high than
when they are low. Sometimes, however, the effect in question may be more
pronounced and readily demonstrable at low levels of food availability and
intake than at high levels. Therefore, in looking for possible inter-
ference with food-conversion efficiency, it is always advisable to perform
an experiment in which a small, uniformly restricted ration of food is pro-
vided to each fish, regardless of whether the appetite of the fish has or
has not been found to be stimulated by the poison tested. This ration
should not be much greater than the maintenance ration for the controls,
which is the ration that brings about neither growth nor loss in weight of
these fish. Should the food-conversion efficiency of fish receiving the
small, restricted ration be found to be impaired more markedly by a poison
than that of fish consuming much larger amounts of food, tests with this
small ration should be performed to determine the highest concentration of
poison that does not impair the conversion efficiency. Ideally, such
series of tests of different concentrations are performed at four or more
levels of food availability, ranging from unrestricted supplies to re-
stricted supplies barely sufficient or insufficient for maintenance of the
initial body weight of controls (Warren, 1971). Such more laborious experi-
ments certainly can be very instructive and can increase confidence in the
reality of observed small differences of food-conversion efficiency. They
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have not been shown, however, to be absolutely necessary for achieving the
not very ambitious objective of the practical investigations proposed here.
The maximum concentration of a toxic pollutant that impairs neither the
appetite nor the food-conversion efficiency of the experimental fish so
much that their growth under natural conditions must be judged likely to be
seriously affected having been determined, what should be the next step?
May we now assume that this concentration can affect the growth of the fish
adversely only if it causes a reduction of the natural food supply? This
assumption is not usually justifiable, because an impairment of water
quality can reduce also the activity of animals and, consequently, their
success in seeking, pursuing, and capturing their prey, as well as in
eluding their enemies. Experiments have shown that exposures of salmonid
fishes to exceedingly low concentrations of sodium cyanide, for example,
has a very pronounced, rapidly produced, and lasting effect on their swim-
ming ability. Though not known to be otherwise demonstrably affected, they
become unable to resist currents of moderately high velocity nearly as long
as individuals not exposed to the cyanide (controls). Such effects may or
may not interfere with normal feeding activities. In experiments with
other fish that proved less susceptible, Cichlasoma bimaculatum, Leduc
(1966) found that the duration only of moderately rapid swimming, and not
that of very rapid swimming, was reduced by exposure of the fish to low
cyanide levels. A reduction of maximum swimming speeds sustainable for
long periods of time cannot be said obviously to impair the foraging
efficiency of fish; only short bursts of speed in the pursuit of prey are
commonly observed and are clearly essential to successful feeding of many
species. Still, any interference by toxic substances with the ability of
fish to exert themselves may, in some subtle way, cause food consumption
under natural conditions to decline, thus reducing growth. Some poisons
may interfere mainly with very rapid swimming.
In experiments with artificial, concrete-lined ponds stocked with known
numbers of mosquitofish, Gambusia affinis, and several largemouth bass,
Micropterus salmoides, which fed on the mosquitofish, the food consumption
and growth rates of the bass were decidedly reduced when dissolved oxygen
concentrations were reduced (Brake, 1972; Warren, Doudoroff, and Shumway,
1973). The bass consumed fewer of the mosquitofish and grew less in a pond
with a moderately reduced oxygen concentration than in a control pond, even
though laboratory tests had shown that they were capable of consuming much
more food and of growing much faster at the same oxygen concentrations and
temperatures when the food was more available. The mosquitofish were pro-
vided with artificial cover (rolls of wire netting placed in the shallow
water near the periphery of each pond) and were not easily caught by the
bass. Consequently, the food-consumption and growth rates of the bass even
at high oxygen concentrations were dependent on the density of the prey,
that is, they increased when the number of mosquitofish placed in the ponds
was increased. Aquarium tests had shown that the appetite of the bass, or
the amount of food that they are able to consume when the supply of food is
unlimited and the prey easily captured (not protected) is reduced at
moderately low oxygen concentrations just as the food consumption in the
ponds was reduced (Lee, 1969; Warren et al., 1973). The impairment of
appetite could have been somehow partly responsible for the reduction of
60
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food consumption at the low oxygen concentrations in the ponds. However,
since the bass in the ponds were obviously unable fully to satisfy their
appetite for food at any of the tried oxygen levels, this reduction of
their food consumption must have been due primarily to their reluctance to
expend as much energy in pursuing their prey at the low oxygen levels as
they expended in the presence of more dissolved oxygen.
The amount of energy that the bass expended in capturing and assimila-
ting their food in the ponds at like temperatures and high oxygen concen-
trations was virtually independent of the density of the prey. In other
words, the average rate of their metabolism was not appreciably affected by
changes in availability of the food. As food became less available, so
that less was consumed, less energy had to be expended in the process of
assimilation of the food ingested and more could be expended, therefore, in
pursuing the prey, but the total energy expenditure did not increase or de-
crease demonstrably. The average metabolic rate of the bass in the ponds
was estimated by the energy balance (or caloric apportionment) method. The
caloric value of the unassimilated portion of the food and of metabolic
wastes (estimated through laboratory experiments, using small aquaria) and
the measured increment in caloric value of the bodies of the growing bass
were subtracted from the caloric value of the food found to have been
consumed by the bass during an experiment; the remainder, in calories, was
then divided by the mean caloric value of the bodies of the bass, in kilo-
calories, and by the duration of the experiment, in days. Fairly uniform
values of about 26 cal/kcal per day were thus obtained at temperatures near
21 C (Lee, 1969; Doudoroff and Shumway, 1970).
At moderately reduced oxygen concentrations the feeding activity of the
bass was reduced evidently because the metabolic rate was limited by the
oxygen supply, and this must have been the reason also for the reduction of
the appetite of the bass in aquaria at the same oxygen concentrations.
When held in aquaria with an unlimited food supply (easily captured
mosquitofish), the bass had at high oxygen concentrations an average
metabolic rate approximately equal to that of more active bass in the
ponds. Reduction of the oxygen concentration thus can be expected to limit
the appetite of fish whenever it impairs their feeding, and vice versa.
Effects of some toxic substances on feeding activity and appetite may be
similarly related, but those of other poisons may be quite unrelated
phenomena. It is highly probable that some toxic substances, at concen-
trations that impair neither the appetite nor the gross food-conversion
efficiency of fish held in aquaria or even improve one or both of them,
nevertheless reduce the rate of growth of the fish under natural conditions
by limiting their feeding activity. Such an effect has not yet been
clearly demonstrated, probably only because the appropriate experiments
have not been performed. But we surely may not assume that a reduction of
feeding activity will always be accompanied by an impairment of appetite as
it apparently is when oxygen concentrations are reduced.
As I have pointed out already, cyanide poisoning can greatly restrict
one kind of activity of fish, at least, while causing an increase of their
appetite for food; for reasons to be soon apparent, I believe that it can
restrict spontaneous (not enforced) activity also without impairing the
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appetite. Also, we certainly may not assume that feeding activity cannot
be materially restricted at cyanide levels that have very little or no ad-
verse effect on the efficiency of utilization of food for growth by fish in
laboratory aquaria and on their consumption of food that can be procured
with almost no effort. Although extremely low cyanide concentrations
greatly impaired the swimming ability of young coho salmon, Oncorhynchus
kisutch (Broderius, 1970), much higher concentrations not far below lethal
levels were found, in a sinlgle experiment performed by Leduc (1966), to
have no persistent, adverse effect on their food consumption and conver-
sion efficiency to aquaria. Indeed, after an initial reduction of both
food intake and food-coversion efficiency during the first 12 days of ex-
posure the gross conversion efficiency considerably exceeded that of
controls. This result needs verification, but there is no very good reason
to doubt its validity. The efficiency of food conversion probably
increased, as compared with that of the controls, because of reduction of
the activity of the usually quite active fish in the cyanide solutions, in
which more of the assimilated food consequently could be utilized for
growth. Had the fish been required to remain normally active, a very
different result probably would have been obtained. To ensure complete
validity and comparability of laboratory measurements of food-conversion
efficiency, uniform, moderate activity of all test subjects must be somehow
enforced, but this is very difficult to accomplish. When this is not done,
the results of detailed studies of food-conversion efficiency at a number
of different levels of food intake and toxicant concentrations are
certainly not without interest or value, but, for the reasons indicated,
such a laborious study may not be quite as profitable an exercise as it may
appear to be. I believe that effort devoted to feeding-activity studies
can be more profitable, and that whenever an impairment of the efficiency
of food utilization for growth is masked in aquarium tests by a depression
of activity, the harmful effect of a poison will be revealed by appropriate
tests for reduction of feeding activity. The activity of fish in the
aquaria is largely spontaneous and unrelated to feeding, but feeding
activity, which is not enforced activity, can be expected to be depressed
by a poison whenever spontaneous activity is suppressed.
I have discussed in much detail the relation and distinction between
appetite for food and feeding activity or foraging efficiency, and how they
can be affected by water quality changes, because I believe that many
biologists do not sufficiently realize the need for food-consumption
studies designed to measure something other than the appetite or
assimilative capacity. Very few studies of effects of water pollution on
the foraging activity and success have been undertaken in the past. I
believe that much effort can be profitably devoted to the development of
methods for such investigation.
Sufficiently instructive tests for impairment of the feeding activity
or efficiency of small fishes that feed on plankton or on benthic inverte-
brates such as amphipods in standing or gently flowing waters apparently
can be quite simple, requiring little space and no elaborate facilities.
One can introduce a limited number of the food organisms into each of
several large aquaria and determine how many of them are consumed in a
certain period of time by hungry fish that have been held in the aquaria
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for some time in the presence or in the absence of a poison. The number of
food organisms introduced into each aquarium should be such as to make it
impossible for the fish to become satiated. It can be less than the number
that can be consumed by the fish at once when the food organisms are very
abundant and easily found and caught, or it can be greater than that number
if the initial density of the food organisms or the cover provided for them
are such that these organisms are not too vulnerable to predation. The
food organisms should be as uniform in size as possible. The food
organisms remaining in the aquaria with and without the toxicant being
tested can be counted when only a few remain in the control aquaria. If
the foraging activity and efficiency of the fish are unaffected at a tested
concentration of poison, about as many of the food organisms, on the
average, should remain in the aquaria with the contaminated water as in the
control aquaria at the end of the test period. Because of the progressive
decline of the numbers of food organisms in the aquaria during a test, the
experimental and control fish will be confronted with a desirable variety
of foodorganism densities in such tests.
For experiments with large fish, small, artificial ponds like those
that have been used in the already mentioned experiments on the influence
of dissolved oxygen on the feeding and growth of largemouth bass can be
used. However, such ponds are costly and require much space, and the
maintenance in them of constant concentrations of toxic pollutants, by
sufficiently rapid replacement of the water or otherwise, can be
difficult. Therefore, experiments with laboratory models of more modest
size have been undertaken recently. In these exploratory tests long, rec-
tangular aquaria are being used, with a shelf made of fine-mesh wire or
plastic netting suspended at each end a short distance below the water sur-
face. Mosquitofish (Gambusia) with which these tanks are stocked soon
learn to use the area over each shelf as a refuge, remaining there most of
the time and escaping to one of these sanctuaries if they can when they are
pursued by a largemouth bass also placed in the aquarium. The bass catch
some of the mosquitofish that spontaneously leave the protection of the
cover from time to time, or that the bass are able to flush from the cover
by some maneuver, but they are unable to follow the prey in the shallow
water above the shelves and capture it there. Consequently, they cannot
fully satisfy their appetite, and their food consumption and growth are
dependent on the density of the prey, just as were those of the bass in the
ponds. Other things being constant, any reduction of their foraging vigor
and agility must result in a reduction of food intake and slower growth.
Because of differences of the foods and feeding habits of fish of
different kinds in various waters, experimental methods suitable for the
study of effects of water pollution on the foraging activity and effi-
ciency of some species are unsuited to other forms. The contriving of the
most appropriate methods sometimes may not be easy, challenging the most
imaginative and inventive biologist's ingenuity. But experiments that are
very easily designed or standardized soon cease to be interesting. Artifi-
cial streams with circulated water can be used for tests with fishes that
normally inhabit rapidly flowing waters.
63
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In all such experiments it is most important to ensure that the
activity and behavior of the food organisms (prey) will be unaffected by
the pollutant at the concentration tested, or will be affected, at most,
far less than the foraging activity of the test subjects is affected, if
there is any effect. A debilitating effect on prey of the kind selected
more serious than the effect on the predator can result in improvement of
the predator's foraging efficiency under the experimental conditions. This
improvement would not necessarily occur in nature, where important food
organisms can be ones relatively resistant to the pollutant, and it
certainly would conceal an actual reduction of the predator's activity.
The food organisms selected should, therefore, be of a kind found through
preliminary experiments to be relatively insensitive to the pollutant that
is being tested, as compared with the fish that is the test subject. For-
tunately, the mosquitofish, Gambusia affinis, and other related species
with similar habits are handy fishes highly resistant to many poisons and
highly suitable for use as prey in the experiments and in other respects.
On the other hand, many of the predaceous species most valued by man as
food and game fishes, especially those of the family Salmonidae, are re-
latively sensitive. Suitable food organisms to be preyed upon by small
fishes usually should not be very difficult to find. Air-breathing aquatic
insects such as mosquito larvae and pupae, which are not sensitive to many
dissolved toxic substances, as well as to dissolved oxygen deficiency, and
young of the hardy brine shrimp, Artemia salina, hatched in the laboratory
from eggs that can be easily purchased, should not be overlooked in seeking
suitable forms.
Since most poisons at low concentrations at which the appetite and food-
conversion efficiency of fish used as test subjects are not materially
affected can so affect the activity of resistant food organisms only after
fairly long exposures, the duration of exposure of the food organisms to a
poison usually should be minimal. They should be kept usually in clean
water while the test subjects are being exposed to a toxicant in prepara-
tion for a foraging activity test. Small fish to be used as prey can be
accustomed to the experimental environment and trained to avoid the preda-
tor by holding them for some time under the test conditions but in the ab-
sence of the poison until they are subjected to attack by fish previously
exposed to the poison. However, if adverse effects of relatively high con-
centrations of poison are found to be produced very rapidly and subse-
quently to become less pronounced because of acclimation of the organisms
to the poison, a different procedure may be advisable. All of the food
organisms, including those to be fed to control fish, then can be
acclimated before a test to the lower concentration to be tested. Various
other ways to minimize effects of the tested impairment of water quality on
the food organisms may be possible. For example, some clean water could be
continuously introduced into experimental aquaria over the productive
shelves described above that serve as cover for the prey (mosquitofish).
The resulting, unavoidable dilution of the pollutant in the bulk of the
aquarium water could be compensated for by continuous introduction of a
sufficiently strong solution elsewhere in the tanks.
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By the various experimental methods that have been discussed in some de-
tail, a limit of water quality alteration not likely to have any consid-
erable effect on the feeding and growth of the experimental subjects when
the availability of food is constant can be determined. One can then pro-
ceed to the final step of the proposed scheme for the estimation of the
limit of alteration having no material effect of any kind on the growth of
the fish in their natural habitat. Possible effects on the food supply
(production and availability of food organisms) next can be investigated
experimentally with model environments, such as artificial or modified
natural streams or ponds, in which foods are produced naturally and fish
depend entirely on this natural production (except for some consumption of
terrestrial.organisms that may unavoidably enter into their diet). Experi-
mental facilities and methods for such investigations cannot be adequately
described here. Suitable methods that have been used at Oregon State Uni-
versity to study the effects of enrichment of water with sucrose on the
foods and growths of trout in a modified natural trout stream have been des-
cribed by Warren et al. (1964). Those used more recently in a similar
study of effects of pulp and paper mill wastes in outdoor, artificial
streams have been described briefly by Warren et al. (1974). Only the
highest concentration of a toxic water pollutant that has been found not to
affect fish directly so as to impair their feeding and growth may need to
be tested in the difficult and costly experiments in which natural condi-
tions must be reproduced as nearly as possible. This concentration may be
found to have no demonstrable, adverse effect also on the food supply and
growth of the fish in the simulated natural environment. However, if a
considerable reduction of the growth of the fish is observed at this con-
centration, lower concentrations must be tested to determine the level at
which there is no such effect and, therefore, no material effect on the
availability and consumption of food.
In the foregoing discussion emphasis has been placed on toxic water
pollutants, but effects of reduced concentrations of dissolved oxygen on
the growth of fish also have been considered. The literature on the latter
subject has been critically reviewed and the significance of the then
available data thoroughly discussed by Doudoroff and Shumway (1970). In
much the same way as oxygen deficiency, some toxic substances that inter-
fere with external respiration may very markedly impair the appetite of
fish but have little or no effect on their food-conversion efficiency when
the fish receive uniformly restricted food rations. Because of the well-
known influence of temperature on the metabolic rates of fish, thermal
pollution, now of great importance, presents some special bioenergetic
problems (Doudoroff, 1969) that can be only very briefly discussed here.
The temperature optimum for the growth of fish is a function of the
food supply. At moderately elevated temperatures, fish may be able to grow
much faster than they do at lower, normal temperatures when the food supply
is unlimited, but more food is needed for mere maintenance of their body
weight, because of the elevated metabolic rate. When the supply of food is
limited so that the daily consumption cannot increase, growth is reduced as
the temperature rises, not because of any impairment of, or interference
with, metabolic processes, but only because of their acceleration and conse-
quent reduction of the fraction of the energy of .food that remains to be
65
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utilized for growth. Recent studies on salmonid fishes in aquaria and in
artificial streams at Oregon State University have well demonstrated the
importance of this effect. An increase of the activity of fish with rise
of temperature can, of course, sometimes result in greatly increased exploi-
tation of available food resources. However, much improvement of foraging
efficiency may not be possible because of the nature of the food supply and
of the feeding habits of the fish. The effect of a temperature increase on
growth in the natural environment than can be just the opposite of that ob-
served in laboratory tests when the food supply is unlimited. Since the
gross efficiency of conversion of a limited amount of food is reduced at an
elevated temperature at which the appetite for food is increased, there is
superficial similarity between the thermal effects and those of poisons
that impair metabolism while stimulating the appetite. Obviously, however,
there are important physiological differences of these effects that should
be recognized. Growth is more likely to be impaired markedly by a rise of
temperature when uniformly restricted food supplies are small than when the
daily rations are relatively large.
Although deposits of fat can be of great value to fish during periods
of nutritional deficiency, mere deposition of fat should be distinguished
from true growth, which is largely an increase of protein. Appropriate
measurements of body composition should be made, therefore, in connection
with studies of growth. Finally, I want to point out that low concentra-
tions of some poisons may not only be harmless to fish in nature but also
favor growth. There is some evidence that the vigor and foraging
efficiency of fish, as well as their appetite, increase at low levels of
some substances that generally are regarded as poisons only, so that growth
may be promoted not only under artificial conditions. Such data should not
be judged obviously erroneous.
REFERENCES
Brake, L.A. 1972. Influence of dissolved oxygen and temperature on the
growth of juvenile largemouth bass held in artificial ponds. M.S.
Thesis. Orgeon State Univ., Corvallis, Ore.
Broderius, S.J. 1970. Determination of molecular hydrocyanic acid in water
and studies of the chemistry and toxicity to fish of the nickelocyanide
complex. M.S. Thesis. Oregon State Univ., Corvallis, Ore.
Chapman, G.A. 1965. Effects of sub-lethal levels of pentachlorphenol on
the growth and metabolism of a cichlid fish. M.S. Thesis. Oregon State
Univ., Corvail is, Ore.
Davis, G.E., and C.E. Warren. 1968. Estimation of food consumption rates,
ln_ W.E. Ricker (ed.) Methods for assessment of fish production in fresh
waters. Intern. Biol. Prog. Handb. 3. Blackwell Scientific Publica-
tions, Oxford, p. 204-225.
66
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Doudoroff, P- 1969. Discussion of paper by D.I. Mount, "Developing Thermal
Requirements for Freshwater Fishes,". Ir± P. A. Kvenkel and F.L.
Parker (ed.) Biological aspects of thermal pollution. Vanderbilt Univ.
Press, Nashville, Tenn.
Doudoroff, P. and D.L. Shumway. 1970. Dissolved oxygen requirements of
freshwater fishes. FAO Fish. Tech. Pap. 86. Food and Agriculture
Organization of the United Nations, Rome.
Leduc, G. 1966. Some physiological and biochemical responses of fish to
chronic poisoning by cyanide. Ph.D. Thesis. Oregon State Univ.,
Corvallis, Ore.
Lee, R.A. 1969. Bioenergetics of feeding and growth of largemouth bass in
aquaria and ponds. M.S. Thesis. Oregon State Univ., Corvallis, Oreg.
Warren, C.E. (with P. Doudoroff). 1971. Biology and water pollution con-
trol. W.B. Saunders Company, Philadelphia, Pa.
Warren, C.E., P. Doudoroff, and D.L. Shumway. 1973. Development of dis-
solved oxygen criteria for freshwater fish. U.S. Environmental Pro-
tection Agency, Washington, D.C. Ecol. Res. Ser. EPA-R3-73-019.
Warren, C.E., W.K. Seim, R.O. Blosser, A.L. Caron, and E.L. Owens. 1974.
Effect of kraft effluent on the growth and production of salmonid fish.
TAPPI 57: 127-132.
Warren, C.E., J.H. Wales, G.E. Davis, and P. Doudoroff. 1964. Trout pro-
duction in an experimental stream enriched with sucrose. J. Wild!.
Manage. 28: 617-660.
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SECTION 7
MONITORING THE CONDITION OF FLOWING WATERS BY
BIOLOGICAL ORGANISMS
Ruth Patrick
Whether one is concerned with the assimilative capacity of the river or
its sport and commercial fisheries, one eventually has to be concerned with
the whole ecosystem. In monitoring we may study one or a few selected
species or we may assay the condition of the whole aquatic ecosystem in
certain areas. It is important that natural streams be set apart and kept
in their natural state, so we can have base lines against which we can
measure natural change and changes due to man's effects.
In monitoring one may wish to learn about immediate change or more long-
term, subtle changes. In such studies one must remember that time is a
relative parameter. Organisms that reproduce once a day would show effects
in two days that might take years to show in an organism that reproduces
much more slowly. One type of monitoring the condition of organisms is by
bioassay tests. The time and duration of the test to show acute or
subacute effects depend considerably on the organism under study. However,
generally we use bioassay tests of a few hours or a few days to determine
acute effects—that is, those effects that show up immediately. In the
United States these short-term tests are designed to determine the concen-
tration at which 50 percent of the organisms die in a given length of
time. They are, if batch tests, often considered as more-or-less "rough
and ready" tests to get an idea of the effects of a given substance on
aquatic organisms in an ecosystem. In carrying out such tests it is ad-
visable to use organisms representing various stages of the food web, be-
cause the food web may be altered if any stage of nutrient and energy trans-
fer is impaired. For this reason an alga that is a good source of food, an
invertebrate, and a fish are often tested.
Long-term monitoring tests are aimed at showing sublethal effects and
often follow acute tests, because by the acute tests one has found that con-
centration or range of concentrations that probably will not kill the or-
ganism being studied. Whereas acute tests are concerned with
concentrations that cause death; sudden morphological changes such as the
sloughing of mucus by fish; or avoidance reactions to low oxygen; long-
term tests are more concerned with physiological changes and changes re-
lating to the fecundity of the organism.
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In such tests one may examine histological changes in gills, liver, and
pancreas; physiological changes in respiration rates; or enzymatic
changes. For example, Hinton, et al. (1973) have found that DDT inhibits
the formation of ATP—that is, adenosine triphosphatase. Other enzymes
often studied as to the effects of a given waste are d-hydrogenase, acid
phosphatase, and carbonic anhydrase.
In such long-term or chronic tests one often measures the build-up or
accumulation of materials within the cells such as heavy metals, radio-
active materials, or some of the chlorinated hydrocarbons. Behavioral
tests are often used in these long-term chronic tests. For example, Cairns
and Scheier (1964) found the dieldrin will interfere with the sight of
certain fish at extremely low concentrations, thus they are not able to see
their food as well nor, in the case of schooling fish, are they able to
school. One also is concerned about the effect upon the fecundity of
organisms. Shifts in temperature or chemicals that alter the food species
may affect the fecundity of the female and success of offspring. If the
fecundity of a species is changed very much it may alter the whole food web.
Recent work of Patrick et al. (1975) has shown that minute amounts of
heavy metals such as nickel, vanadium, and chromium may alter the species
composition of the algae in a community and thus greatly change the plant
food source of the food web. If such changes occur and the primary produc-
tion is carried out by species of low food value, the productivity of the
rest of the food chain will be greatly reduced.
Monitoring may be concerned with changes in individual species in the
waters in which they live or with changes in communities. If one is
studying one or a few species, such studies are usually carried out by
isolating the species under study either in the field or in semi natural
conditions. For example, oysters or clams are often sorted as to size and
age class and placed into large trays with each oyster or clam being
marked; thus over time one can study growth, the attack by disease, and the
condition of the oyster or clam in question by sacrificing the individual
organism (Figures 1,2). One can also determine the accumulation of heavy
metals, or radioactive materials, or of chlorinated or polycyclic hydro-
carbons. Thus such studies are valuable not only to monitor the potential
of the commercial crop in the area but also to monitor whether or not cer-
tain toxicants have passed through the estuary.
Fish in aquaria are sometimes used to monitor the effect of a given
waste as it is being discharged. Cairns et al. (1973) have described such
a methodology- Fish can be sacrificed from time to time to discover
histological and physiological changes as well as the observations from day
to day of death. In such studies of fish or oysters the same organisms may
be studied over time and thus the monitoring is continual, although observa-
tion may not be continual. More recently, Dr. Burton of the Academy labora-
tories has developed methods of inserting probes into crabs and thus being
able to continually monitor the heartbeat and various kinds of biochemical
changes within the crab.
69
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Figure 1. Platform from which oyster trays are suspended.
70
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^^^^^^NV^S'/X^S^^
Figure 2. Oyster tray.
71
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Sometimes a given species such as oysters in an oyster bed is intermit-
tently monitored by taking grab samples; however, if this methodology is
used one must very carefully design the experiment so that the total number
of grabs will give reproducible results—in other words, that there will be
given degree of statistical reliability that if the procedure is repeated
the same kind of data will be obtained if no change occurs. This type of
monitoring has been developed by the laboratories of the Academy of Natural
Sciences (Patrick, in press).
Another type of monitoring is that which has been developed for monitor-
ing communities of organisms. The most sophisticated of these has been
those developed by Patrick et al. (1954) for algal communities growing on
glass slides. In this method an apparatus known as a diatometer is intro-
duced into a body of water. It has been found that diatoms grow success-
fully on these slides. This fact was first pointed out by Butcher (1947).
However, the method used by Patrick is the first one to model the community
and to note by changes in the structure of the community as well as in the
kinds of species the effects of a pollutant. For example, it has been
found that under natural conditions the structure of the diatom community
conforms to a truncated normal curve (Figure 3) and that this curve remains
fairly constant over time (Table 1). However, if pollution high in
nutrients is introduced such as those high in nitrogen, phosphorus, and
carbon, certain species will become extremely common and produce a long
tail to the curve (Figure 4). Under toxic conditions typically one finds a
reduction in numbers of species and the sizes of populations, although in
some cases a few species that can withstand or tolerate the toxicant become
very common because there is little competition by other species for the
nutrients in the system and predator pressure has been greatly reduced
(Figure 5).
By this diatometer method of studying algal communities one cannot only
study shifts in the diatom community, but can determine whether or not
shifts from diatoms to other species are occurring.
These diatometers can also be inserted into various reaches of a river
to determine the relative degrees of eutrophication of the areas by the
total biomass and kinds of species produced on the slides. Thus they are
valuable in regional studies of eutrophication. In some instances they
have been found extremely useful in determining the presence of small
amounts of heavy metals or radioactive materials because some metals are
concentrated by the algae to amounts many thousands of times the concentra-
tion of the ambient medium. Algae growing on these slides can also be used
in determining primary productivity and P/R ratios of algal communities.
We have found that these measures are important in determining small or
sublethal shifts in the community. Likewise, one can extract pigments
from them and determine a shift in pigment concentrations.
Periphyton are particularly good organisms to study because they have
short life cycles and often produce chronic effects due to low amounts of
toxicants much more rapidly than many larger macroinvertebrates. Further-
more, we know a considerable amount about the kinds of species and what
they indicate. For example, the diatoms Ni'tzschia pa lea and Gomphonema
72
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40 p
GO
V)
W
I— I
u
w
&
w
pa
35
30
25
20
B 15
10
5
0
INDIVIDUALS =1-2 2-4 4-8
INTERVALS =0 1 2
8-16 16-3232-64 64- 128- 256- 512- 1024-2048-4096-8192-16384-32768-65536-
128 256 512 1024 2048 4096 8192163843276865536131072
10
11
12
13
14
15
16
17
Figure 3. The structure of a natural diatom community.
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TABLE 1. SUMMARY OF CATHERWOOD DIATOMETER READINGS AT STATION 1
OCTOBER 1953 TO JANUARY 1958
Date
Oct. 1953
Jan. 1954
Apr. 1954
July 1954
Oct. 1954
Jan. 1955
Apr. 1955
July 1955
Oct. 1955
Jan. 1956
Apr. 1956
July 1956
Oct. 1956
Jan. 1957
Apr. 1957
July 1957
Oct. 1957
Jan. 1958
Specimen number
of modal interval
4-8
4-8
2-4
2-4
4-8
4-8
2-4
2-4
2-4
2-4
4-8
2-4
2-4
2-4
2-4
4-8
2-4
2-4
(Apr. 1954-1958
averages)
Species in
mode
22
19
24
23
21
19
25
20
27
30
35
24
23
29
21
29
25
27 '
24
Species ,
observed
150
151
169
153
142
132
165
132
171
185
215
147
149
177
132
181
157
152
151
Species in theo-
retical universe
178
181
200
193
168
166
221
180
253
229
252
185
206
233
185
203
232
212
194
74
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to
W
I— I
O
w
w
M
40 r
35
30
25
20
15
—I
en
10
5
0
INDIVIDUALS = 1-2 2-4 4-8 8-16 16-3232-64 64- 128- 256- 512- 1024-2048-4096-8192-16384-32768-65536-
128 256 512 1024 2048 1096 8192163843276865536131072
INTERVALS — 0
8
10
11
12
13
14
15
16
17
Figure 4. The structure of a diatom community under the effects of pollution high
in nutrients.
-------
OT
w
t-t
u
w
a,
OT
w
w
40
35
30
25
20
15
10
CTl
0
T
INDIVIDUALS = 1-2 2-4 4-8 8-16 16-3232-64 64- 128- 256- 512- 1024-2048-4096-8192-16384-32768-65536-
128 256 512 1024 2048 4096 8192163843276865536131072
INTERVALS =0 1 2 3 4 5 G 7 8 9 10 11 12 13 14 15 16 17
Figure 5. The structure of a diatom community under the effects of toxic conditions.
-------
parvulum typically develop large populations under nutrient-rich conditions.
The presence in abundance of Cyclotella meneghiniana in contrast to
Cyclotella stelligera and £. kutzingiana indicates an increase in the
nutrient levels of the water. There are a great many diatoms that indicate
these shifts. Studies by Patrick (1956) have shown that diatoms have a
toxicity threshold similar to that of fish and invertebrates to many
industrial wastes, and thus a very short-term test can tell a considerable
amount about the effects of a toxicant in a body of water on other members
of the food web.
Various types of substrates have also been developed for monitoring
invertebrate communities. Whereas the diatometer substrates (Patrick et
al., 1954) have been devised to reliably represent the community in the
river under study, this has not been done so far as I know for inverte-
brates. However, various people have studied how many invertebrate
samplers one needs in an area in order to get reproducible results (Beak,
et al., 1973).
One type of substrate that has been used are panels which collect
sessile organisms (Figure 6). This is a substrate sampler consisting of a
series of flat substrates placed in the water. This has been found to be
very good for the collection of certain invertebrates. Other people have
used various kinds of baskets from simple chicken-wire baskets filled with
rotted wood to baskets of very definite structure such as barbeque
baskets. These have been found very useful, but the important thing is to
calibrate them so that one can obtain reproducible results and know the
extent these organisms represent the area under study. In this way one can
compare over time changes in a given area and the shifts between areas.
In the macroinvertebrate studies, as in periphyton or diatom studies,
one is concerned about shifts in the dominant forms or shifts in the sizes
of populations; shifts in kinds of species; and shifts in numbers of
species. For example, by insect traps placed by Dr. Roback of the Academy
in the Savannah River he was able to clearly define the effects of
dredging, because the filter-feeders such as caddisflies disappeared as
long as the water had a high suspended solids load. One can also determine
different degrees of nutrient loading in the water by shifts in the faunas
of insects such as shifts in a mayfly-stonefly dominated fauna to one domi-
nated by damselflies and dragonflies, to one dominated by chironomids and
worms. Since mayflies and stoneflies are particularly sensitive to oxygen
concentrations, the loss of these species indicates an oxygen sag at times
in the river even though it is not noted chemically. In other cases shifts
of Hydropsyche caddisflies to dragonflies and Chematopsyche caddisflies
have indicated intermittent toxicity.
A second type of monitoring is one in which given selected areas of a
river are studied over time. This type of monitoring is extremely valuable
because it tells a great deal more than monitoring by means of substrates.
In such monitoring a team of scientists is sent into a river. They deter-
mine not only changes in the chemical and physical characteristics but also
the characteristics of the aquatic ecosystems. They are able to determine
by increased growth of various types of organisms -such as submerged or emer-
77
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cinder block
Figure 6. Invertebrate sampler.
78
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gent aquatic plants if the nutrient content of the water and sediments are
high. Beds of Sphaerotilus indicate a significant increase in carbona-
ceaous materials in the water. Furthermore, such studies are able to deter-
mine shifts in the diversity of habitats or physical changes in the river
channel. This information is not determined by monitoring with substrates,
for by the use of substrates one simply determines facts concerning the
organisms, but not the condition of the area in which they live.
Another difference is that when one sends a team of scientists into a
body of water one determines facts concerning many different groups of
organisms. This is extremely important if one wants to determine subtle
changes. Often an ecological change that has nothing to do with pollution
will affect a single group of organisms, but it is extremely rare that an
ecological change will affect many groups of organisms such as the algae;
invertebrates such as molluscs and worms; insects; and fish. This
methodology was first developed by my staff at the Academy of Natural
Sciences in 1948. Therefore, the more lines of evidence from different
kinds of organisms that one has the more sure one is of his diagnosis of
conditions. Such studies are particularly valuable when one is concerned
with small sublethal effects which are important to detect before they
become problems. Wth the possible problems in our country due to increases
in chlorinated hydrocarbons, heavy metals, and radioactivity, this type of
thorough examination of conditions in monitoring becomes more important.
Of course, this type of monitoring is intermittent, and it is more expen-
sive, and therefore between studies things may occur that one does not
realize. For this reason it is best to combine a continuous monitoring
system with this more thorough, intermittent system. I often compare these
types of studies to medical treatment. If one wants to know if something
is wrong a simple procedure can be used such as examining the condition of
a single group of organisms. This compares with taking one's temperature
or doing a cardiogram. But if one wants to understand trends or causes of
change a thorough study of aquatic areas or a detailed physical examination
of an individual is needed.
As noted above, the kinds of changes which one observes are first the
changes in relative sizes of populations of species. If we find that those
species that are tolerant to a given type of pollution are becoming more
common, then one strongly suspects that it is present. For example, in a
stream in eastern United States if one finds a shift from mayflies and
stoneflies and certain species of caddisflies being very common to a great
increase in chironomids, dragonflies, Physa snails, the limpid, Ferrissia,
and tubificid worms, we know that the nutrient load of the river has in-
creased. Furthermore, if we find that only organisms such as the flatworm
Dugesia tigrina, Ferrissia tarda, Physa heterostropha, and tubificid worms
are present we know that the degradation is caused by increased nutrients
and resultant increase in bacteria, and probably no toxicity. However, if
we find and increase in certain of the chironomids and of certain dragonfly
larvae without an increase in the above mentioned species, we can infer
that the organic load may occasionally have low levels of toxicity present
(Patrick observations).
79
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Temperature is another common pollutant that is often difficult to
detect unless continual monitoring is carried out, and then it is sometimes
difficult to predict the temperature regime in the river. However, shifts
in algal species will clearly denote these kinds of changes. For example,
we have found that if the temperature of the water consistently remains
below 30 C and no other pollutant is prevalent, diatoms will be dominant
throughout the year in most streams, particularly in eastern United
States. If the temperature during the summer increases to between 30 C and
33 C green algae will predominate, and some blue-green algae will become
very common. Thus one can estimate the temperature regimes in various
parts of a body of water.
Another type of determining change, particularly in lakes, has been to
examine the fossil record. In such studies sediment cores are taken and
dated. The shift in diatoms and invertebrates species enables one to deter-
mine trends toward eutrophication.
From this discussion it is evident that the monitoring of biological
organisms can be very valuable in determining the effects of wastes. As
contrasted with chemical and physical determinations of water quality, the
organisms integrate over time all deleterious effects, whereas a chemical
examination only determines the presence of the chemical for which analysis
is made at the particular time. Actually it is important that both types
of studies be made. The biological studies often give an indication of a
certain type of chemical or deleterious conditions being present. It is
then necessary to determine exactly what chemical is causing the effect.
Therefore, both types of studies become important, but the biological
studies are the better continual monitoring studies if only one type of
monitoring is to be made, because it integrates all changes which may occur.
In the United States we are also realizing the importance of preserving
specimens from monitoring studies. Recently there was a considerable scare
about the accumulation of mercury in fish. However, an examination of fish
in our museum collections showed that these older specimens had similar
amounts of mercury and the sudden awareness of the presence of mercury was
due to better analytical techniques. Without these specimens in our
museums such comparisons would not have been possible.
From these various examples it is evident that biological monitoring
can be useful to determine the extent and degree of harm in an area of a
specific waste. Diatometers and similar samplers can be very useful in
determining trends over long reaches in a river system of increases in
pollution. Organisms such as diatoms and some invertebrates can pick up
and concentrate over time amounts of chemicals infrequently discharged that
would probably not be picked up by ordinary chemical monitoring.
Continuous monitoring as well as intermittent monitoring is extremely
useful in comparing changes over long periods of time in various bodies of
water.
The program of monitoring depends on the questions one wishes to ask,
particularly whether one wants to determine general changes or whether one
wants to determine trends or more precise causes of change.
80
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REFERENCES
Beak, T.W., T.C. Griffing, and A.6. Appleby. 1973. Use of artificial sub-
strate samplers to assess water pollution. Ir± J. Cairns and K. Dickson
(eds.) Biolgical Methods for the Assessment of Water Quality. American
Society for Testing and Materials, Philadelphia, Pa. STP 528. p. 227-
241.
Butcher, R.W. 1947. Studies in the ecology of rivers. VII. The algae of
organically enriched waters. Ecol. 35(1/2): 186-191.
Cairns, J. Jr., and A. Scheier. 1964. The effect upon the pumpkinseed sun-
fish Lepomis gibbosus (Linn.) of chronic exposure to lethal and sub-
lethal concentrations of dieldrin. Notulae Naturae, Academy of Natural
Sciences, Philadelphia, Pa. No. 370. 10 pp.
Cairns, J. Jr., R.E. Sparks, and W.T. Waller. 1973. A tentative proposal
for a rapid in-plant monitoring system. J_n J. Cairns and K. Dickson
(eds.) Biological Methods for the Assessment of Water Quality. Ameri-
can Society for Testing and Materials, Philadelphia, Pa. STP 528.
p. 127-147.
Hinton, E.E., W.M. Kendall, and B.B. Silver. 1973. Use of histologic and
histochemical assessments in the prognosis of the effects of aquatic
pollutants. Jjl J. Cairns and K. Dickson (eds.) Biological Methods for
the Assessment of Water Quality. American Society for Testing and
Materials, Philadelphia, Pa. STP 528. p. 194-208.
Patrick, R., M.H. Hohn, J.H. Wallace. 1954. A new method for determining
the pattern of the diatom flora. Notulae Naturae, Academy of Natural
Sciences, Philadelphia, Pa. No. 259. p. 12.
Patrick, R. 1956. Diatoms as indicators of changes in environmental condi-
tions. Jjl C.M. Tarzwell (ed.) Biological Problems in Water Pollution.
R.A. Taft Sanitary Engineering Centrer, Cincinnati, Ohio. p. 71-83.
Patrick, R., T. Bott, and R.A. Larson. 1975. The role of trace elements in
management o'f nuisance growths. U.S. Environmental Protection Agency,
Corvallis, Ore. Environmental Protection Technical Series, EPA-660/2-
75-008.
81
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SECTION 8
THE ROLE OF ALGAE IN THE POLLUTION OF RESERVOIRS
AND PROBLEMS OF CONTROLLING THEIR NUMBERS
1
V.G. Khobot'ev
Water pollution, both marine and freshwater, from industrial wastes
changes the living conditions for all living organisms and disturbs the es-
tablished communities. In some polluted waters, with nutrient enrichment
and slightly increased temperature, favorable conditions are created for
massive development of algae. Algal cell counts reach millions or even
billions in a single liter.
The massive development of phytoplankton creates a considerable nui-
sance in water supplies, since it often disturbs treatment processes and im-
pairs the quality of water produced. Algal blooms in water bodies of this
type promote the intensified growth in underground mains and equipment, com-
plicating treatment and sometimes causing equipment failure.
In cooling ponds blooms facilitate the formation of thick, compact sur-
face films that hinder evaporation and heat loss from the surface, thereby
reducing normal cooling of waste water. The growth of algae also reduced
C02 through intense photosynthesis and produces scum on the inner surfaces
of heat-exchange equipment. Removal of this scum frequently requires a con-
siderable expenditure of labor, time, and resources. Blooms caused by blue-
green algae also considerably degrade drinking water quality, giving the
water an unpleasant taste and odor. The taste is caused by the emission of
sulphur-containing compounds produced by the blue-green algae. Methymercap-
tan, dimethylmercaptan, isopropylmercaptan, dimethyl sulphide, and others
have been identified in decaying cultures of these algae. The odor of dime-
thyl sulphide, brought about by the presence of such amines as methyl amine
and ethylamine, strongly resembles the smell of fish. A similar smell in
natural water associated with the development of certain species of algae
is caused by dimethylsulphide.
The massive development of algae creates difficulties in the operation
of other water plants as well as in canals and irrigation systems. For con-
trol of phytoplankton, different means have been widely and effectively
used in many cases. Biological, physical, mechanical and chemical methods
of controlling "blooms" in reservoirs are well-known.
'Moscow State University
82
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The biological method uses biological filtration. One method is that
proposed by S.N. Skadovskiy--a method in which a cascade arrangement of
selected aquatic organisms in the water purifies it. In the upper portion
of the cascade large sections of the bottom are populated with filter
feeders such as freshwater mollusks (Unionids and Anodonts), which are cap-
able of filtering up to 2000 liters of water per day at a distribution
density of up to 70 individuals per square meter. These are followed by
water plants, which reduce the dissolved nitrogen in the water to a minimum.
Biological communities growing on surfaces as overgrowths will further re-
duce the nuisance organisms. Such communities reduce the number of phyto-
plankton cells by 60-70%, saprophytic bacteria by 70-80% and intestinal
bacilli by 30-50%.
The mechanical method is based on various filter systems through which
the water is filtered and suspended particles are removed. During heavy
bloom, however, the filters quickly plug and they must be cleaned, sometimes
every 20-30 minutes. The efficiency in such filters fluctuates from 20-60%.
The physical method of combatting bloom is based on the destruction of
algal cells using ultrasonics or an electric current. The shortcoming of
this method is the need for additional equipment to remove the slimy mass
that is obtained.
The chemical method is most used for preventing blooms. To date, hun-
dreds of chemical compounds have been tested as algicides to suppress the
development of algae. Copper sulphate and chlorine are utilized most fre-
quently. Their effectiveness, however, depends on the acidity of the en-
vironment. The copper ion in copper sulphate is toxic for algae, therefore,
the effect of this ion depends on the concentration of hydrogen ions, i.e.,
the lower the pH, the more toxic the ion becomes. Since pH is greater than
7 in the majority of water bodies, the effectiveness of this agent is con-
siderably reduced. Active complexing of copper ions with ligands in indus-
trial water supplies also often leads to a reduction in the toxicity of
copper sulphate.
The chlorine in hypochlorous acid is toxic. Because of its strong oxi-
dizing efect, chlorine penetrates into plant cells and damages vital
centers. Chlorine concentration also depends on the pH of the solution and
it increases only with a reduced pH.
Besides pH, the selection of algicides depends on the capacity of the
algae to adapt to the effect of the preparations. If only 2-3 species of
the 50-60 algae species encountered in a water body adapt, then difficulties
can occur for the industrial utilization of the water since these species
can cause a bloom and occur in huge numbers. The selection of different
chemical substances which are toxic to these specific organisms can be the
best solution. A collection of 2-3 algicides provides the possibility of
completely suppressing blooms in the water. However, in selecting algi-
cides, one must consider the features of the water body being treated. The
substance should possess toxicity with regard to the greatest number of
algal species, have the capacity to penetrate easily into vitally important
83
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centers of the organism, not react with chemical compounds contained in the
water and be available for use. In addition, the algicides should be safe
for man, harmless to fish and non-corrosive to metallic parts of equipment.
Chemicals such as 2,3-dichloronaphthoquinone (also called figone, frigone,
2,3DNA, 2,3SNA) and hexachlorobutadiene are suggested as algicides, as well
as rosein-amide-D-acetate, monuron, simazine and many other algicides.
The physiological effect of some algicides, monuron and diuron on blue-
green algae in particular, amounts to an acute inhibition of photosynthesis.
Diuron, the molecules of which contain two chlorine atoms and methyl groups
in addition to the phenol ring, possess the most clearly-pronounced effect
on blue-green algae. However, the introduction of these algicides into a
reservoir impairs the organoleptic properties of the water, and the pro-
cesses of nitri- and nitrofication are disturbed. Through toxicity tests
on other aquatic organisms 2 and 10 mg/liter were shown-to have a substan-
tial effect on blood morphology, phagocytic activity of leucocytes and
other changes in test animals.
Many algicides are volatile or readily hydrolyzed, thereby requiring re-
peated application, sometimes 3-4 times in a summer.
Searchers for new compounds to effectively protect the water from blooms
led us to study the effect of complex ores and products of their processing
on various species of algae. Together with A.P- Terent'ev and N.S.
Stroganov, we established that complex ores containing zinc, copper, cad-
mium, nickel, lead, silver, and other chemical elements act effectively on
the algae which produce blooms in water bodies. In tests 2 mg/liter reduced
the number of algae cells in 30 days and eliminated them in 45 days. Under
the effect of complex ores, the filamentous alga, Cladophora, decreased
sharply in biomass and the colony darkened and decomposed into separate
cells. Green algae proved more resistant to the effect of ores, but was
severely depressed. First the chlorophyll disappears, then the cells lose
pigment and in 30 days they are almost entirely dead.
The effect of complex ores is a complex process, but the investigations
conducted show that the slow change of complex metals into a soluble state
and their low initial concentration evidently favor their inclusion into
the algae's biochemical reactions and lead to inhibition of various vital
functions. The slow and weak dissolution of the ores is easily overcome if
the ores are introduced into the water body 10-15 days prior to the begin-
ning of phytoplankton development. In some cases, it is convenient to
introduce the algicide into the ice in the spring so that water contains a
solution of substances from the complex ore in the required concentration
prior to the spring outbreak of algae development. Thus, one can prevent
an increase in the number of algae by controlling their numbers, and not
just decreasing them. Complex ores, usable as algicides, are resistant to
hydrolysis and not susceptible to destruction by bacteria. Their slow dis-
solution in water makes it possible to maintain a concentration in the re-
servoir which is toxic for algae for a long period of time.
84
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For the purpose of controlling the quantity of plankton organisms in in-
dustrial water supplies, we tested another c lass--organostannic compounds
and trialkyl/aryl/substituted compounds in particular, which possess fun-
gicidal and insecticidal properties. We tested trimethylstannanol and trim-
ethyl acetoxy-stannane which strongly suppressed the vital functions of
algae even at a concentration of Q.2 mg/liter. In the same concentrations,
these compounds suppressed the reproduction process in representatives of
zooplankton. The advantage of organostannic compounds over other algicides
is that they are toxic for plankton organisms at considerably lower concen-
trations.
However, it is important to use substances that are selectively toxic
to the undesirable algae. Often blooms must be suppressed without harming
the development of zooplankton and other aquatic organisms. Phytoplankton
toxicity tests with silicone compounds (vinyl-triethoxysilane, tetraethoxy-
silane, trifluoropropinyldiethoxysilane, and others) showed that 0.01
mg/liter depresses the number of algae by roughly 94-96% in the first days
of the test; however, in the subsequent 15 days the number of cells in-
creases slightly and reaches 50% of the control. Tests on Daphnia and
Mollusca showed 100% survival even at 1 mg/liter of silicone compounds.
The number of young borne by Daphnia in the test proved to be greater than
in the control.
By comparison of the sensitivity of representatives of zoo- and phyto-
plankton to silicone compounds, the high selectivity of phytoplankton is re-
vealed, as well as low sensitivity or stimulation of Daphnia. The ability
of silicone compounds to reduce algae cells can be utilized to suppress the
development of phytoplankton during blooms.
Alkoxysilanes have the advantage over previously-used algicides to com-
bat blooms because they are comparatively quickly destroyed in water and
have a selective effect on phytoplankton, while not suppressing the vital
activity of Daphnia and Mollusca.
The development and utilization of toxic substances, which possess a
narrow selective effect, opens a path towards the synthesis of substances
with prescribed toxicity for certain aquatic organisms and is the basis for
the development of the best methods of controlling the number of zoo- and
phytoplankton in reservoirs.
No one method is best and there should be several such methods for spe-
cific purposes for each water body.
REFERENCES
Terent'ev, A.P., Stroganov, N.S., Rukhadze, Ye. G., Khobot'ev, V.S. DAN
SSR, 164(4) (1965).
Stroganov, N.S., Kochkin, D.A. Khobot'ev, V.G., Kolosova, L.V. DAN SSSR,
170(5) (1966).
85
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Stroganov, N.S., Khobot'ev V.6. Priroda, (12) (1966).
Stroganov, N.S., Khobot'ev, V.G., Kolosova, L.V., Kadina, M.A. DAN SSSR,
181(5) (1968).
86
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SECTION 9
EUTROPHICATION IN THE UNITED STATES: PAST-PRESENT-FUTURE
A.F. Bartsch, K.W. Malueg, C.F- Powers, and T.E. Maloney
Chief Seattle, leader of the Suquamish tribe in the Washington Terri-
tory, delivered a prophetic speech in 1854. The occasion was to mark the
transferral of ancestral Indian lands to the federal government. His words
indicate much greater understanding of man's position in the natural system
than we seem to have today, for he stated:
This shining water that moves in the streams and rivers
is not just water but the blood of our ancestors. If we sell
you land, you must remember that it is sacred, and that each
ghostly reflection in the clear water of the lakes tells of
events and memories in the life of my people. The water's
murmur is the voice of my father's father.
The rivers are our brothers, they quench our thirst. The
rivers carry our canoes, and feed our children. If we sell you
our land you must remember, and teach your children, that the
rivers are our brothers, and yours, and you must henceforth
give the rivers the kindness you would give any brother.
...The earth does not belong to man; man belongs to the
earth.
In the United States we have failed for too long to teach our children
that the waters are sacred. Our rivers still carry our canoes, our boats,
our ships, but very few of them could feed our children. Most of our lakes
are not clear. Many offer no lovely reflection to a potentially admiring
glance because most of the time, they are covered with algae or higher
aquatic plants.
As elsewhere throughout the world, eutrophication is now a familiar
problem in the United States. This progressive nutrient enrichment of
lakes and their responding increased biological production with its related
consequences are the major threat to many lakes. Lakes typically evolve
from a state of low productivity and relative high purity to one of in-
creased productivity and lessened quality. This process often is marked by
nuisance algae or other plant growths, drastically reduced oxygen content
in the deeper waters, and bad tastes and odors. Reaching this stage
usually is a lengthy process, sometimes even requiring thousands of years.
87
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However, when a lake is subjected to heavy pollution and other forms of
human population pressure, eutrophication proceeds much more rapidly. This
accelerated process has been termed "cultural" eutrophication.
THE PAST (1850-1967)
Cultural eutrophication was recognized as a problem in this country at
least as early as 1850 when complaints about unpleasant odors from Lake
Monona that assailed the citizens at Madison, Wisconsin, were published in
local newspapers. Eutrophication research in this country appears to have
started in the early 1900's with the work of Birge and Ouday at the Univer-
sity of Wisconsin. Because eutrophication is fundamentally an expression
of the metabolism of lakes, studies of limnology and eutrophication go hand
in hand.
In these early years limnology centered around the taxonomy, and to a
lesser degree, the ecology of the zooplankton and around descriptive
investigations of lake phenomena. Included particularly were the areal and
seasonal distributions of temperature, dissolved gases, and solar radiation.
Increased interest in the chemistry of lake waters developed during the
second quarter-century. There was special interest in nutrients, pH, Eh,
organic matter, and oxygen consumed, all parameters related directly to
lake productivity and trophic state, and hence eutrophication.
In the 1930's and 1940's more attention was given to cycling of nutri-
ents in lakes. The awareness, for example, that algal populations could be
maintained or increased with no apparent changes in lake-water concentra-
tions of available phosphorus, or in the presence of no detectable avail-
able phosphorus (or nitrogen), led to renewed investigations in these areas.
It then became evident that in many lakes the nutrient elements were under-
going very rapid cyclical changes, moving between bottom sediments and over-
lying water, from dead organic matter to the water, from the water to
actively photosynthesizing plants and to bacteria. These precepts are
fundamental to an understanding of the eutrophication process and are basic
to the concept of limiting nutrients.
Following World War II increased attention was devoted to the accele-
rated eutrophication of lakes by nutrients from cultural sources. For
example, experience with the Madison, Wisconsin, lakes, Lake Washington at
Seattle, and a number of European lakes (particularly Zurichsee) made it
increasingly evident that nutrients introduced by numerous activities of
man—from both point and non-point sources—could lead to rapid and serious
deterioration of water quality. As a consequences, a number of remedial
approaches were taken to curb nutrient contributions to lakes. At Madison,
where municipal sewage had long been discharged into the chain of lakes,
the following diversions took place: (a) from Lake Mendota in 1899, (b)
from Lake Monona in 1936, and (c) from Lake Waubesa in 1958. At Seattle's
Lake Washington, effluents from 11 treatment plants were diverted to Puget
Sound between 1963 and 1968. After the first diversion the lake's condi-
tion began to improve and has continued to do so. The abundance of phyto-
-------
plankton has decreased. Secchi disc transparency, which had fallen from
3.7 in 19bO to 1.0 in 1963, returned to 3.5 m in 1971. The lake is now no
longer considered eutrophic by many investigators.
In the 1950's sufficient evidence had already come to hand to strongly
suggest that even huge bodies of fresh water such as the Great Lakes are
not immune to cultural eutrophication. Lake Erie, the shallowest and most
polluted of the five, was the first to exhibit serious deterioration. Sub-
sequently, similar changes in lake water quality were detected in southern
Lake Michigan, Lake Ontario, and other areas of the Great Lakes. Although
these lakes had been studied for many years from the point of view of
fisheries management, very little limnological work had been done prior to
the fifties. These evident eutrophication trends, however, served to
stimulated greatly increased research efforts at a number of United States
and Canadian universities, and by the governments of the two countries as
well. Great strides have since been made in furthering the knowledge of
these important lakes, leading to more intelligent management and utiliza-
tion of the resource.
Also in the 1950's, and indeed as far back as the 1940's, emphasis
began to shift from descriptive to experimental limnology, in which mani-
pulations are carried out on the full-scale or pilot-scale level, utilizing
lakes, ponds, or physical models. Such experimental methodology has become
fundamental to the development of lake-restoration techniques and proce-
dures, wherein results from the laboratory are carried to the field for
testing under controlled conditions.
Studies of the nutrition of algae, both freshwater and marine, have
been going on since the 18th century. However, progress was slow compared
to other branches of biology, probably because of difficulties in culture
and manipulation. To define the nutritional requirements of algae, they
must be obtained in pure culture. As a consequence, progress in these
studies had to await the development of improved laboratory equipment that
would make this possible. As a result, during the 19th century algae were
studied almost exclusively from the morphological and taxonomic points of
view. By 1920 many species had been isolated in a bacteria-free state thus
setting the stage for investigations under controlled conditions. Today
the nutritional requirements of many species of algae are well documented.
Through the first half of the century limnology was more or less
centered in the midwestern universities. Since World War II, however,
teaching and reseach programs in limnology, with related centers of
excellence in eutrophication, have been established in all the major geo-
graphic areas of the country. As a result, eutrophication problems pecu-
liar to various regions can now be considered much more thoroughly than
ever before. We hope these scientific resources will facilitate in many
ways the control of this problem through sound management decisions. One
such way is the continued synthesis of available information that bears on
eutrophication and control possibilities.
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An excellent beginning at such a synthesis occurred at the Inter-
national Symposium on Eutrophication sponsored by the National Academy of
Sciences and held on the campus of the University of Wisconsin in 1967.
Proceedings of this symposium were published under the title "Eutrophica-
tion: Causes, Consequences, Correctives" (National Academy of Sciences,
1969). This went far to bring together much of the existing knowledge of
eutrophication processes and controls. A second review document, "Eutro-
phication—A Review," was published in 1967 (Stewart and Rohlich, 1967).
Together, these two publications emphasized the state of the art, pointing
the way for future work. Their appearance at this critical time marks the
year 1967 as an especially significant milestone in eutrophication and lake-
restoration research.
THE LAST EIGHT YEARS
The period from 1850 to 1967 provides the historical preclude to the
modern reaction to the eutrophication problem in the United State. No
doubt, the manpower and dollars spent in the past 8 years to understand and
cope with eutrophication far exceeds those spent in all of the preceding
117 years. The outlook today is that expenditure to activate remedial
technology soon will be greater than dollars spent to develop new techni-
ques.
Two factors lead to this conclusion: (a) several remedial approaches
are available, and (b) the number of lakes to be addressed is very great.
In the United States, excluding Alaska, there are about 100,000 lakes. At
best, this is only an estimate because no standard definition of a lake is
used by all states1. It is estimated, also, that 12,000-15,000 lakes are
over 4.5 ha and that 10-20% are eutrophic. Generally, lakes in or near
urban development are eutrophic, whereas many of those in relatively
sparsely populated areas are more likely to be oligotrophic.
Today, one can still ask: What causes eutrophication? Many inter-
acting factors contribute to the overall process. Productivity depends on
solar radiation, temperature, lake-basin morphology, water-retention time,
and perhaps most important, the availability of adequate nutrients. It is
generally agreed that algae and higher aquatic plants require 25-30 differ-
ent nutrients for growth. Large amounts of carbon, nitrogen, hydrogen, and
phosphorus and smaller amounts of approximately 25 others, such as magne-
sium, calcium, boron, zinc, copper, molybdenum, and manganese, are
necessary. In addition, vitamins such as B12, thiamine, and biotin, and
hormones play a part in nutrition. In theory, since all of the above are
essentially for growth, the unavailability of any one could control eutro-
phication. Generally, however, nitrogen and phosphorus emerge as the criti-
cal elements in controlling aquatic plant nuisances.
states report, as lakes, bodies of water over 1.21 ha (3 acres);
Others, over 4,05 ha (10 acres); and others over 40.5 ha (100 acres).
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The Carbon Problem
In the early 1970's a major controversy developed in regard to the rela-
tive importance of carbon, phosphorus, and nitrogen in regulating eutrophi-
cation. This controversy centered on a contention that carbon rather than
phosphorus or nitrogen limits algal productivity in many aquatic ecosystems.
Since phosphorus in detergents is linked to cultural eutrophication of
lakes and streams, the controversy became emotionally charged following pro-
posals to remove phosphorus from detergent formulations. Because of this,
the American Society of Limnology and Oceanography sponsored a special
symposium entitled "Nutrients and Eutrophication: The Limiting-Nutrient
Controversy" (Likens, 1972). This symposium was held in February 1971 in
an effort to provide a clear statement on the relative importance of
various regulating or limiting nutrients in the eutrophication of aquatic
ecosystems. The papers and discussion focused not only on phosphorus,
nitrogen, and carbon, but also considered other nutrients and environmental
factors that affect eutrophication. As various ideas, views, and data were
openly and authoritatively debated, there emerged a general agreement that
phosphorus is the critical limiting nutrient in most North American lakes
and hence should be the center of focus for management programs. This ex-
pression in the published proceedings provides guidelines to the public and
to officials concerned with lake protection.
Nutrient Loading
Preventive and remedial programs based on nutrient-control measures can
be initiated effectively only after the origins of the nutrient supplies
have been determined. To obtain an accurate nutrient budget, all sources
must be considered, including all tributaries, industrial discharges, muni-
cipal discharges such as sewage and storm waters, precipitation, ground
water, unchanneled surface runoff, and last but not least, feedback from
the lake sediments. Synthesis of such a budget for even one nutrient is
time-consuming, difficult, and expensive. Few accurate nitrogen and phos-
phorus budgets exist for U.S. lakes. Of these, the phosphorus budgets are
generally more accurate than those for nitrogen.
Several actions are underway to improve our understanding of the rela-
tionship between lake loading and lake response. One program is the OECD2
North American Project. Approximately 40 scientists from the United States
and Canada are collecting and analyzing limnological data from selected
lakes. Correlations of nutrient loading with mean depths and water-resi-
dence times are being examined to determine relationships of these factors
to the prevailing trophic level in the studied lakes. The results of this
study will be available soon.
o
Organization for Economic Cooperation and Development.
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The U.S. Environmental Protection Agency (EPA) initiated "The National
Lake Survey" in the summer of 1972. This project determines the location,
severity, and extent of eutrophication in lakes and impoundments that act
as receiving waters for municipal waste-treatment-plant effluents. At a
cost of $7-8 million and a time span of at least 4 years, the study will
examine 812 major lakes and impoundments in 48 states. For each lake the
trophic state is estimated, and the sources and magnitudes of nitrogen and
phosphorus supplies are identified, to judge whether reduction in phos-
phorus loading will restore or protect the lake.
To sample the lakes and measure limnological characteristics, EPA uses
three helicopters equipped with special remote and contact sensors. Each
lake is sampled, at multiple sites, three times during the growing season.
The pontoon-equipped helicopters land on the lakes where probes are lowered
into the water to measure dissolved oxygen, conductivity," temperature, and
turbidity at different depths. Samples from various depths are analyzed
for approximately 15 parameters. Algae are identified and counted. Algal
assays are conducted on the lake waters to determine the productivity poten-
tial of the water and to assist in the identification of the limiting nutri-
ent. The major streams entering or leaving the lakes are sampled for nitro-
gen and phosphorus. Stream flows are measured, and sewage treatment plant
effluents are also sampled.
Another phase of the survey's work is to develop techniques to deter-
mine relationships between land-use patterns and nitrogen and phosphorus
supply by using aerial photography. Land-use types will be correlated with
the trophic condition of lakes and refined nutrient flux factors developed
for various land-use types and geographic areas. This and other informa-
tion from the survey has made it possible to focus on the following aspects:
1) Relationships between drainage-area characteristics and
non-point source nutrients in streams (Figures 1 and 2);
2) quantitites of nitrogen and phosphorus in wastewater
effluents;
3) the relationships of phosphorus and nitrogen to the trophic
state of northeast and north-central lakes and reservoirs
(Figures 3 and 4); and
4) an approach to a relative trophic index system for classi-
fying lakes and reservoirs.
The final step of the survey will be the interpretation of the data for
each lake and, in cooperation with appropriate state agencies, the deriva-
tion of subsequent recommendations for remedial action.
Algal Assays
Algal assays have been for some time an extremely valuable tool for
evaluating water quality in relation to eutrophication. Many investigators
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FOREST
MOSTLY
FOREST
MIXED
MOSTLY
URBAN
MOSTLY
AGRIC.
AGRI-
CULTURE
-
0
MEAN TOTAL NITROGEN CONCENTRATIONS
vs. LAND USE
—I 1 T 1 T~
0,850
0.885
1.282
1.286
.812
1,0
MILLIGRAMS
2.0
PER
LITER
T
4,170
4.0
Figure 1. The relationship between mean total nitrogen concen-
trations in streams and land use in the Eastern United States.
MEAN TOTAL PHOSPHORUS CONCENTRATIONS
vs. LAND USE
FOREST
MOSTLY
FOREST
MIXED
MOSTLY
URBAN
MOSTLY
AGRIC.
AGRI-
CULTURE
|o.OI4
0.035
0.040
0.066
0.066
0
0,05 OJO
MILLIGRAMS PER LITER
0.135
0.15
Figure 2. The relationship between mean total phosphorus concen-
trations in streams and land use in the Eastern United States.
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u>
100 ^
"I 1 I I (Hl| 1 1 Mill!) 1 1 I II II Tj 1 1 I I I III
• EUTROPHIC
: A MESOTROPHIC
10 - • OLIGOTROPHIC
x
-------
improvised algal assays to meet their specific needs, but because they were
nonstandard, they offered no basis for comparing results among laboratories
or among samples obtained from different geographic areas. Early in 1968
EPA prepared a tentative procedure for a proposed standardized algal growth
test. It was intended to (1) identify and determine the availability of
algal growth-limiting nutrients; (2) quantify biological response to
changes in concentrations of algal growth-limiting nutrients; and (3) deve-
lop a rational framework for application of assay results to practical
problems.
Early in the developmental effort it became apparent that emphasis
should be placed on a static bottle-type test. In August 1971 the "Algal
Assay Procedure: Bottle Test (AAP)" (National Eutrophication Research Pro-
gram, 1971) was published after inter laboratory precision tests at eight
laboratories showed excellent agreement in the data. It was then concluded
that the bottle test had undergone sufficient evaluation and refinement to
be considered reliable. It has been applied to numerous situations to
assist in solving and understanding eutrophication problems. The assay pro-
cedure is included in the 14th edition of "Standard Methods for the Examina-
tion of Water and Wastewater" by the American Public Health Association, et.
al.
Eutrophication Control
Since the early use of copper sulfate to control algae (Moore and
Kellerman, 1904), an array of ecologically based procedures has evolved.
What are today's options for controlling eutrophication? How can lakes be
protected from further degradation and, equally important, how can
eutrophic lakes be restored? Without question, the most promising preven-
tive and restorative measure is to curb nutrient supply. This objective
can be approached in several ways.
Nutrient Diversion--
A commonly used method to reduce nutrient input is to divert point-
source nutrient-rich wastewater around or away from the receiving body of
water. The early programs at Madison, Wisconsin, and Seattle, Washington,
are well known. The same approach has been used more recently at Lake
Sammamish, Washington, and Twin Lakes, Ohio. Although these lakes have im-
proved, they have not responded to the same degree as Lake Washington.
As an alternative to diverting wastewater the effluents can be sprayed
on the land and still protect the lake. This approach required much more
land area than conventional treatment facilities; hence, when land is
costly or not available, it may not be a viable remedy. Land requirement
is about 1 ha per 300-400 population served. The process requires very
little construction facilities, has a low energy requirement for operation,
and can be easily automated for unattended operation at small facilities.
Success is also very dependent on climatic factors, and year-round use will
obviously be impractical in regions with severe winters.
The State of Michigan has completed a program of soils testing for phos-
phate absorption capacity, and New York State has undertaken such a program.
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Processing organic wastes and returning them ultimately to the land is one
of the most far-sighted methods for dealing with them. One of the largest
demonstrations of this technique is currently underway at Muskegon,
Michigan.
Another way to reduce the nutrient input is to remove specific nutri-
ents by advanced waste treatment. Either nitrogen or phosphorus or both
can be removed. One of EPA's current major research undertakings is a
unique project to demonstrate the feasibility and dynamics of restoring a
deteriorating lake by removing phoshorus from municipal wastewater flowing
into it. Shagawa Lake, in northeastern Minnesota, was selected to demon-
strate this technique. High inputs of phosphorus supplied by the lake
shore city of Ely have caused excessive productivity and undesirable
conditions in the lake. Ely, producing at a maximum about 3500 m3
(1,000,000 gal) of wastewater daily, has had a municipal sewer system since
1901 and a secondary treatment plant since 1954. The effluent has always
been discharged into Shagawa Lake. As a result, over the last 70 years,
the lake has become increasingly eutrophic, a condition in great contrast
to the near-pristine surrounding lakes. A $2.3 million tertiary waste-
treatment facility, designed to remove more than 99% of the phosphorus in
the wastewater from the secondary sewage treatment plant, was constructed
with 95% financing by EPA. Full-scale operation began in early 1973.
Phosphorus is removed by chemical treatment, primarily with lime and lesser
amounts of ferric chloride, settling, and filtration. Only about 68 kg of
phosphorus now enter the lake each year from this source, instead of the
6,800 kg before tertiary treatment. This plant is unique in the United
States in removing phosphorus from all of the municipal wastewater to a
residual of 0.05 mg/liter.
According to existing mathematical models of Shagana Lake, recovery
should be rapid, very likely reaching a new phosphorus equilibrium in 1% to
2% years. However, when taking into account the phosphorus contained in
the bottom sediment, and its exchange with the overlying water, additional
time must be allowed for depletion of this nutrient source. Nevertheless,
significant reduction in the phosphorus level of lake water has been noted,
and the chlorophyll a_ concentration has been reduced from pre-treatment
levels.
Product Modification—
Still another way to limit nutrient loading to lakes is to modify
nutrient-rich products to reduce their growth-promoting potential. The
best example is phosphorus compounds in detergents. It has been estimated
that on a per capita basis 0.96 kg of phosphorus per capita year was
utilized in the household and 0.24 kg per capita year was utilized in indus-
try (Porcella et al., 1974).
On the average, roughly half of the phosphates entering U.S. streams
come from municipal wastes and urban runoff. The other half comes from
natural runoff, industrial and agricultural wastes, and animal feed lots.
About half of the phosphates in domestic wastes are of detergent origin
(Hatling and Carcich, 1973). Thus, detergents account for about one-
quarter of the phosphates discharged into lakes and streams.
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Onondaga Lake, New York, portrays the result of modifying detergent
compounds. Following the implementation of local and state legislation in
1971-72 that limits the phosphorus allowed in detergents to 8.7% as P, a
decrease of 54% in the concentration of total inorganic phosphate occurred
in the Syracuse sewage treatment plant discharge to Onondaga Lake. The
average total inorganic phosphate concentration in the lake also decreased
by 57%. In the first full growth season after implementation of the law,
the blue-green alga Aphamizomenon was newly absent in the succession of
phytoplankton.
In-Lake Treatment—
Once the nutrients have entered a lake, the problem of eutrophication
control is more complex. However, various control methods are under in-
vestigation. One can increase the nutrient output, immobilize the nutri-
ents, withdraw nutrient-rich hypolimnetic waters, or dredge to remove
nutrient-rich sediments. One may also treat the symptoms, such as nui-
sance algae, plants, and fish, by applying poisons or toxins, by harvesting,
or by biological grazing.
Algicides and herbicides—Chemical treatment has been a widely used
method to improve the appearance and usefulness of lakes. It is intended
to limit specific populations of organisms, such as blue-green algae,
higher aquatic plants, or unwanted fish populations that become nuisances.
The chemicals vary in their cost, effectiveness, toxicity, and persistence.
In any event, the result is only temporary or "cosmetic" in that it treats
only the symptoms and not the cause of the problem. In addition,
decomposition of the target species serves to regenerate nutrients that
allow for continued biological development of the same or different popula-
tions.
Mechanical harvesting—A method of eutrophication control that is being
used with decreasing frequency is the harvesting of plants or animals from
the lake. The product may be stumps or sunken logs as at Marion Pond,
Wisconsin, rough fish, or higher aquatic plants. Weed-harvesting equipment
is available, but attempts to develop equipment and procedures to harvest
algae have not been successful. Harvesting obviously removes some
nutrients from a body of water, but the amount of phosphorus and nitrogen
removed is exceedingly small. In the makeup of plants an average reported
value for P is 0.24% and for N is 2.3%, dry weight concentration. A re-
cently published report (Peterson, Smith and Malueg, 1974) on a harvesting
study on Lake Sallie, Minnesota, states: "Perhaps the most significant
conclusion to be derived from this study is that continuous harvest of
aquatic plants from Lake Sallie during the growing season could not offset
the high loading of phosphorus and nitrogen. The net-weight harvest of
428,000 kg of plants was successful in removing only 1.3% of the total
phosphorus to the lake, or 1.03% of the phosphorus contained in the water
volume of the lake during the fall circulation period.
In spite of this, harvesting often can be justified on the basis of
aesthetic values alone. Research by the University of Wisconsin on Lake
Mendota indicates that one harvesting will reduce the amount or regrowth to
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about 50% of the controls, two harvests will result in about 75% reduction,
and three harvests almost totally eliminate the plants for that year- The
researchers recommended two harvests, one in June and the other in July,
for that climate. None of the treatments had an appreciable effect on the
subsequent year's growth. So far little is known of the effects of har-
vesting of higher aquatic plants on the phytoplankton.
A demonstration in the State of Florida provides another example of
weed harvesting. The St. John's River is the largest river entirely within
the state, approximately 480 km long. Problem weeds such as water hya-
cinths occur in areas where water use for navigation is extensive. The
weeds also retard irrigation and drainage and reduce game fish and water
fowl. The decomposition of detritus from these plants also depletes the
oxygen from the lower waters. Control of water hyacinths on the St. John's
River has been carried out with chemicals for a long time. Now researchers
are experimenting with harvesting techniques. In Florida alone, more than
40,000 ha of water are covered with water hyacinths, despite extensive and
continuous programs of control by various governmental agencies.
Dredging--A procedure that can be thought of as an extension or modifi-
cation of harvesting is dredging of the sediments of a eutrophic lake. Per-
haps in many lakes the sediments are an important source of nutrients that
may be cycled to the overlying waters, especially at certain times of the
year. In theory, dredging would remove this nutrient source, but there are
several problems, not the least of which is disposal of dredged material.
These problems are being addressed in an extensive research program
recently undertaken by the U.S. Army Corps of Engineers.
Dilution or Flushing—Another method of eutrophication control is flush-
ing or dilution. Use of this method is limited by the availability of
fresh water. Two nutrient dilution procedures have been attempted: (1)
pumping water out of the lake, thus permitting increased inflow of nutrient-
poor groundwater, and (2) routing additional quantities of nutrient-poor
surface waters into the lake. The first has been used at Snake Lake, Wis-
consin. The second has been tried in several places. One of the most
successful experiments was done at Green Lake, Washington, where, after 5
years of flushing (plus some initial dredging), the blue-green algal
standing crop was suppressed, with elimination of Aphanizomenon. Flushing
has also been tried on a small scale at Moses Lake, Washington.
Aeration—It is also possible to immobilize nutrients in eutrophic
lakes through aeration of hypolimnetic waters where large reservoirs of
phosphorus may accumulate. Aeration methods generally fall into two
groups: those that destratify the lake and thus affect all depths, and
those that aerate only the bottom waters and do not destratify the lake,
i.e., hypoh'mnetic aeration. When destratification is accomplished, the
lake becomes isothermal with oxygen present at all depths and other chemi-
cal conditions fairly uniform. Hypolimnetic aeration has certain
advantages over destratification. Nutrients are not upwelled into the
surface waters where they may promote algal growth. Further, hypolim-
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netic aeration permits the establishment of a cold-water fishery such as
trout or salmon, whereas destratification may preclude such a fishery by
eliminating the cold-water region.
Examples of aeration projects for eutrophication control are at Cline's
Pond, Oregon, where destratification aeration was used; at Lake Waccabuc,
New York, where the "Limno" hypolimnetic aerator is being used; and at
Ottoville Quarry, Ohio, where hypolimnetic aeration is achieved by a pro-
cess called "side stream pumping".
Nutrient inactivation—A promising approach for a wide range of situa-
tions is nutrient inactivation. This involves treatment of lake waters in
situ with a chemical to precipitate phosphorus. Inactivant materials that
have shown particular promise in laboratory and field studies are aluminum,
zirconium, and fly ash. Experiments with aluminum compounds are presently
being conducted on lakes in Wisconsin, New England, Ohio, and Washington.
Zirconium is being tested in a controlled pilot field study in Oregon, and
a similar experiment, utilizing fly ash, is under way in Indiana. It is
anticipated that such treatments will be particularly efficacious in lakes
with very long retention times.
Hypolimnetic withdrawal/selective discharge--Hypo1imnetic withdrawal
has been used to improve dissolved oxygen conditions near the bottom of a
lake and to increase nutrient export. In bodies of water that stratify,
this technique permits the removal of anaerobic, nutrient-rich deep
waters. The technique is suitable for waters with outlet controls, such as
reservoirs, or in lakes with surface withdrawals by installation of a
siphon from a point of maximum depth. The surface discharge is, or can be,
completely blocked off. This technique has been used in Wisconsin, Ohio,
and other states. A potential problem with its use is the triggering of
increased macrophytic growth and low dissolved oxygen in the downstream
channels.
Drawdown—Sediment exposure and desiccation via lake drawdown has been
undertaken on impoundments for various purposes. In favorable sites this
procedure can reduce the rooted aquatic plants by desiccation. The effect
of drying on sediment chemistry and possible nutrient release is now being
studied, particularly in Florida and Louisiana. Presently 13 Louisiana
impoundments are being managed by water drawdown to aid in control of
aquatic vegetation and fish populations.
Biological control—The control of particular problem species by mani-
pulation of biotic interactions has been a much desired goal in recent
years. Evaluation of biological controls has been limited, however; most
testing has been done in the laboratory or experimental ponds. Con-
siderable publicity has been given to programs which have sought to
decrease the density of weed species through the introduction of host-
specific predators, and a great amount of research has been expended in
predator control of macrophytes. One such program has been relatively
successful: the flea beetle has reduced populations of alligator weed
considerably in some areas of the Southeast.
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A somewhat less specific herbivore, the grass carp or white amur, has
been released in numerous lakes in Arkansas, where it has apparently been
able to successfully control undesirable submerged weeds. Its widespread
introduction into this country, however, is still the subject of much
apprehension and study.
Similar control programs involving crayfish, a specific weevil for
water hyacinth, and other insects are currently under investigation.
Aquatic mammals such as the manatee and other animals such as snails and
swans have also been tried. Although most of these animals have been some-
what effective on a local basis, few are effective over a broad geographic
range. The need to carefully consider and anticipate the total effects of
introduced or exotic species on the natural ecology is well known.
Relatively little work has been done with biological control of algal
populations. Bacteria and viruses have been isolated that destroy blue-
green algae, but to date only laboratory tests have been conducted. No
full-scale, in-lake treatment has ever been tried in the United States.
The control of undesirable macrophytes with plant pathogens, mostly fungi,
may show some potential and is currently being evaluated on a small scale.
Biomanipulation, or facilitating desirable interactions among different
segments of the whole ecosystem, has long been a desirable goal. Some
possibilities in this direction are under study. Attempts are being made
to reduce phytoplankton abundance by increasing the number of grazers by
either direct innoculation or by controlling the zooplankton by disease or
carnivorous fish introductions. Attempts to exploit the competitive or
inhibitive reactions among aquatic weed species are also being studied as
possible control measures.
Legislation
Lake-restoration measures are in a very early stage of development.
Much of the technology is still being applied in laboratories, in experi-
mental ponds, or in pilot lake studies. No technique can be applied indis-
criminately to every problem lake; each must be studied and evaluated suffi-
ciently to assure that the most appropriate course of action is taken. Ob-
viously, a whole range of remedial methods must be made available. Public
Lake 92-500, the Federal Water Pollution Control Act Amendments of 1972,
will certainly help in this regard because it authorizes funds to support
state programs for lake restoration. The Congress has appropriated $4
million to EPA for this purpose; approximately 10% of the funds have been
designated for evaluation purposes. It is expected that funding will be
increased next year.
To limit fertility in lakes, several states have passed laws setting
forth regulations affecting nutrient loading. Some of them are the
following:
Minnesota has passed legislation to set effluent standards for phos-
phorus in municipal discharges. If the discharge enters a lake directly,
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the phosphorus concentration must not exceed 1.0 mg/liter; if to a lake via
a river, 2.0 mg/liter. Similarly, Illinois has adopted an effluent stand-
ard of 1.0 mg/liter phosphorus for discharge to Lake Michigan.
In Iowa, laws were passed in 1971 that provide for mandatory soil con-
servation. Iowa's Conservancy District Act established conservancy dis-
tricts and declared soil erosion resulting in siltation damage to be nui-
sance. The act also directed the commissioners to establish soil-loss
limits for their districts.
The Wisconsin Shoreland Protection Statute authorizes and requires
counties to adopt pollution-control regulations for the shoreland areas.
The law sets out zoning, sanitary code provisions, and subdivision regula-
tions.
Indiana, Iowa, and Minnesota have adopted various kinds of farm-animal-
waste regulations, aimed chiefly at problems related to large feedlot
installations.
A number of states have acted to limit the phosphorus content of deter-
gents. New York and Indiana passed such laws in 1971. The New York law
reduced the phosphorus content to 8.7% by January 1972 and to a trace by
July 1, 1973. The Indiana law limited phosphorus to approximately 5% after
the beginning of 1972, and to 3% after January 1973. The law has changed
in 1972 to read 8.7% after January 1, 1972 and to zero after January 1,
1973, Indiana thus becoming the first state to completely ban phosphorus in
household laundry detergents.
Laws limiting phosphorus in detergents were also passed in Florida,
Maine, Michigan, Minnesota, Connecticut, and Oregon, as well as in Chicago,
Illinois, Akron, Ohio, and Dade Country, Florida.
The Environmental Protection Agency is presently drafting phosphorus
criteria for recreational waters. Although the agency does not propose a
limit of acceptability for phosphorus, it gives guidelines for the
establishment of total phosphorus criteria in receiving waters. These will
include both a concentration, which prescribes maximum acceptable levels,
and a loading value in the form of an annual allowable specific loading to
the receiving water.
The United States and Canada joined together on April 15, 1972, under
the Great Lakes Water Quality Agreement "to restore and enhance water
quality in the Great Lakes system" (Great Lakes Water Quality, 1972).
Annex 2 of the agreement pertains to control of phosphorus. It specifies
effluent requirements for municipal waste treatment plants, goals for in-
dustries, and reductions in input from animal husbandry operations.
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THE FUTURE
The science of limnology and the study of eutrophication have come of
age since 1900, but considerable work still remains if this nation is to
have clean lakes and streams. The following statements identify specific
areas where intensified research is needed if we are to succeed in these
goals:
1) Develop and utilize remote sensing so that water bodies can
be quickly trophic level.
2) Develop methods to examine and manage lakes as part of an
entire watershed.
3) Delineate the role of sediments as a source or sink of nutri-
ents, to facilitate predictions of impact on lake recovery
prior to initiating control or restorative practices.
4) Evaluate the role of the thermocline as a barrier to the
transfer of chemical and biological material and possibilities
for beneficial manipulation.
5) Understand the reasons for seasonal succession of algal types
and, in particular, the reasons for the appearance and domi-
nance of blue-green algae.
6) Determine interactions between macrophyte and phytoplankton
populations and the effects on one when the other is mani-
pulated.
7) Develop methods to control macrophytes to achieve a balance with
desirable uses of the lake.
8) Develop and evaluate methods of aquatic ecosystem mangement
through biological manipulations so that the water body pro-
duces the most desirable product.
9) Evaluate useful products derived from harvestable material
from water bodies. Such products would include soil condi-
tioners, pharmaceutical materials, animal feed, and energy
source.
10) Develop techniques to predict the success of control or
management methods on lakes through mathematical modeling.
11) Determine the socioeconomic aspects of cultural eutrophica-
tion and lake restoration including recreational impact,
effects on commercial fisheries, public health, and cost of
water treatment.
102
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Despite the research and refinements that are still to come, many signi-
ficant advances in eutrophication and lake-restoration research have been
made. Some techniques now available can be and are being used with
reasonable success on individual lakes. Other options need to be developed.
The continuing interdisciplinary interest in this area of water-resources
mangement is heartening. Scientists, engineers, economists, and others
working together will ultimately find solutions to save and protect our
aquatic habitats. We must act to protect mankind's great natural heritage
without destroying it. In trying to find the ways to do it we must cherish
the words of Chief Seattle:
"The earth does not belong to man; man belongs to the earth."
REFERENCES
Great Lakes Water Quality. 1972. Agreement with annexes and texts and
terms of reference, between the United States of America and Canada.
Signed at Ottawa, Canada, April 15, 1972. 75 p.
Hetling, L.J., and 1.6. Carcich. 1973. Phosphorus in wastewater. Water
and Sewage Works 120: 59-62
Likens, 6.E. (ed.). 1972. Nutrients and eutrophication: the limiting-
nutrient controversy. Special symposia - Vol. 1. Proc. Symposium
Michigan State Univ., February 11 and 12, 1971. American Society of
Limnology and Oceanography, Inc. 328 p.
Moore, G.T., and K.F. Kellerman. 1904. A method of destroying or pre-
venting the growth of algae and certain pathogenic bacteria in water
supplies. Bur. Plant Ind., U.S. Dept. Agric. Bull. 64. 44 p.
National Academy of Sciences. 1969. Eutrophication: causes, consequences,
correctives. Proc. International Symposium on Eutrophication, Univ.
Wisconsin, Madison, June 11-15, 1967. Washington, D.C. 661 p.
National Eutrophication Research Program. 1971. Algal assay procedure
bottle test. U.S. Environmental Protection Agency, Corvallis, Ore.
82 p.
Peterson, S.A., W. Smith, and K.W. Malueg. 1974. Full-scale harvest of
aquatic plants: nutrient removal from a eutrophic lake. J. Water
Pollut. Contr. Fed. 46: 697-707.
Porcella, D.B., A.B. Bishop, J.C. Anderson, O.W. Asplund, A.B. Crawford,
W.J. Grenney, D.I. Jenkins, J.J. Jurinak, W.D. Lewis, E.J. Middlebrooks,
and R.M. Walkingshaw. 1974. Comprehensive management of phosphorus
water pollution. U.S. Environmental Protection Agkency, Washington,
D.C. Socioeconomic Environmental Stud. Ser. EPA-600/5-74-010. 411 p.
103
-------
Stewart, Kenton M., and Gerard A. Rohlich. 1967. Eutrophication--A review.
A report to the state water quality control board, California. The
Resources Agency, Sacramento, Cal. Pub!. 34. 188 p.
104
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SECTION 10
DETERMINING THRESHOLD AND BIOLOGICALLY DANGEROUS
CONCENTRATIONS OF BLUE-GREEN ALGAE IN WATER BODIES
L.A. Sirenko, A.Ya. Malyarevskaya, and T.I. Birger^
The economic activities of man have caused eutrophication in many water
bodies. One of the significant results of this is a change in the species
composition and number of aquatic organisms in an affected area. This often
causes a disturbance in the regulation processes in the ecosystem of the
water body. The blue-green algae bloom may be the most important example
of disturbance in the ecological balance under the influence of anthropo-
genic factors. In this case, excessive development of individual species
of algal flora determines the whole complex of the internal water body pro-
cesses and the ecosystem's final biological productivity. As is known, in
their vital processes algae excrete into the environment about 30% of the
total carbon absorbed by them during 24 hours or about 40% of the daily
pure photosynthesis production.
Excreted products include: Organic acids which determine the environ-
mental buffer action and its pH; amino acids and peptides, which contribute
to the formation of the complex and lower the toxicity of heavy metals;
polysaccharide substances which adsorb on their surface the most varied
types of ions, aldehydes, terpenes, polyphenol compounds; and also other
biologically active substances which dominate biological activity.
Toxins found in individual species of algae (blue-green, flagellateae,
peridium) also play a very important role among exogenous metabolites.
Blue-green algae excreting some substances during life processes, and
decomposition not only form a definite biotic environmental background, but
also change the hydrochemical indices of the environment if their accumula-
tion is considerable. In this case, the quantity of oxygen diminishes, the
content of carbonic acid increases and the environmental reaction changes.
In other words, the effect of algae on the aquatic organisms depends on a
whole complex of biotic and abiotic factors.
In addition to this direct influence, blue-green algae may also affect
the aquatic organisms indirectly. Namely, the substances they excrete may
intensify or weaken to a considerable degree the action of different chemi-
Institute of Hydrobiology, Academy of Sciences, Ukrainian SSR.
105
-------
cal substances entering the water body. This effect may result in the form-
ation of complex compounds, suppression or stimulation of the bacterioflora,
and the appearance of free radical compounds if photosynthetic oxygen and
some other factors are present in the environment.
This indicates that when determining the Maximum Permissible Concentra-
tion for toxic substances, it is necessary to take into account the influ-
ence of exogenous metabolites of algae on toxic compounds and likewise it
is necessary to determine biologically dangerous concentrations of the
natural metabolites for fish and other aquatic organisms.
This information examines the possibility of determining, for fish, the
threshold and biologically dangerous concentrations of metabolites of blue-
green algae that cause "blooms".
During cell reproduction of Microcystis, the main stimulus for algae
"blooms", the environment accumulates exogenous metabolites which have a
high biological activity (Goryunova, 1966; Sirenko, 1971). They in-
clude polyphenol compounds, and likewise polynucleotides. We determine
that in concentrations up to about 0.28 ppm, phenol compounds suppress the
growth of other examples of algal flora, not affecting the life processes
of the species-producer. When the concentration of phenol compounds in-
creases to 1.3 ppm, autoinhibition of the algal growth and reproduction is
observed.
This indicates the role of the metabolite concentrations in the cell-
division regulation processes; in this case, it concerns the algae cells in
culture and natural conditions. If we changed the concentration, we would
be able to regulate to a definite limit the number of algae cells in a pop-
ulation.
The metabolites produced by algae are not of less importance for the
life processes of other aquatic organisms entering into communities and im-
portant in the self-purification processes. For example, the effect of
blue-green algae metabolites on the decrease in the producing capacity of
Daphnia was shown (Braginskiy et al., 1965).
As for fish, there has been little study of the influence of natural
blue-green algae metabolites on them. The threshold and biologically dan-
gerous concentrations of these metabolites for all practical purposes, have
not been formulated.
At present, when determining the maximum permissible concentration of
some basically artificial substances entering the water body, the following
criteria are usually considered (Stroganov, 1972):
1. Mineralization processes of organic substances.
2. Oganoleptic indices of water and water organisms (especially
of fish).
106
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3. Survival, growth, reproduction, fertility, and quality of the
aquatic organisms' progeny.
All the enumerated criteria most completely characterize the ecological
and toxicological situation in the water body and are widely used to deter-
mine the maximum permissible concentration of artificial substances.
In view of the necessity to develop express methods which would help to
detect the aquatic organism-metabolism links changing in the first instance
under the influence of toxicants, a study of the biochemical mechanisms of
toxicant effect on the aquatic organisms has been planned. There are some
data affirming that the effect of many toxicants is the result of effects
on enzyme processes in the aquatic organism. For example, an anticholine-
sterase effect of phosphorus organic compounds on fish has been ascertained
(Metelev, et al., 1971).
The investigation carried out in our Institute (Malyarevskaya, Birger,
Arsan, Solomatina, 1973) give an idea of the threshold and biologically dan-
gerous concentrations of biological toxicants for fish (e.g., effect of
blue-green algae Microcystic aeruginosa Kutz. emend. Elenk). The experi-
ments were carried out from 1964 to 1974 in laboratories and in water
bodies exposed to a heavy "bloom" caused by mass development of blue-green
algae. Effects of various blue-green algae concentrations on fish (pike
perch Lucioperca lucioperca L., perch Perca fluviatilis L., ide Leuciscus
idus L., cruc i an Carassius carassius L., and Hypophthalmichtys molitrix
Val.) were studied.
In long-term experiments, small quantities of algae were applied (0.03-
0.30 g/liter). Since the fish did not die, they were conditionally named
"nonlethal". When more considerable concentrations (0.6 - 5.0 g/liter)
were used in acute experiments, the fish died within a period of 6-64
hours, depending on the fish species and algae concentration and condition
(living, decaying). Such algae concentrations were called "lethal".
The biochemical composition of fish (content in their bodies of dry or-
ganic substances, ashes, protein and its amino acid composition, lipids,
vitamins B.i, B2 and enzyme activity of thiaminase, cholinesterase, trans-
aminse and content of nicotinamide coenzymes) and also the fish metabolism
(interchange of gases and nitrogen exchange) have been investigated and the
effect of various algae concentrations on them has been ascertained.
The effect of blue-green algae on fish is conditioned by the complex of
biotic and abiotic factors which includes the effects of metabolism pro-
ducts, algae decomposition, and changes in the hydrochemical indices in the
environment.
Non-lethal algae concentrations (0.03-0.30 g/liter) do not cause the
death of fish, but if a fish inhabits waters characterized by such a
quantity of algae (especially 0.3 g/liter), it results in determinable
changes in the metabolism, i.e., suppression of plastic processes and
intensification of energetic processes. The growth of dry and organic sub-
107
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stances and protein in fish drops and nitrogen consumption is reduced. The
expenditure of nitrogen for energetic processes increases (Table 1).
Considering that the most evident negative deviations from the control
are observed in fish in concentrations of 0.30 g/liter, this quantity of
algae may be considered a threshold concentration.
Lethal concentrations of blue-green algae caused more significant
changes in the biochemical composition and the fish metabolism. The data
show that fish lost dry and organic substances, protein and lipids. The
nitrogen balance was negative. Respiration intensified at a stage of
heightened movement activity and slowed down before death. A change in the
content of free amino acids (Table 2) and protein hydrolysates amino acids,
and also transaminase activity, pointed to the protein synthesis disturb-
ance. A decrease in nicotinamide coenzyme oxidized forms"confirmed that
considerable changes in oxidation-reduction processes in tissues took place.
Given the effect of the blue-green algae lethal quantities on fish,
changes in the thiaminase enzyme activity and total thiamine content
(vitamin Bx) were most significant. A thiaminase activity increase of 21-
40% in organs and tissues of fish influenced by the blue-green algae and
thiamine content drop of 38-50% in comparison with the control value, re-
sulted in a convulsion stage. A thiaminase activity increase of 28-55% and
a thiamine content drop of 49-74% caused the death of fish (Table 3). A
number of experiments, in which thiamine chloride injected at the initial
paralysis stage stopped the convulsions and prolonged the fishes' life, cor-
roborate that avitaminosis B , given the influence of lethal concentrations
of blue-green algae, is the reason for the fishes' death. This is also in-
directly confirmed by the fact that thiaminase activity in fish in natural
conditions (ponds, reservoirs) during the "bloom" period was heightened,
but the total thiamine content was lower than in autumn when no "blooms"
were observed (Table 4).
A shortage of vitamin B1 in organs of fish, given the influence of
lethal concentrations of blue-green algae, causes a disturbance in all meta-
bolism processes in which it participates. As mentioned above, in this
case the protein exchange, and likewise biosynthesis of lipid structure
compounds and normal transformation of substances in Krebs' cycle, are dis-
turbed. Changes in biochemical processes result in disturbances of some
functions. In particular, Bi-avitaminosis involves changes in the nervous
system's functional state. The latter is corroborated by a drop in the
cholinesterase activity in the fish's brain, given the influence of lethal
concentrations of blue-green algae. Some other symptoms typical of the
thiamine shortage are also observed in these fish, namely: Disturbance in
liver functions, the alimentary canal, and cardiovascular system; hemorrhage
in organs; and pathological changes in blood-formation. Blood analyses of
fish caught in waters covered with "bloom" areas support the latter
(Komarovskiy, 1970).
Thus specific phenomena, e.g., Bj-avitaminosis, resulting in a number
of nonspecific changes—in particular, non-coordination of energetic and
108
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TABLE 1. CHARACTERISTICS OF NITROGEN EXCHANGE IN THIS YEAR'S GROUP OF PERCH
Concentra-
tion of
blue-green
algae (g/1
of live
weight)
--
0.03
0.30
0.30
Physiologi-
cal state
of
algae
Control
Living
Living
Decomposed
Average
daily
ni trogen
ration of
1 fish
(mg)
2.8
2.2
1.9
1.4
Nitrogen
accumulated
per 24 hrs
in organism
of 1 fish (%
of the ave-
rage daily
nitrogen
ration)
30.0
10.9
12.3
2.8
Nitrogen
by one fi
daily
wi th li-
quid ex-
cretions
46.6
75.9
71.8
84.6
excreted during 24 hrs
sh (% of the average
nitrogen ration)
with ex-
crements
23.4
13.2
15.9
12.6
total
excreted
nitrogen
70.0
89.1
87.7
97.2
-------
TABLE 2. CONTENT OF FREE AMINO ACIDS IN ORGANS AND TISSUE OF IDE
YEARLINGS (mg/g of Fresh Tissue)
Ami no
Acids
Cystina
Lysine
Histidine
Agrinine
Asparagine
Acid+Serin
Glycine
Glutamine
Acid+Treonine
Alanine
Tirosine
Metionine+
Valine
Phenilala-
nine
Lei cine
Indices
Control
Experiment
Control
Experiment
Control
Experiment
Control
Experiment
Control
Experiment
Control
Experiment
Control
Experiment
Control
Experiment
Control
Experiment
Ccnt.ro!
Expc-riment
Control
Experiment
Control
Experiment
Intestine
M ± m
0.43±0.04
4.74±0.13
0.17±0.02
0.13±0.02
0.23±0.01
0.31+0.01
0.34±0.10
0.17±0.03
0.68±0.09
0.52±0.13
0.13±0.02
0.58±0.06
1.08±0.14
1.43±0.23
0.55±0.03
0.75±0.04
0.08+0.01
0.04±0.01
0.49±0.01
0.03+0.01
0.16±0.01
0.07±0.01
0.19±0.06
0.17±0.03
diff
31.3
0.4
2.6
1.6
1.0
7.38
1.15
4.0
0.49
1.44
9.44
0.36
Liver
M ± m
0.34.t0.10
1.29±0.04
0.29+0.05
0.31±0.03
0.13i0.05
0.25±0.00
0.42+0.01
0.31±0.07
0.22+0.03
0.64+0.14
0.04±0.00
0.1U0.01
0.15±0.04
2.0U0.15
0.09±0.02
0.36±0.04
0.08±0.00
0.50±0.01
0.20+0.00
C. 13+0. 02
0.05±0.00
0.17±0.04
0.01 ±0.00
0.12±0.03
diff
8.46
0.78
2.29
1.26
2.94
4.93
11.99
5.87
47.8
5.0
2.91
9.9
Muscles
M ± m
0.09±0.01
0.26±0.04
0.93+0.07
0.32±0.05
_..
~
0.05±0.01
0.04±0.08
0.33±0.01
0.35±0.08
0.16+0.01
0.12±0.02
0.46±0.04
0.13+0.01
0.50±0.06
0.08+0.01
0.10±0.02
0.07±0.01
0.03+0.02
0.03+0.05
0.04±O.OG
0.03±0.01
0.04+0.00
0.06±0.01
diff
4.12
6.5
—
5.5
0.3
0.23
6.22
6.28
0.6
0.47
0.85
1.82
-------
TABLE 3. EFFECT OF LIVING BLUE-GREEN ALGAE (5 g/1) ON CONTENT OF TOTAL THIAMINE
(yg/g) AND THIAMINASE ACTIVITY (yg/hr) IN LIVER AND INTESTINE OF FISH
Fish
Species
Pike perch
Perch
Ide
Pike perch
Perch
Ide
Control
Total
Thiamine
M ± m
4.70±0.
7.27+0.
4.62+0.
4.29±0.
3.32±0.
4.71±0.
20
23
17
10
18
23
Thiaminase
Acti vi ty
M ± m
709. 23± 8.50
565.64±25.00
663.22±21.80
517. 64± 9.54
699.14±18.00
530.34±26.00
Initial
Total
Thiamine
M ± m
LIVER
1.89±0.10
5.53±0.26
2.65±0.15
INTESTINE
1.97±0.10
1.90 ±0.10
2.22+0.13
state
sis
of paraly-
Thiaminase
Activity
M ± m
992
775
808
720
849
661
.21+16
.04±17
.23±16
.62+ 1
.41 ±28
.42±22
.00
.00
.40
.08
.60
.00
Death stage
Total
Thiamine
M ± m
1.
3.
1.
1.
1.
1.
21+0.10
45±0.18
60+0.16
28±0.10
67±0.28
42+0.13
Thiaminase
Activity
M + m
1095. 80± 8
841.14±22
861.17±22
777. 69± 1
901.83±39
767.33±24
.75
.70
.30
.80
.30
.40
-------
ro
TABLE 4. EFFECT OF BLUE-GREEN ALGAE ON CONTENT OF TOTAL THIAMINE (yg/g)
THIAMINASE ACTIVITY (ng/hr) IN LIVER AND INTESTINE OF FISH
Indices
Fish
Species
SCARDINIUS
erythrophtalnus L.
Roach
Bream
SCARDINIUS
erythrophtalnus L.
Roach
Bream
Total
thiamine
M±m
7.13±0.31
7.37±0.40
7.34±0.84
3.83±0.20
2.88±0.30
3.20±0.20
Autumn
(no "bloom")
Thiaminasae
Activity
M±m
LIVER
363.07±6.55
438.18±4.82
445.93+9.20
Summer
("bloom" period)
Total
thiamine
M±m
4.62±0.22
4.60±0.63
4.40±0.25
INTESTINE
280.10+5.20
384.12+9.00
354.92±3.13
2. 3U0.18
1.46+0.10
1.72±0.14
Thiaminasae
Activity
M+m
550.66±65.40
614.75+5.50
538.81+5.70
347.00±8.00
507.79±5.80
461.67+13.10
-------
plastic processes and changes in the above mentioned factors including
blood analysis—develop in the fish when exposed to lethal concentrations
of blue-green algae.
A question arises as to symptoms (specific or nonspecific) that may
serve as indicators when determining biologically dangerous concentrations?
Oviously, specific changes, and, in the present case, changes in the thia-
minase activity and the total thiamine content under the influence of blue-
green algae, must serve as indicators. Judging from our experimental data,
biologically dangerous concentrations of blue-green algae must range from
0.3 to 0.6 g/liter of raw substances.
However, it is important to remember that biologically dangerous concen-
trations of blue-green algae may change, depending on the effect of the
algae's natural metabolites and synergism or antagonism with other biotic
or abiotic water substances. In particular, the toxicity of blue-green
algae will depend both on a series of chemical indices (temperature effect,
content of oxygen in water, carbon dioxide, the presence of salts of such
metals as manganese, zinc and lithium) and on the physiological state of
algae cells (living, dead, decomposing). Thus, in our experiments, decom-
posing algae proved to be more toxic for fish.
The nature of an aquatic organism's reaction to the algae is important.
Thus, predators are the first to react to the algal toxicant effect be-
cause they are organisms characterized by a more intensive metabolism and
belong to the final link in the trophic chain. When estimating natural
toxicants, it becomes necessary to consider indicator organisms.
We have data showing that anlysis of the biological toxin effect re-
quires that we examine indicator organs which change more appreciably and
begin to show changes at an earlier period of time. In experiments investi-
gating the effect of lethal concentrations of blue-green algae on fish, the
liver may be considered such an organ. The reversibility of fish intoxica-
tion is a very important problem for man and animals.
No doubt, several metabolic changes observed in threshold concentrations
are reversible. Judging from our observations, even the changes in fish
metabolism which occur in the fish under the influence of lethal concentra-
tions of blue-green algae are reversible at early stages. Namely, respira-
tion and several biochemical indices in fish transferred to pure water be-
come normal. It is assumed that in the absence of sizable algal concentra-
tions, metabolic processes in fish organisms will be normalized since algal
concentrations may not only increase but also decrease due to the fact that
the wind concentrates or disperses them in the water body. The important
problem is the degree that normalization will affect the enzyme systems,
and whether the thiaminase activity is lowered enough to avoid Gaff's
disease if the fish is consumed by man or other animals. According to our
data (Birger, Malyarevskaya, Arsan, 1972) the disease is an acute B^avita-
minosis.
113
-------
In addition, since each lethal concentration needs a definite period of
time for action, short-term effects of even considerable algal concentra-
tions do not always result in the fishes' death, but can change their meta-
bolism and that will affect in the future the species reproduction.
A considerable effect of algal concentrations in the environment on the
metabolism indices of some fish species is seen in an example of blue-green
algae "bloom" stimuli. As a result of the investigations, threshold concen-
trations of algae accumulation in the environment have been determined. Ex-
ceeding those threshold concentrations appreciably results in negative
effects of algae on vital activity of fish even in an environment completely
deprived of other chemical pollutants.
REFERENCES
Braginskiy, L.P., S.L. Gusynskaya, I.M. Papchenko, A.M. Litvinova, and A.F.
Sysuyeva. 1965. Toxic and bactericide properties of extracts from
plankton blue-green algae. Vopr. Gidrobiol. Is'zd. Vses. Gidrobiol.
OB-VA. M. Izd. "Nauka".
Birger, T.I., A.Ya. Malyarevskaya, and O.M. Arsan. 1973. On the etiology
of Gaff's (Yuksov-Sartlan) disease. Gidrobiol. Zh. 2.
Goryunova, S.V. 1966. Formation of algae, their physiological role and
influence on the general regime of water bodies. Gidrobiol. Zh. 24.
Komarovskiy, F-Ya. 1970. On several pathological changes in fish under
the influence of blue-green algae". Gidrobiol. Zh. Vol. VI, 2.
Malyarevskaya, A.Ya., T.I. Birger, O.M. Arsan, and V.D. Solomatina. 1973.
Influence of blue-green algae on fish metabolism. "Naukova Dumka"
Kiev.
Metelev, V.V. 1971. Toxicity of pesticides for fish, effect mechanism,
indicator methods. Ekspep. Vodn. Toksikologiya, Vol. 2. "Zinatne"
Riga.
Sirenko, L.A. 1972. Physiological bases for reproduction of blue-green
algae in water reservoirs. "Nauka Dumka" Kiev.
Stroganov, N.C. 1972. Scientific bases for establishing the MPC for toxic
substances in open water bodies. Biolog. Aspec. Scientific Bases for
Establishing MPC in the Water Environment and Self-Purification of Sur-
face Waters.
114
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SECTION 11
TOXIC ORGANIC RESIDUES IN FISH
Howard E. Johnson
INTRODUCTION
The distribution of synthetic organic chemicals in the environment has
emerged as a major problem of industrialized nations throughout the world.
The discovery of widespread environmental contamination by DDT and dieldrin
has led to the recognition of many other environmental contaminants includ-
ing such industrial chemicals as polychlorinated biphenyls (PCB), pthalate
esters, and hexachlorobenzene.
The development of sophisticated analytical techniques and intensified
chemical monitoring efforts has shown that a wide variety of synthetic or-
ganic chemicals or their degradation products is present in the aquatic en-
vironment. As many as 40 potentially hazardous chemicals have been identi-
fied in some rivers that receive domestic and industrial effluents
(Kleopfer and Fairless, 1972; Hites, 1973). Significant contamination may
also be occurring in some regions because of chemical fallout from the
atmosphere. The ecological and public health hazard of these contaminants
is largely unknown, but the potential effect is considerable.
PRODUCTION AND DISTRIBUTION
The problem of environmental contamination is increased because of the
magnitude of commercial chemical production. The United States has in-
creased its production of chemicals by nearly 10% a year with present pro-
duction exceeding 140 billion pounds. As many as 500 new chemicals are
produced each year with little or no knowledge of the potential hazard of
their behavior in the environment (Lee, 1964). Some compounds have highly
toxic, carcinogenic, or mutagenic properties that may be especially
damaging if they are accumulated in aquatic systems. In some instances the
degradation products or metabolites may be of equal or greater consequences.
Aquatic ecosystems are especially vulnerable to the effects of chemical
pollutants. Acutely toxic concentrations resulting from accidental spills
or direct application have caused extensive fish kills over broad areas of
the environment, but many chemicals occur in the environment at concentra-
tions that are not directly lethal to fish. These compounds are distri-
buted as microcontaminants, i.e., concentrations of a few parts per million
115
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or less, in various substrates within the aquatic ecosystem. Because of
their very low concentration, such chemical contaminants may not be
detected until they appear in undesirable levels within some trophic level.
Serious ecological damage or effects on human health have sometimes
occurred before we have taken action to prevent further contamination. An
example is the environmental mercury problem, which first received world-
wide attention when human beings were poisoned by eating contaminated fish
and shellfish during the 1950's in Minamata, Japan.
EFFECTS ON FISHERY RESOURCES
Synthetic organic chemical residues that accumulate in aquatic
organisms can have far-reaching effects on an entire fishery resource. In
particular, the problem is exemplified by the impact of polychlorinated
biphenyls (PCB) on fishery resources of the Great Lakes."
Occurrence and Accumulation
Polychlorinated biphenyls are a widely used class of chlorinated
hydrocarbons found in a variety of manufactured products and in many
industrial processes. Environmental monitoring has shown that PCB are
distributed throughout the Great Lakes ecosystem, but the highest
concentrations are generally found near industrial and urban area. Some
specific industries are known to discharge PCB in their effluents, but
non-specific sources such as municipal wastewater effluents are more
difficult to control. Atmospheric contributions in the form of rain, snow,
and particulate fall-out also may be significant.
Concentrations of PCB in the Great Lake waters are generally only a few
parts per trillion (nanograms per liter), but because of biological concen-
tration residues in some fish exceed 20 parts per million (milligrams per
kilogram). Laboratory studies indicate that fish can accumulate PCB by
more than 40,000 times the exposure concentration (Stalling and Mayer,
1972). The residues are most concentrated in the lipids of body tissue.
Careful monitoring studies have shown that residue concentrations vary with
different species in proportion to their fat content. The highest concen-
trations are found in mature salmon and trout just before or during their
spawning migration (Veith, 1975).
It has been suggested but not proven that PCB and other chemical
residues accumulated in the eggs are responsible for high mortalities of
some fish during the early stages of development. Mortality of young
salmon has been high where the eggs contained PCB, DDT, and some other
chemical residues (Johnson and Pecor, 1969; Halter and Johnson, 1974).
More recently PCB are suggested as the cause for losses of northern pike
(Esox lucius) embryos in Michigan hatcheries (Waybrant, 1975). In Sweden
Jensen, Johansson, and Olson (1970) suggested a correlation between PCB
residues and mortality of salmon eggs and fry.
116
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Multiple Residues
The difficulty of assessing the effects of chemical residues on wild
populations of fish is compounded by the simultaneous occurrence of several
contaminants. Potentially harmful levels of DDT, dieldrin, PCB, and other
toxic chemicals are present in the Great Lakes ecosystem. These multiple
residues present analytical uncertainty as well as potential additive
effects on aquatic populations.
Prior to 1970 PCB residues in Great Lakes fish were not identified and
in many cases were probably mistakenly included in the results given for
pesticide residues. The use of gas chromatography-mass spectrometry has
improved our ability to identify contaminants, but the procedure remains
difficult and expensive. A much more complex problem is the interpretation
of the ecological significance or hazards associated with exposure to
multiple residues.
Some research indicates a "synergistic" or more-than-additive toxic
action between PCB and certain pesticides. Joint action of PCB and DDT was
found in chronic exposure tests with Daphnia magna (Maki and Johnson, 1975).
Toxicity tests with insects have also shown joint action between PCB and
some carbonate and organochlorine insecticides (Lichtenstein e^t aj_., 1969;
Plapp, 1973).
Thus, we find that some compounds occur in the environment at concentra-
tions that are known to have adverse effects in laboratory tests. Where
these levels are not directly lethal, effects on growth or reproduction may
be expressed as slow changes in the size and abundance of fish populations.
Therefore, although it appears likely that adverse effects are occurring,
it is very difficult to show this conclusively.
Effects on Higher Trophic Levels
Toxic organic chemical residues present a serious hazard to consumer
organisms at the higher trophic levels, including man. The biological
accumulation of residues and their trophic levels are especially hazardous
to animals that utilize fish as a major food source.
Piscivorous birds in the Great Lakes region have suffered unnaturally
high mortalities in recent years, and depressed reproduction of some popula-
tions has been correlated with chemical residues (Hesse, 1975). The high
residue concentrations of PCB in gulls in some regions of the Great Lakes
may seriously threaten these bird populations, but the full extent of the
problem is unknown.
Even before PCB and pesticide contamination in the Great Lakes was re-
cognized, fur farmers in the region reported reduced reproduction of mink
that were fed with Great Lakes fish. Surplus coho salmon or salmon by-
products caused death or reproductive failure of mink when these products
formed 30% of their diet (Aulerich, Ringer, and Iwamoto, 1973). The rela-
tively high concentrations of DDT, dieldrin, and PCB in the fish were sus-
117
-------
pected as the causative agents. Subsequent laboratory tests in which mink
were fed various doses of DDT and dieldrin in excess of the levels found in
the fish did not reproduce the effects. However, 5 ppm PCB added to the
experimental diet markedly reduced reproduction, and 15 ppm totally
inhibited reproduction and caused death of the adults (Ringer, Aulerich,
and Zabik, 1972). These tests established that mink are highly sensitive
to PCB toxicity and clearly indicated that residues accumulated in Great
Lakes fish were responsible for death and reduced reproduction of commer-
cially reared mink. Because of the high losses resulting from feeding coho
salmon, fur farmers have discontinued the use of Great Lakes fish in mink
diets.
The residues of PCB in Great Lakes fish pose a potential health hazard
to humans. To protect consumers the U.S. Food and Drug Administration has
restricted the distribution and sale of fish that contain more than 5 ppm
PCB; shipments of such fish from commercial outlets have been confiscated
and destroyed. This ruling has curtailed the commercial utilization of
most major food fish species in the Great Lakes. Although recreational
fisheries are not restricted, state health authorities have warned sport
fishermen to limit their consumption of Great Lakes fish. A new informa-
tion on PCB effects is developed, greater restrictions may be necessary.
NEW TEST PROCEDURES
The problems currently associated with PCB in the Great Lakes are only
a single example of the serious impact of synthetic organic chemicals on
aquatic ecosystems. Residues of other potentially harmful chemicals (e.g.,
hexachlorobenzene, the chlorinated napthalenes, pthalate plasticizers) have
been found in increasing concentrations in aquatic systems. Clearly there
is a need to identify and restrict the distribution of harmful residues
before serious damage has occurred.
Industrialized nations throughout the world have a responsibility to
develop new strategies for identification and control of harmful chemicals.
We can neither afford to wait to study these problems after contamination
has occurred, nor can we afford the time and resources to thoroughly
investigate each new chemical before it is released to the environment. It
is imperative that we develop a systematic approach for evaluation of new
materials and new technology. Important new efforts are being made to find
correlations between chemical structure and biological activity (Veith and
Konasewich, 1975). A chemical classification system based on physical pro-
perties, chemical structure, and biological activity would provide some
indication of potential hazard. Simple model ecosystems (Metcalf, Sanga,
and Kapoor, 1971) and food-chain models (Johnson, 1974) offer additional
promise for preliminary testing to identify harmful properties of chemicals.
118
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REFERENCES
Aulerich, R.J., R.K. Ringer, and Susumo Iwamoto. 1973. Reproductive fail-
ure and mortality in mink fed on Great Lakes fish. J. Reprod. Pert.,
Suppl. 19: 365-376.
Halter, M.T., and H.E. Johnson. 1974. Acute toxicities of a polychlori-
nated biphenyl (PCB) and DDT alone and in combination to early life
stages of coho salmon (Oncorhynchus kisutch). J. Fish. Res. Board Can.
31: 1543-1547.
Hesse, J.L. 1975. Contaminants in Great Lakes fish. Staff Report, June
1975. Michigan Water Res. Comm., Dep. Nat. Resour., Lansing, Michigan.
15 p.
Hites, R.A. 1973. Analysis of trace organic compounds in New England
rivers. J. Chromatograph. Sci. 11: 570-574.
Jensen, S., N. Johannsson, and M. Olson. 1970. PCB-Indications of effects
on salmon. Swed. Salmon Res. Inst. Rep. LFI Meod. 7/1970.
Johnson, B.T. 1974. Aquatic food chain models for estimating bioaccumula-
tion and biodegradation of xenobiotics. Proc. Int. Conf. on Transport
of Persistent Chemicals in Aquatic Ecosystems. National Research Coun-
cil of Canada, Ottawa.
Johnson, H.E., and C. Pecor. 1969. Coho salmon mortality and DDT in Lake
Michigan. Trans. 34th N. Am. Wild!. Natl. Resour. Conf. p. 159-166.
Kleopfer, R.D., and B.J. Fairless. 1972. Characterization of organic com-
ponents in a municipal water supply. Environ. Sci. Tech. 6: 1036-1037.
Lee, D.H.K. 1964. Environmental health and human ecology. Am. J. Public
Health 54 (Suppl.): 7-10.
Lichtenstein, E.P., K.R. Schultz, T.W. Fuhremann, and T.T. Liang. 1969.
Biological interactions between plasticizers and insecticides. J. Econ.
Entomol. 62: 761-765.
Maki, A.W., and H.E. Johnson. 1975. Effects of PCB (Aroclor 1254) and p,p'
DDT on production and survival of Daphnia magna Strauss. Bull. Environ.
Contain. Toxicol. 13: 412-416.
Metcalf, R.L., G.K. Sanga, and I.P. Kapoor. 1971. Model ecosystems for the
evaluation of pesticide biodegradability and ecological modification.
Environ. Sci. Tech. 5: 709.
Plapp, F.W., Jr. 1973. Polychlorinated biphenyl: An evironmental contami-
nant acts as an insecticide synergist. Environ. Entomol. 1: 580-582.
119
-------
Ringer, R.K., R.J. Aulerich, and M. Zabik. 1972. Effect of dietary poly-
chlorinated biphenyls on growth and reproduction of mink. 164th
National Meeting American Chemical Society. 12: 149-154.
Stalling, D.L. and F.L. Mayer, Jr. 1972. Toxicities of PCBs to fish and
environmental residues. Environ. Health Perspect. 1: 159-164.
Veith, G.D. 1975. Baseline concentrations of polychlorinated biphenyls
and DDT in Lake Michigan fish - 1971. Pest. Mon. 9: 21-29.
Veith, 6.D., and Dennis E. Konasewich (ed.). 1975. Structure-activity
correlations in studies of toxicity and bioconcentrations with aquatic
organisms. Proc. Symposium Intern. Joint Comm., Windsor, Ontario, Can.
Waybrant, R.C. 1974. Northern pike fry mortalities attributed to poly-
chlorinated biphenyls. Staff Report. Mich. Bur. Water Manage., Mich.
Dep. Nat. Res., Lansing, Mich.
120
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SECTION 12
BALANCE OF ORGANIC MATTER IN THE
ECOSYSTEM OF THE RYBINSKIY RESERVOIR
V.I. Romanenko^
The Rybinskiy reservoir and its drainage basin are located in the south-
ern "taiga" zone, within the boundaries of three districts: Yaroslavskiy,
Vologodskiy and Kalininskiy. It was constructed in 1941, and it is one of
the largest man-made water bodies in the world with a surface area of 4450
km2, water volume of 25.4 km3 and mean depth of 5.6 m. Water inflow,
according to data taken over a period of several years, is around 37 km3/-
year. The Volga, Mologa and Shekana River inflows amount to 2/3 of total
inflow, with the rest supplied by small rivers.
The reservoir freezes in November and thaws (melts) in April or the be-
ginning of May. According to Secchi disc readings, transparency during the
summer is 1-3 m. The content of particles in water is 3-7 nig/liter. The
water is of bicarbonate-calcium type according to chemical analysis. The
pH is 7.0-7.5, content of organic matter is around 15 mg C/liter, total N
is 1.2 mg N2/liter, total phosphorus is 0.04 mg P/liter, and bicarbonate is
10-20 mg C/liter (monograph "Rybinskiy reservoir and its life", 1972).
Data presented below are based on results of long term observations,
for some parameters 5-10 years, up to 20 years for others. Analyses were
carried out at 15 day intervals from May through November at six stationary
stations distributed along the base area of the reservoir (Figure 1).
INPUT OF ORGANIC MATTER
In the Rybinskiy reservoir, as in all water bodies, there are two basic
sources of organic matter: Internal (authocthonous) and external (alloch-
thonous). The photosynthetic production of organic matter by phytoplankton
and macrophytic vegetation, which is not large in this reservor (Belavskaya
and Kutova, 1963) (approximately 1.8% in the input balance), are the main
sources of authocthonous organic matter. Diatoms (Melosira and Aterionella)
Institute for Biology of the Inland Waters, Academy of Sciences, USSR.
121
-------
Cherepovets
Breytovo )
Borok
Fig. 1. Map-diagram of Rybinskiy reservoir. Figures indicate
number and location of standard stations.
122
-------
are the dominant species of phytoplankton in spring and autumn, but during
July and August the blue-green algae are predominant.
Primary productivity of organic matter was determined by the C method
(Steeman Nielsen, 1952). The intensity of photosynthesis in the water mass
correlated well with water transparency according to Secchi disc readings
(Romanenko, 19/3a) according to the formulas:
Fm = F^ x 0.7 x 3 1 x 1000, where
Fm = Phytoplankton primary productivity/m2/24 hours
F.J = Phytoplankton primary productivity in a water sample
integrated according to depth, exposed at surface
illumination during 24 hours.
1 = Secchi disc transparency.
The gross phytoplankton primary productivity (mean for several years)
ranged from 100 to 500 thousand ton C for the whole water body; for the sur-
face it changed from 30 to 150 g C/m2 (Figure 2), with a mean of 76 g C/m2.
Long-term (meaning for several years) fluctuations in the intensity of
photosynthesis depended on meterological factors during the year, rate of
outflow, and an increase in the reservoir drainage zone. Estuaries of
large rivers are the most productive areas. As a rule, the primary produc-
tivity is 1.5-2.0 higher than in the central area. In all, 0.05-0.20%
energy from solar radiation penetrating the water is used by the algae.
Solar radiation is used less effectively during the spring when there are
few phytoplankton and the water temperature is low (0.02-0.07%). During
the blue-green algae bloom in July and August, given the highest water temp-
eratures, it is used considerably more effectively (0.37%).
By using the data of chemical oxygen demand (COD) and the water balance,
the total input of allochthonous organic matter was calculated by Kuznetsev
and Bezler (1971). According to their data, 4.75 x 103 ton C as organic
matter enters the reservoir with the melted ice and snow. Since this reser-
voir is located in a low populated region, and the pollution of it as a re-
sult of man's activity is not great, the input of organic matter by such
sources as wastewaters can be neglected. The input of organic matter due
to atmospheric precipitations, in particular in winter with snow on the re-
servoir surface, is 680 ton C, which is only 0.5% of the value of primary
productivity (Romanenko and Bezler, 1971). An additional input of organic
matter is also due to rainfall precipitation during summer months. There-
fore, it is necessary to consider all these sources of organic matter to-
gether, because each of them isolated represents a very small input of allo-
chthonous organic matter.
The total bacterial assimilation of C02 and hydrocarbonate (hetero-
trophic assimilation of C02 and chemosynthesis) were determined by the C
123
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150-
w IOO
£
50
O
Fig. 2. Average values for phytoplankton production for 13 years
in the Rybinskiy reservoir.
124
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method for the whole area of the reservoir during the navigable period. It
is not high, (9600 ton C), which means 1.2% of the total organic matter
input.
EXPENDITURE OF ORGANIC MATTER
From the very beginning, a part of the organic matter is expended by
primary productivity, i.e., by phytoplankton. According to our data, this
is equal to 20% of the primary production for 24 hours incubation. When
analyzing the destruction of organic matter by the oxygen method in the
water mass, the given value is a general sum of the organic matter des-
troyed by all plankton organisms. Due to the low mean depth and the severe
wind regime, the water mass of the Rybinskiy reservoir is very well mixed
to the bottom, and the organic matter is mostly oxidized in aerobic condi-
tions by bacterial action. The average number of bacterioplankton by
direct counting (Fazumov, 1932) for 20 years of investigation is 1.5 mill/
ml and ranged between 0.5-2.5 mill/ml in different years (Figure 3).
By cultivating a sample of bacteria from the reservoir in sterile re-
servoir water, it is possible to prove that the amount of bacteria deter-
mined by direct count is true. In this medium, the bacteria prepared from
the water grow very quickly, and it is possible to determine the limits of
their development very easily (Romanenko, 1973b).
The bacterial cell volumes in the water are small (0.3-0.5 y3). The
horizontal and vertical distribution of bacterial biomass is uniform in the
reservoir, and the total number of bacteria changes little from one point
to another. Only near the shores and in the bottom layers is it a little
higher. Wet bacterial biomass is equal to 1-2 mg/liter of water.
The generation time, i.e., time needed for the total number of bacteria
to double, fluctuated within wide limits and ranged from 5 to 10 hours for
individual periods. In the hottest period (July and August) with a doubling
in the number of bacteria, generation time is between 16-20 hours; but for
temperatures 5-13 C, it can be up to hundreds of hours. For 20 years of in-
vestigation, the mean value for the navigable period was 48 hours with temp-
erature fluctuations of 2 to 23 C.
Data on heterotrophic assimilation of C02 was used for calculating bac-
terial biomass (Romanenko, 1964). Its mean value was 35 g C/m2 or 145 x 103
ton C on the whole reservoir (Table 1).
125
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If) (0
0)O —
O—
Fig. 3. Average values of the number of bacteria according to
data for 2 years of standard observations.
126
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TABLE 1. PRODUCTION OF BACTERIAL BIOMASS IN THE RYBINSKIY RESERVOIR
(in C)
During the navigable period (May-November)
Year
1964
1965
1966
1967
1968
Mean
Number
of
days
140
158
151
150
187
157
In 1
1 i ter ,
mg
6.7
10.6
7.4
5.3
3.6
6.7
Under
1 mg2,
g
33
55
39
28
21
35
For the
whole
reservoir
ton
1 1 7000
231000
1 74000
1 1 7000
86000
145000
Mean for 24 hours
In 1
liter,
mg
0.047
0.067
0.049
0.035
0.019
0.043
Under
1 mg2,
g
0.24
0.35
0.26
0.19
0.11
0.23
For the
whole
reservoir
ton
836
1460
1150
780
460
937
A very strong purification effect is produced by this bacterial biomass.
A great amount of reduced organic compounds is oxidized through this effect,
where the compounds are oxidized by oxygen in enzymatic reactions. Accord-
ing to the data, in essence we did not observe a significant increase in
bacterial biomass. This total value for a large amount of time and for
each short time interval is not very great.
The destruction of organic matter was determined on the basis of oxygen
consumption in dark bottles during the incubation of samples for 24 hours
at ambient water temperature (Table 2). From this table, it is possible to
see that this value for the whole reservoir for the navigable period ranged
from 270,000 to 950,000 ton C of organic matter decomposed, with a mean
value of 550,000 ton C for 5 years of observations (129 g C/m2) and a fluc-
tuation between 64-214 g C/m2 in different years.
TABLE 2. DESTRUCTION OF ORGANIC MATTER IN THE WATER
(in C)
During the navigable period (May-November)
Year
Number
of
days
In 1
1 i ter ,
mg
Under
1 mg2,
g
For the
whole
reservoir,
ton
Mean for 24 hours
In 1
1 i ter ,
mg
Under
1 mg2,
g
For the
whole
reservoir
ton
1958
1965
1966
1967
1968
Mean
135
162
167
154
194
162
18
20
40
28
11
23
101
116
214
150
64
129
450000
480000
950000
633000
270000
556000
0.13
0.12
0.24
0.18
0.06
0.14
0.75
0.72
1.28
0.97
0.33
0.91
3340
2960
5670
4110
1390
3494
127
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If we take into account that the rest of the population destroys only
around 20% of the total organic matter (especially algae and invertebrates),
it is easy to calculate the correlation between bacterial biomass and de-
struction. The production of bacterial biomass was 30% of the amount of
destroyed organic matter (expressed as carbon). By means of pure cultures
it was demonstrated that the correlation between these data (Kj coefficient)
ranged from 25-35%, which proves the reality of our calculation of bacterial
production by means of the data on heterotrophic assimilation of C02.
Also in the bottom sediments a great amount of organic matter is decom-
posed in aerobic or anaerobic conditions according to the redox potential
through the activity of aerobic or anaerobic organisms or both. In this re-
servoir, the bottom layers are very well supplied with dissolved oxygen,
and oxygen deficits are only observed in the flood areas in which the rH2
fluctuates from 10-12, and sulfate reduction and methane formation pro-
cesses are most intense. The rH2 is 10-17 in the peat slimes and 5-20 in
the sandy ones with small amount of debris. In such conditions aerobic de-
struction of organic matter prevails.
The mean number of bacteria by direct count is 0.5-30 billion/g of wet
slime. Fifteen years ago, when the amount of organic matter from the re-
cently flooded areas was high enough, a very strong production of methane
gas derived from the activity of methane-producing bacteria was observed
and, especially in winter season, the big bubbles of this gas induced the
death of fish by asphyxiation. The methane formation process at this time
has sharply decreased.
To determine the destruction of organic matter in the bottom sediments,
the oxygen consumption of a column of slime isolated in a glass tube (Hayes
and MacAulay, 1959) was used, and the balance of C02 extracted from the
slime into the water (Romanenko and Romanenko, 1969) was employed to deter-
mine the aerobic decompositon of organic matter.
The results showed that 74,000 ton C as organic matter are decomposed
in the slime by aerobic processes during the navigable period, which means
20 g C/m2. Nearly 10 g C/m2 are decomposed in anaerobic conditions. This
means that 23% of the organic matter is destroyed in the water mass.
The loss of organic matter through the lower outlet was calculated ac-
cording to oxidizability and water balance as 179,000 ton C by Kuznetsov
and Bezler (1971).
For the balance (Table 3) of the input and expenditure of organic
matter in the ecosystem of the Rybinskiy reservoir, taking into account
only the most important parameters, we can say that as a whole, in this eco-
system, the input of organic matter is 826,000 ton C or 199.9 g C/m2.
By bacterial destruction or outlow through the outlet, the expenditure
of organic matter is 838,000 ton C, so the difference is -8400 ton C. If
we discount the loss of organic matter through the outlet (837,000 -
179,000), it is possible to see that the purification effect due to the
128
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TABLE 3. MAIN ELEMENTS OF THE BALANCE OF ORGANIC
MATTER IN THE RYBINSKIY RESERVOIR
For the whole Under
Element reservoir x 103, 1m2, % Note
ton C g C
Input
Gross production of
phytoplankton
Production of
macrophytic
vegetation
Bacterial assimila-
tion of C02
Input of organic matter
from the watershed and
atmospheric precipita-
tions
Total Expenditure
Destruction in
water in summer
Destruction in
water in winter
Aerobic destruction
in slime in summer
Aerobic destruction
in slime in winter
Outflow of organic matter
through the outlet
Total
Difference between input
330
2
14
9.6
475
516
39
84
19
179
837
-8.4
78 39 Mean for 11 years
(from 1955 to 1970)
During 1956, 57, 62
and 63 the analyses
were not done
3.5 1.8 According to data
for 1965 (Belav-
skaya and Kutova,
1966)
2.4 1.2 Mean for 5 years
(1964-1969)
116 58 According to data
for 1965
129 62 Mean for 5 years
(1958,1965,1968)
9 4.3 According to data
for 1965
21 10 Mean for 2 years
(1967, 1968)
5 2.4 Calculated accor-
ding to the coeffi-
cient of Van Hoff
(K = 3)
44.7 21 According to data
for 1965
208.7 100
-8.8
and expenditure of
organic matter
129
-------
activity of the population of water and slime (fundamentally microorganisms)
is 658,000 ton C as organic matter.
Of course, we know that the accuracy of the data depends on the analyti-
cal possibilities, and because of this we are inclined to examine such a
good agreement between the input and expenditure of organic mater as possi-
bly affected by some random element. Nevertheless, the main sources of
input and expenditure of organic matter in such a large and complex man-made
reservoir as the Rybinskiy reservoir are clear.
Only a very small amount of matter and energy, in comparison with the
amount involved in the fundamental processes, flows through the rest of the
population of the reservoir. Taking into account the data from the biomass
and cycles of life of the different animals (Yu.I. Sorokin, see monograph
"Rybinskiy reservoir and its life," 1972), in the second-trophic level
(herbivorous, detritous protozoa, rotatoria and Crustacea) nearly 20% of
the incoming organic matter is utilized. In the third trophic level (preda-
tor zooplankton, zoobenthos and fish larvae) only 0.2% is used and at least
as much for forage and predator fish. For man, it reaches only 0.05%. Of
course, it is necessary to remember that the last values expressed as ab-
solute units are fairly inexact because the calculation of such parameters
as species composition, zooplankton biomass and life cycles are only approx-
imated. So, as useful products on man's table, he receives only tenths or
hundreths of parts of the initial input of organic matter, and a large part
of it is utilized by bacteria.
REFERENCES
Belavskaya, A.P. and Kutova, T.N. 1966. The vegetation of the ephemeral
flood areas of the Rybinskiy reservoir. Publications of the Institute
for the Biology of Inland Waters, USSR Academy of Sciences. Vol 11(14)
Print "Nauka", M.-L.
Hayes, F.R. and MacAulay, N.A. 1959. Lake water and sediment. V. Oxygen
consumed in water over sediment cores. Limnology and Oceanography, V.
4, N. 3.
Kuznetsov, S.I. and Bezler, F.I. 1971. An experiment for determining the
balance of organic matter in the Rybinskiy reservoir. In: "Biology
and productivity of freshwater organisms". Publications of the Insti-
tute for the Biology of Inland Waters. USSR Academy of Sciences, Vol.
21(24). Print "Nauka", M.-L.
Romanenko, V.I. 1964. Heterotrophic assimilation of C02 by the bacterial
microflora of the water. Mikrobiologia XXXIII, Vol. 4.
Romanenko, V.I. 1973a. Interrelation between the intensity of photosynthe-
sis of a population of algae uniformly distributed through the water
column and Secchi disc transparency. Information Bulletin of the Insti-
tute for the Biology of Inland Waters, USSR Academy of Sciences. No.
19. Print "Nauka", L.
130
-------
Romanenko, V.I. 19736. A new method for determining living bacteria in
water bodies and a comparison with the Razumov method. Information
Bulletin of the Institute for the Biology of Inland Waters. USSR
Academy of Sciences. No. 22. Print "Nauka", L.
Romanenko, V.I. and Bezler, F.I. 1971. Chemical and microbiological analy-
sis of the snow over the ice of Rybinskiy reservoir. Information Bulle-
tin of the Institute for the Biology of Inland Waters. USSR Academy of
Sciences. No. 11. Print "Nauka", L.
Romanenko, V.I. and Romanenko, V.A. 1969. Destruction of organic matter in
the slime of the Rybinskiy reservoir. In "The physiology of aquatic
organisms and their role in the cycle of organic matters". Publications
of the Institute for the Biology of Inland Waters. USSR Academy of
Sciences. Vol. 19(22). Print "Nauka", L.
Sorokin, Yu.I. 1972. Chapter "Biological Productivity" in the monograph
"Rybinskiy reservoir and its life". Print "Nauka", L.
Steeman, Nielsen E. 1952. On the use of radioactive carbon (llfC) for
measurement of organic production on the sea. J. Cons. Exp. Mer. V.
18.
131
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SECTION 13
THE IMPORTANCE OF TROPHIC BONDS IN THE
BACTERIAL DESTRUCTION OF ORGANIC MATTER
P.P. Umorinl
In connection with the problem of purifying industrial and domestic
wastewaters and ascertaining the role of bacteria as a main factor in puri-
fying the dissolved organic matter (DOM), it is necessary to show the role
of organisms that feed on bacteria during the process. Unfortunately, work
carried out in this field and concerned with the role of the protozoa is
contradictory (Kryuchkoya, 1968). In the experiments of a great number of
authors (Butterfield, Purdy, Theriault, 1931; Phelps, 1953; Javorinsky,
Prokesova, 1963; Nikolyuk, 1965; Straskrabova-Prokesova, Legner, 1966;
Jensen, Ball, 1970), the oxygen consumption and nitrogen fixation were more
intensive in mixed cultures of bacteria and protozoa than in pure cultures
of bacteria. The results of these and other authors serve as a foundation
for the hypothesis first put forward by Butterfield (Butterfield, Purdy,
Theriault, 1931) that organisms preying upon bacteria keep the latter in a
state of continuous reproduction or physiological "youth" by a simple de-
crease of their number. This should facilitate a greater rate of decom-
position of organic matter. The above-mentioned authors, however, do not
analyze the correlation between the actual conditions of the experiment and
its results. In addition, the role of separate species and the quantitative
characteristics of the bacteria and protozoa relationships are far from
clear. This is especially true of the colourless flagellates. Their role
in the life of water bodies has almost not been studied. The aim of this
work is to study the rate of organic matter decomposition with bacteria and
protozoa in continuous cultures, maximizing the similarity to natural condi-
tions.
METHODS
All the experiments have been performed using a continuous culture de-
vice described previously (Umorin, 1975) with various dilution rates (D).
Pratt's solution served as the nutrient medium. Phenol at a concentration
of 25 nig/liter or glucose at a concentration of 50 mg/liter was added to
the solution as the only source of organic carbon. Before the experiments,
the reactor, 1.2 liter in volume, was filled with the nutritive medium and
innoculated with bacteria taken from a water-body and previously kept in a
Institute of Biology of Inland Waters of the Acad. Sci. USSR.
132
-------
continuous culture on the Pratt medium with an organic matter (phenol or
glucose) used in the experiment. Then a continuous run was started with
the required dilution rate. When a steady state (constant concentration of
organic matter and constant number of bacteria) was reached, protozoa were
added to the reactor and the experiment was continued until a new steady
state was attained. Thus, the first part of the experiment (organic matter
+ bacteria) served as a control for the second part (organic matter +
bacteria + protozoa). As for the protozoa, the infusoria ciliates
Paramecium caudatum, Ehrb. were used with the experiments with phenol, and
the zooflagellates Pleuromonas jaculans, Perty, in the experiments with
glucose.
To clarify the interrelations of bacteria and protozoa in a medium de-
ficient in nutrients, experiments have also been performed with infusoria
ciliates P_. caudatum in Pratt solution in which the concentration of nit-
rate nitrogen was diminished to 1 mg/liter. All the experiments were con-
ducted at a temperature of 22 C. Altogether, the following experimental
variables have been used:
1. Bacteria on complete Pratt medium with phenol at D = 0.02 hrs"1.
2. Bacteria and infusoria on complete Pratt medium with phenol
at D = 0.02 hrs"1.
3. Bacteria on complete Pratt medium with phenol at D = 0.04
hrs-1.
4. Bacteria and infusoria on complete Pratt medium with phenol
at D = 0.04 hrs l.
5. Bacteria on pratt medium given deficient nitrogen with phenol
at D = 0.02 hrs 1.
6. Bacteria and infusoria on Pratt medium given deficient nitrogen
with phenol at D = 0.02 hrs"1.
7. Bacteria on complete Pratt medium with glucose at D = 0.06 hrs"1.
8. Bacteria and flagellates on complete Pratt medium with glucose
at D = 0.06 hrs'1.
9. Bacteria on complete Pratt medium with glucose at D = 0.08 hrs"1.
10. Bacteria and flagellates on complete Pratt medium with glucose
at 0 = 0.08 hrs"1.
All the variables of the experiment were conducted with three replica-
tions. Every day the numbers of bacteria and flagellates were determined
in the reactor by direct counts in the Goryavev chamber at microscope magni-
fication of 900x for bacteria and 150x for flagellates. Infusoria were
counted in the Bogorov chamber under a MBS-1 microscope. To see the
balance of organic matter, the numbers of organisms were expressed in dry
133
-------
weights in rag/liter. The latter were calculated from their biovolume;
specific gravity of the wet biomass was accepted as an equal unit and the
dry weight - 15% of the biomass. The mean volume of bacteria developing on
the medium with phenol was 1 y3, and on the medium with glucose 0.5 y3; in
this case, 1 mg of dry weight of bacteria per liter is equivalent to 7.5
million and 15 million cells per milliliter, respectively. With the mean
volume of P_. jaculans 25 y3, 1 mg of dry weight per liter corresponds to
the number 300 thousand cells per ml of this flagellate. The dry weight of
one infusorian P_. caudatum was taken from G.6. Vinberg's work (Vinberg,
1949), 0.08 x 10~3 mg. Then their dry weight of 1 mg/liter is equivalent
to the number 12.5 cells/ml. Concurrently with a calculation of the number
of organisms, the phenol concentration was determined by pyramidone method
(Lurie, Rybnikova, 1966), and that of glucose by a reagent with phenol
(Bikbulatov, Skopintsev, 1974).
RESULTS
In all the experimentally conditions, a steady component value was esta-
blished in the reactor, according to the type of damped oscillations in ex-
periments 2, 4, and 6, and aperiodically in the remaining ones. On the com-
plete Pratt media, the concentration of the organic matter in the reactor
was established at a higher level in the experiments with protozoa than
without them. This indicates that when mineral nutrients are provided and
the growth rate of bacteria is limited only by the concentration of organic
matter, the bacteria alone decompose it faster than in the mixed cultures
with protozoa. Since the number of bacteria in the presence of protozoa was
significantly lower than without them, the decrease in the rate of decom-
position of organic matter may be explained by the decrease in the number
of bacteria due to predation by protozoa. Nevertheless, in the experiments
with nitrogen deficiency, destruction of organic matter proceeded faster in
the presence of protozoa, in spite of the decrease in the number of bac-
teria; i.e., one and the same factor, protozoa, may either accelerate or
slow down destruction of organic matter depending upon the condition.
TABLE 1. VALUES OF THE COMPONENTS IN MG/LITER (DRY WEIGHT)
AT THE STEADY STATE (AVERAGE THREE REPETITIONS).
Experiment
number
1
2
3
4
5
6
7
8
9
10
D
hrs
0.02
0.02
0.04
0.04
0.02
0.02
0.06
0.06
0.08
0.08
Organic
Matter
4.9
19.6
10.4
20.6
23.2
21.5
6.7
11.0
8.9
9.5
Bacteria
11.1
0.36
7.5
1.6
0.89
0.83
12.2
6.3
11.5
10.6
Protozoa
1.2
_
0.8
_
0.39
_
1.9
_
0.3
Time of establ
steady state
8
36
8
24
30
30
4
25
4
25
ishing
(days)
134
-------
To reveal the character and quantitative aspect of interrelations be-
tween bacteria and protozoa in the process of destroying organic matter, it
is necessary to use mathematic models. We used models similar to Canale's
(1970) models, in which the dependences between specific growth rates and
feeding of organisms and the concentration of their food was linear. Such
a linearization has appeared to be quite acceptable for test conditions
allowing only a small change interval in the component values.
For the system—organic matter-bacteria:
=D(Si-S) -klSH
- = k2SH - DH
dt
For the system—organic matter-bacteria-infusoria
= D(SX-S) - kjSH - k3SP
-= k2SH - DH -
J= k5HP + k6SP - DP
(2)
where S - is the concentration of organic matter in the reactor
$! - the same, in the inflowing medium
H - concentration of bacteria in dry matter
P - the same, for protozoa
D - dilution rate
t - time
k.j - coefficients of proportinality.
For the system—organic matter-bacteria-zooflagellates, a model was con-
structed taking into account the possibility of feeding the zooflagellates
with the dissolved organic matter (DOM) (Goryacheva, 1975):
135
-------
^r= D(S,-S) - kjSH - k3SP
dH
dt
= k2SH - DH -
dP_
dt
k HP + k£SP - DP
5 6
'(3)
Here the term k SP is the consumption of DOM by zooflagellates and k6SP is
their growth (per unit time) due to DOM consumption.
The mathematical models (1) - (3) are not applicable for experiments
with nitrogen deficiency, since in this case, the concentration of organic
matter is not the factor limiting the growth of bacteria'. A system of equa-
tions which consider the role of nitrogen as the limiting factor has been
made up.
For the system—organic matter-bacteria:
dt
:}§• = k,NH - DH
dt 2
dt
- k3NH
(4)
For the system--organic-matter-bacteria-protozoa:
dJN
dt
dH
dt
dP_
dt
= DINj-N) - kjNH +
k2NH - DH - k5HP
k6PH - DP
°|= D(SrS) - k3NH
(5)
where N: is the concentration of nitrogen in the inflowing medium, N is the
concentration of nitrogen in the reactor, and the remaining designations
136
-------
are the same as in models (1) - (3). In models (4) and (5), the term
D(Nj-N) shows the influx of nitrogen into the reactor; kxNH is the net con-
sumption of nitrogen by bacteria (difference between consumption and exre-
tion); ki>P is the excretion of nitrogen by protozoa; k2NH and k3NH are the
growth of bacteria and the consumption of organic matter by the bacteria
per unit time.
To determine the rates of feeding and reproduction of the living com-
ponents, it is necessary to calculate the Kj coefficients. The coefficient
kj. and k2 were calculated by substituting into model (1) the steady-state
values of the components and equating the right members of the equations to
zero. For the experiments with phenol, they appeared to be ki = 0.008 and
k2 = 0.004; for the experiments with glucose, ki = 0.032 and k2 = 0.009*.
Coefficients k4 and k5 of model (2) and k3 and k4 of model (3) were cal-
culated by introducing into these systems the coefficients ki and k2 from
model (1). They were for model (2) k4 = 0.051, k5 = 0.024; for model (3)
kit = 0.02. Coefficients ks of model (3) was determined using Y (yield co-
efficient) taken from our previous work on feeding of the zooflagellate P_.
jaculans on bacteria (Umorin, 1975). That coefficient was equal to 33%.~
Since kg/k^ x 100% = ki, then ks = 0.006. Coefficient k6 was calculated
from the third equation of model (3). It appeared to be equal to 0.002.
Since in the experiments the concentration of nitrogen with its defici-
ency was not determined, coefficients k2 and k3 of model (4) were taken
from another experiment which had been performed to evaluate the dependence
of the specific growth rate of bacteria y upon the source concentration of
nitrate nitrogen (N). Phenol at a concentration of 25 mg/liter was used as
the source of organic carbon. The results of this experiment (average 6 re-
petitions) are provided in Table 2.
TABLE 2. DESTRUCTION OF PHENOL WHEN LIMITED BY NITROGEN
Initial con-
centration
of nitrate
nitrogen
mg/liter
0.5
1.0
1.5
Specific
growth
rate of
bacteria,
hrs"1
0.016
0.021
0.027
Initial
biomass
of bacteria
mg/liter of
dry weight
0.8
0.8
0.8
Growth of
bacteria
in dry
weight for
10 hrs,
mg/liter
0.138
0.176
0.248
Phenol
consumed
for 10
hrs,
mg/liter
0.24
0.39
0.54
Judging by the amount of bacterial growth for a period of 10 hrs,, the
nitrogen concentration in the experiments decreased by a negligible value
the k.,- coefficients have a dimension of liter-nig"1 hrs"
137
-------
(parts per hundred mg/liter) and may be considered constant during this
time period. The dependence described in the experiments y on N may be
approximated by the linear function y = 0.023 N as shown in Figure 11,
i.e., k2 = 0.023. The quantity of the consumed phenol, according to the
data given above, is approximately twice as great as the growth of bac-
teria; therefore, we accepted k3 = 2 k2. The coefficient kj was chosen so
that model (4) would most closely simulate the experimental data; it was
taken to be 0.0028. The values of coefficients k2 and k3 of model (4) do
not suit model (5). The coefficients ka, k2, k3, and k^ for model (5) were
chosen, preserving their ratios, as in model (4), in such a way that mathe-
matical model (5) would most closely approach the results of experiment 6;
in this case, coefficients k5 and k6 were taken as equal to ki, and k5 of
model (2). They were accepted to be ki = 0.0055, k2 = 0.0438, k3 = 0.0876,
k!( = 0.0025, ks = 0.0510, and k6 = 0.0240.
After determining or while choosing the coefficients, the processes de-
scribed by the models were computed on a "Minsk-22" digital computer to com-
pare the calculated and experimental curves. The calculated and experi-
mental curves correspond quite well, that is the mathematical models given
above describe rather correctly the processes taking place in the experi-
ments. An analysis of the models allows us to obtain quantitative data on
the growth and feeding rates of the living components and on the character
of their interrelationships in the process of organic matter decomposition.
The product k^H in model (2) is the rate of bacteria consumption by unit
dry weight of infusoria. Its calculation and conversion for the number of
cells show that one infusorian £. caudatum consumes about 50 thousand bac-
terial cells per hour at D = 0.04 hrs'1, and 25 thousand at D = 0.02 hr'1
at the steady state in the reactor.
Analysis of models (4) and (5) renders it possible to understand why
the destruction of organic matter by bacteria accelerates in the presence
of infusoria given a nitrogen deficiency. The necessity to make the coeffi-
cients k2 and k3 in model (5) almost twice as great as those in model (4)
indicates that in the former case nitrogen was more easily assimilable as
present in the reactor, than nitrate nitrogen inflowing with the medium.
As is known, some of the substances excreted by infusoria are urea and uric
acid (Dogel, 1951). The infusoria apparently play a role in stimulating
the decomposition of organic matter by creating nitrogen circulation and
liberating it in a form more readily accepted by the bacteria in the envir-
onment. In model (3) the values of k3S and k4H are correspondingly the
values of glucose and bacteria consumption by zooflagellates. Calculation
of them has shown that at steady state with a dilution rate of D = 0.06
hrs-1, 0.126 mg of dry weight of bacteria and 0.055 mg of glucose are con-
sumed per hour by 1 mg dry weight of zooflagellates. At a dilution rate of
D = 0.08 hrs"1 these values are equal to 0.212 and 0.047 mg, respectively.
In such a manner, in test conditions, from one-sixth to one-third of
the zooflagellates' food consumption is satisfied by consumption of organic
matter. To verify the fact that the zooflagellate P_. jaculans feeds on dis-
solved organic matter, model (3) was calculated by computer, with coeffi-
cients k3 and k6 equal to zero, so that an exclusion of glucose from the
138
-------
zooflagellate's ration was indicated. The calculation indicated that at a
dilution rate of D = 0.08 hrs'1 in this case, a washing out of the zooflag-
ellates from the reactor takes place (factually not observed), but at D =
0.06 hrs~:, the estimated calculated curve does not correspond to the
experimental one.
When converting the dry weight of bacteria consumed into the number of
cells obtained we found that one zooflagellate consumed 6.3 and 9.7 bacteria
per hour.
Assuming that the coefficients of feeding and reproduction of the organ-
isms obtained in the experiments are close to those in nature, we can esti-
mate the degree of participation of bacteria, zooflagellates and infusoria
in the processes of destruction and transformation of organic matter in any
water body, e.g., in the Rybinskoye reservoir. In this water body the con-
centration of an easily degradable organic matter is on the average 10
mg/liter (Skopintsev, Bakulina, 1966); the biomass and the production of
bacteria biomass are 0.8 mg/liter and 0.6 mg/liter day (Sorokin, 1971); the
biomass of the zooflagellates is up to 0.02 mg/liter (Zhukov, 1973); and
the biomass of infusoria is about 1 mg/liter (Mamayeva, 1971). Using these
data and the calculated coefficients, we find that bacteria consume from
250 to 750 mg of DOM per day in a cubic meter. The growth in their dry
weight must make up from 100 to 250 mg/m3 per day which quite corresponds
to the above mentioned value of the production of wet biomass.
Infusoria consume about 360 mg/m3 of the bacterial biomass per day, or
about a half their daily production. In one cubic meter, the zooflagel-
lates consume daily 4.8 mg of DOM and 1.3 mg of bacterial biomass, i.e.,
they continue to feed in nature mostly on DOM, consuming only 0.2% of the
bacterial production as food.
Thus, bacteria are the main consumer of organic matter in water bodies.
Zooflagellates cannot affect the rate of decomposition of organic matter by
predation on bacteria or by direct consumption of dissolved organic matter.
As for the infusoria, they can notably slow down the decomposition or or-
ganic matter by bacteria by means of reducing their number due to predation.
However, in water bodies, as a rule, a deficiency in nutrients has been ob-
served. Guaranteeing their recycling, infusoria may accelerate the bac-
terial destruction of organic matter.
REFERENCES
1. Bikbulatov, E.S., Skopintsev, B.A. 1974. Gidrokhimicheskiye
materialy. No. 60, 23-28.
2. Vinberg, G.G. 1949. Successes in Contemporary Biology. Uspekhi
sovrem. biol. 28, No. 2(5), 226-245.
3. Coryacheva, N.V. 1975. Transactions of the Inst. Biol. Inland Waters
Ac. Sci. USSR, 23(26). Biology, Morphology and Classification of
Water Organisms.
139
-------
4. Zhukov, B.F. 1973. "Circulation of matter and energy in lakes and re-
servoirs." Listvennichnoye na Baykale. 191-193.
5. Kryuchkova, N.M. 1968. Successes in Contemporary Biology. Uspekhi.
sovrem. biol. 65, No. 3, 466-475.
6. Lurie, Yu. Yu., Rybnikova, A.I. 1966. "Chemical analyses of indus-
trial wastewaters. "Moscow. "Khimiya".
7. Mamayeva, N.V. 1973. Inform, bull, of the Inst. Biol. Inland Waters.
Ac. Sci. USSR. No. 18, Leningrad, "Nauka", 15-19.
8. Nikolyuk, V.F. 1965. "Protista of the Uzbek soils." Tashkent,
"Nauka" UzSSR.
9. Skopintsev, B.A., Bakulina, A.6. 1966. Productivity and Circulation
of Organic Matter in Internal Water Bodies. Moscow-Leningrad
"Nauka". Transactions of the Inst. Biol. Inland Waters. Ac. Sci.
USSR. 13(16), 3-32.
10. Sorokin, Yu.I. 1971. Biology and Productivity of Fresh Water Organ-
ims. Transactions of the Inst. Biol. Inland Waters. Ac. Sci.
USSR 21(24), 5-16. Leningrad, "Nauka".
11. Umorin, P.P. 1975. Inform, bull, of the Inst. Biol. Inland Waters.
Ac. Sci. USSR, 27, 14-17.
12. Umorin, P.P. 1975. Zhurnal obshchei biologii (in print).
13. Butterfield, C.T., Purdy, W.C., Theriault, E.J. 1931. Publ. Health
Rep., 46, 393-426.
14. Canale, R.P- 1970. Biotechnd. and Bioeng. 12, 3, 353-378.
15. Javorinsky, P., Prokesova, V. 1963. Internat. Rev. Hydrobiol. 48(2),
335-350.
16. Jensen, A.L., Ball, R.C. 1970. Ecology 51, 3, 517-520.
17. Phelps, B.B. 1953. "Stream Sanitation", Wiley, New York.
18. Straskrabova,-Prokesova, V., Legner, M. 1966. Internat. Rev. Hydro-
biol. 51, 297-293.
140
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25
20
LJ
8
O
15
10
20
en
en
15 I
g
CD
<
10 S
b
123
45678
DAYS
Fig. 1. Change of component values in time — experiment 1. Left
vertical axis - concentration of organic matter mg/L, -
right - bacteria biomass in mg/L of dry weight. 1 - Bacteria
biomass, calculated curve; 2 - the same, experimental curve;
3 - concentration of organic matter, calculated curve;
4 - same, experimental curve.
141
-------
I5«
"
'I
k
ft \
?
»t
R
I
O
/ ^.
• •*• V • \>*'"*' • s- •-*
v. /
2 4 6 8 10 12 14 16 18 20 22 24 26 28 30
DAYS
Fig. 2. Change in component values in time -- experiment 2.
Vertical axis: A - Phenol concentration mg/1, B -
bacteria biomass in mg/1 of dry weight, C - infusoria
biomass - in mg/1 of dry weight. 1 - calculated
estimated curves, 2 - experimental curves.
142
-------
25
Fig. 3. Change in component values in time — experiment 3.
Symbols are the same as for figure 1.
143
-------
25
20
15
vx~''
B
0
•_ .-'• '
2 468 IO 12 14 16 18 20 22 24
DAYS
Fig. 4. Change in component values in time — experiment 4.
Symbols are the same as for figure 2.
144
-------
o25
-1^24
oS
Q. 022
Z
02I
8 10 (2 14
DAYS
16
18
1.0
o:
0.6
0.4
0.2
2O 22 24
Fig, 5. Change in component values in time — experiment 5. Symbols are the
same as for figures 1, 3.
-------
1.2
0.8
0.4
r\
•
"
" x/
--•^-
x^r^ °
" ***"*'^~^^-~^*^~^
^^ <^^ «^^
O.8
O.4
0
2 4 6 8 10 12 14 16 18 20 22 24
DAYS
Fig. 6. Change in component values in time -- experiment 6.
Symbols are the same as for figures 2, 4.
146
-------
345
DAYS
678
Fig. 7. Change in component values in time — experiment 7.
Symbols are the same as figures 1, 3, 5.
147
-------
50
-9 -
2345
H P
2.5
2.0
1.5
*"""•——6 -
1.0
-3 -
0.5
Fig. 8. Change in component values in time — experiment 8.
Vertical axis: S - glucose concentration mg/1;
H - bacteria biomass in mg/1 of dry weight; P -
zooflagellate biomass in mg/1 dry weight. 1 - bacteria
biomass, calculated estimated curve; 2 - same,
experimental curve; 3 - glucose concentration,
calculated estimated curve; 4 - same, experimental
curve; 5 - zooflagellate biomass, calculated estimated
curve; 6 - same, experimental curve.
148
-------
50
I 2345678
Fig. 9. Change in component values in time — experiment 9.
Symbols are the same as for figures 1, 3, 5, 7.
149
-------
H P
10.5
12 -
9 1
6 -
3 -
0.4
0.3
0.2
O.I
12345 23 24 25
DAYS
Fig. 10. Change in component values in time -- experiment 10.
Symbols are the same as for figure 8.
150
-------
0.03-
0.02-
Fig. 11. Dependence of the specific growth rate of bacteria
M on the nitrate nitrogen N concentration. 1 -
approximating function y = 0.023 N2 - experimental
curve.
151
-------
SECTION 14
SIMULATION OF POTENTIAL POLLUTANT-CAUSED CHANGES IN THE ECOSYSTEM,
RESULTING FROM THE SENSITIVITY OF AQUATIC ORGANISMS TO TOXICANTS
N.S. Stroganovl
The ecosystem changes in a complex manner from exposure to pollutants.
The course of the changes which take place depends both on the properties
and quantity of the pollutants and on the features of the system itself.
Aquatic organisms have a substantial effect on water properties, but their
vital activity, in turn, depends on the physical and chemical conditions of
the water body. The effect of seasonal climatic changes are superimposed
onto these interdependences. Beside the direct effect of pollutants on
aquatic organisms, secondary intermediate effects create a "web" of links
and interdependences.
It is possible to simulate a complex ecosystem using a small volume, in
which there are all of the basic components characteristic of natural aqua-
tic communities. However, such a simulation is expensive, very time-con-
suming, and requires the equipping of a large experimental base. Another
means of simulation could be based on laboratory tests and conducted ac-
cording to a definite plan using representative organisms from basic func-
tional groups which take part in the cycling of materials and of organisms
which are of interest to industry.
Aquatic ecosystems were created over a very lengthy period of time, and
have acquired a definite qualitative structure, which is disturbed on a
small scale by the seasons of the year. The introduction of pollutants, es-
pecially toxic substances, drastically disturbs the established order, and
the system passes into a new state, which is in response to the new condi-
tion. As a rule, the new state of the ecosystem is unstable. The change
which has taken place is not desirable for man, his health, or his indus-
trial activity. The instability is expressed by the disappearance or de-
crease in numbers of commercially important species of organisms and the
deterioration of water quality. Considering the fact that aquatic organisms
are a major component of the ecosystem and determine the desirable pro-
perties of the water body and the productivity of the species useful to man,
we turned our attention first and foremost to them, devoting special atten-
tion to the sensitivity of these aquatic organisms to toxic substances--
responses which can be determined under laboratory conditions.
In,
Biological Faculty of Moscow State University
152
-------
The link between functional groups is expressed diagrammatically in
Figure 1 using only the most general aspects that occur in the water body.
The following considerations are relevant to the interrelations in ques-
tion:
1. The primary organic material (algae, macrophytes) produced is al-
most entirely transformed into a "new organic material", which is partially
utilized by man in the form of commercial animals (fish, crayfish, Mollusca,
etc.).
2. Organisms which have died, as well as food residues, are destroyed
by bacteria, fungi, and protozoans into simple organic and mineral
materials, which enter again into the cycle of materials in the reservoir
as nutrients for algae, macrophytes, etc.
3. A portion of the organic material of animal and plant origin does
not mineralize and is deposited in the bottom sediment. Man, using a given
water body, is interested in high quality water and commercial organisms of
good quality. If a given system of interrelationships is in equilibrium,
then nearly all of the organic material is transformed and little forms
bottom residues. Such a situation exists in oligotrophic reservoirs. The
introduction of toxicants into such a reservoir sharply changes the rela-
tionship between functional groups because the species have both low and
high sensitivity. Some increase in number, others disappear or decrease,
and still others remain in their previous state (Figure 2).
In each functional group, there are several dozens of species in un-
equal numbers. Usually, 2-5 species are dominant in numbers and biomass
and the rest are supplementary species thst play a small or negligible role
in the interrelationships between aquatic organisms. With a change in the
environment as a result of pollution, the relationship between species
changes depending on their sensitivity. The resistant species increase and
achieve dominance (the development of blue-green algae, inferior fish,
etc.). The new relationship of species affects water quality and commer-
cial species. As a rule, this change is less desirable for man's uses.
One can, with sufficient reliability, identify the physiological basis
of the change in relationship of species by means of laboratory bioassays
of the sensitivity of the principal aquatic organisms which cycle materials
in the water body. Based on these data, the weak link in the chain of
transformations and interrelationships during exposure to toxic substances
can be identified.
Given in Table 1 are data on the sensitivity of various aquatic organ-
isms resulting from bioassays of toxicants. Shown in the table are the per-
missible, i.e., almost harmless, concentrations of the toxicants in water
based on the aquatic organism's sensitivity. The numerical data given in
the table reflect both the no-effect and permissible concentrations of the
toxicants in the water body at which vital life processes are possible. In
accordance with the accepted method (N.S. Stroganov, 1971), at these concen-
153
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Invertebrates |—*
Fig. 1. Block-diagram of the basic functional
links in the reservoir
Key:
a.
b.
c.
d.
Man
Fish
Toxicants
Bacteria
Fungi
Protozoans
e. Invertebrates
f. Algae
Macrophytes
154
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Figure 2. Change in numbers of a species with exposure to a
toxic substance. Designations: K--control; l--low concen-
tration—stimulation; 2--concentration of substance at which
95-97% of individuals dies. The individuals which survive
re-establish the population; 3—complete death.
155
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trations effects such as death or abnormalities do not increase by more
than 25% as compared with the control. We regard the stimulation of
fertility, number of the species, growth rate, and other items as desirable
effects of the toxicant. Therefore, such stimulation does not limit the
permissible concentration. From this point of view, such concentrations
are considered harmless to the test species and the system as a whole. It
follows from Table 1 that different aquatic organisms possess different
sensitivities to the same toxicant and different manifestations of vital
activity are affected in different ways. For example, the processes of re-
production and fertility are disturbed more quickly than mortality. As a
result, aquatic organisms which differ in their resistance determine the
wide variation of potential changes that occur in the ecosystem.
Variable damage to aquatic organisms from different systematic groups
eliminates some species and increases others in the system. An ecological
system is unified and all of its components are interdependent. The compli-
cated plant-animal complex not only depends on the abiotic environment, but
on the biotic factors (bacteria, protozoan, algae, and higher plant groups)
that compose the environment. The diversity of changes in the qualitative
and quantitative composition of the organisms is primary a result of the
variable sensitivity of the aquatic organisms composing the system.
It is evident from Figure 1 that at a salicylanilide concentration of 1
mg/liter, biological oxidation of organic substances will occur completely
but nitrification will occur poorly and incompletely. Consequently, basic
self-purification processes will be disturbed and will not be complete.
This in turn will create unacceptable conditions for other aquatic or-
ganisms. If the salicylanilide concentration were 0.1 mg/liter, then min-
eralization will be complete and such algae as Scenedesmus, Anabaena and
Elodea die, and Ceratophyllum will exist normally. Mollusca, worms, crust-
aceans, and fish disappear from the community.
In a natural situation, the changes which have been described will
occur in a more complex manner because the toxicant will decompose or de-
grade to another, less toxic state, and because non-resistant individuals
will be eliminated. As was evident from Figure 2 (curve 2), a portion of
the resistant individuals survive (2-3%), and re-establish the population
to normal. Such population responses to the effect of a toxicant have a
decisive effect on the structure of the community, and the change in its
structure with time. This reaction, as is evident, occurs because in-
dividuals have varying sensitivities. Apparent differences in degree of re-
action of different aquatic organisms are relative and time-dependent in
nature.
By comparing the sensitivity of aquatic organisms of varying systematic
groups, one can see that bacteria and algae are less sensitive than fish
and Daphnia. However, the magnitude of the difference also depends on the
chemical nature of the toxicant. For example, the algae Scenedesmus and
Anabaena, are more sensitive to salicylanilide than the bacteria Nitroso-
monas and Nitrobacter, while, conversely, the bacteria are more sensitive
to 8-oxyquincline. We tested several dozen substances according to the
156
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diagram indicated in Table 1. With rare exception, the bacteria which take
part in the decomposition of an organic substance are more resistant to
toxicants than invertebrate animals and fish. Algae, as a rule, occupy an
intermediate position. Consequently, these empirical data make it possible
to construct a diagram to search for the weakest link in the functional
cycles of materials in the ecosystem, and provide for a means to forecast
changes in the structure of the ecosystem and the relationship of species
in the community. Various concentrations of phenylmercuric acetate cause
some groups of organisms to drop out. At 0.0005 mg/liter the vital pro-
cesses of all aquatic organisms will occur normally (see Table 1) and the
water can be considered acceptable biologically. If mercury should accumu-
late in commercially important organisms and damage their commercial qual-
ity, then the economic norm would be disturbed, but not the biological norm.
In this case, additional analysis should be conducted to limit the concen-
tration of the substance in food organisms. At 0.01 mg/liter the processes
of self-purification will occur normally, but some algae and crustaceans
such as Daphnia will die and the growth of some fish (trout fry) will be
poor. At 0.1 mg/liter the processes of self-purification will be disturbed
but the second phase of nitrification occurs. Some algae, macrophytes,
worms and crustaceans die, fish feeding is disturbed, and fish fry die. At
1 mg/liter all the organisms indicated in the table die, and only the sapro-
phytic microorganisms which carry out biological oxidation of the organic
material remain.
The diagram of sensitivity of aquatic organisms of different systematic
groups to a toxicant provides a scientific basis for predicting potential
changes in communities exposed to toxicants at different concentrations.
By revealing the weak links in the community of aquatic organisms in the
food chain, such as through the effect of blue-green algae metabolites, it
is possible to foresee the nature of the structural rearrangements of the
community.
The means we proposed for simulation of the potential changes in the
ecosystem of a reservoir with pollution cannot be without error. Like any
model, the proposed simulation has an approximate nature, and the natural
situation may be different. But we think that the proposed diagram of simu-
lation is more complete and reliable than those proposed by other re-
searchers to judge the potential changes based on the reaction of one or
two organisms, or on biophysical bases. The simulation has two weaknesses.
1. In a natural water, several dozen, and sometimes even hundreds, of
species of aquatic organisms co-exist. We conduct tests only on representa-
tive species, i.e., predominant ones and species having commercial signifi-
cance. The rest of the species, a quantitatively larger group but less
dominant, does not determine the nature of the community. With a change in
conditions, such as the appearance of a toxicant, some dominant species can
disappear and less dominant species increase in numbers.
2. We analyzed the toxicity of only one substance, but in natural
water under present conditions, several toxicants act simultaneously.
Mutual intensification or weakening of their effect is possible. We think
157
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it is correct to sum the toxicity of toxicants. The occurrence of anta-
gonism and synergism is not encountered frequently, and can be disregarded.
It is difficult to predict the hydrochemical situation into which the toxi-
cant enter. Its effect on the aquatic organisms will depend on tempera-
ture, dissolved ^2, pH, total hardness and other indicators. Prediction is
facilitated by the inclusion of changes in the toxicity of toxicants
(metals, pesticides, etc.) caused by a series of environmental factors.
Thus, for example, a reduction in temperature usually lowers the toxicity
of the substance; a decrease in the environment's pH increases the toxi-
city of metals; and an increase in the water's hardness leads to a reduction
in toxicity. One can also list other factors which have an effect on the
final toxicity. We should not be discouraged by the difficulties in apply-
ing laboratory tests to forecast changes in the ecosytem when exposed to
toxicants. For specific water bodies perhaps it will be necessary to
change the experimental organisms. For examination of aquatic ecosystems,
considerable difficulty arises in finding major components which determine
its behavior. Simplifying the approach to solving the problem, I selected
basic functional blocks in the cycle of materials and the energy flow in the
system. They were known long ago, and have not been disputed. However, we
know little about the links within the blocks, and they often appear as
"black boxes", about which we judge based on what comes in and goes out.
For better predictions it would be desirable to know more.
The matter of the stability of the system and the mechanisms of its reg-
ulation arises in connection with predicting potential changes in the eco-
system. It seemed correct to me to single out the two major aspects in
this matter—the abiotic environment and the organisms. The stability of
the ecosystem is determine primarily by the stability of the environment.
The introduction of toxicants (pollutants) changes the aquatic organism's
living environment and, as a result, a change in the aquatic community
occurs. Specific diversity is primarily determined, apart from the histori-
cally established conditions, by hydrochemical and hydrological conditions.
However, the organisms of the functional groups can maintain the stability
of the system within known limits. This is demonstrated especially well
when some species increased their numbers or when a new species appears.
But the abiotic environment always plays the largest role in the stability
of the ecosystem. Changes in this environment often have a decisive effect
on the nature of the links in the system. The different relationships of
aquatic organisms to the environment, which contains toxicants, is the
basic relationship.
TABLE 1. COMPARISON OF HARMLESS CONCENTRATIONS OF SUBSTANCES
(MG/LITER)
Organisms Indicators Toxicants
123 4
BOD1 50 20 50 5
Bacteria NO formation 0.5 10 0.01 0.1
NO formation 0.5 15 0.01 0.01
158
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Organisms
Indicators
Toxicants
Algae:
Chlorella vulgaris
Scenedesmus quadri-
cauda
Anabaena variabilis
Macrophytes:
CeratojDhyllum
demersum
Elodea canadensis
Lemna minor
Mollusca:
Limnaea stagnalis
Planorbis planorbis
Planorbis planorbis
Planorbis planorbis
Worms:
Tubifex tubifex
Crustaceans:
Daphnia magna
Daphnia magna
Daphnia magna
Daphnia magna
Daphnia magna
Insects:
Chironomus plumosus
larvae
C. plumosus larvae
C. plumosus larvae
Fish:
Rutilus rutilus
Rutilus rutilus
Rutilus rutilus
Number of cells
Number of cells
Number of cells
Growth of stalk
Growth of stalk
Growth of roots
Survival rate of
adults
Number of eggs laid
Hatching
Survival rate of
young
Survival rate
Survival rate
Reproduction of
adults of initial
individuals
Number of genera-
tions
Number of offspring
Molting
Survival rate
Pupation
Emergence
Embryonic develop-
ment
Hatching of fry
Survival rate of
1
0.001
0.01
1
0.01
0.1
0.1
0.7
0.02
0.02
0.02
0.005
10
0.01
0.005
10
0.005
0.005
0.005
0.005
0.005
0.005
2
0.01
0.1
3
0.5
0.005
0.01
0.01
0.001
0.001
0.0005
0.01
0.005
0.01
0.01
5
5
0.1
3
1
10
0.5
1
0.005
0.001
1
0.01
0.001
1
0.01
0.01
0.01
4
0.1
0.005
0.05
0.5
0.01
0.1
0.2
0.05
0.001
0.0005
0.02
0.1
fry
159
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Organisms
Salmo gairdneri
iridens (yearlings)
S. gairdneri iridens
(yearlings)
Indicators
Feeding
Growth
Toxicants
1
0.1
0.005
2
0.1
0.1
3
0.005
0.005
4
0.01
0.002
Toxicants: l~salicylanilide;
3--8-oxyquinoline;
2—sodium pentachlorophenol ate;
4--phenylmercuric acetate
Temperature 18-22. Length of test 30 days (10-20 days for the bacteria)
Survival rate is indicated for 50% of individuals.
^Biochemical Oxygen Demand
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SECTION 15
FISH-POPULATION STUDIES IN THE OHIO RIVER
William C. Klein
INTRODUCTION
In 1957 the Ohio River Valley Water Sanitation Commission (ORSANCO) and
the Kentucky Division of Fish and Wildlife initiated a 3-year study -
Aquatic Life Resources in the Ohio River (ALRP). Chemical fishing in navi-
gation locks of the U.S. Corps of Engineers was used as one of the princi-
pal methods for collecting the samples to be analyzed. Subsequently, from
1968 to 1970, state, federal, and inter-state agencies continued the
investigations by cooperative arrangement and gathered data to establish a
relationship between trends in the fish population and changes in water
quality occurring from the installation of improved wastewater treatment
facilities. The goal of these ongoing studies was to provide agencies re-
sponsible for water quality management in the Ohio River region with infor-
mation necessary for assessing river quality conditions.
In 1974 ORSANCO began to expand its monitoring program on the Ohio
River and the lower reaches of its major tributaries to contribute needed
information to these agencies. An important part of the program, biologi-
cal monitoring at selected locations, was partially initiated during the
fall of 1975; again chemical fishing was used. Some background together
with a summary of the methods used and the results obtained from the pre-
vious studies, is detailed below.
THE OHIO RIVER
The Ohio River is a large canalized stream extending 981 miles from
Pittsburg, Pennsylvania, where it is formed by the Allegheny and Monogahela
Rivers, to Cairo, Illinois, where it flows into the Mississippi River. At
normal pool stages the stream varies in width from approximately 1,000 ft
to 4,000 ft in the lower reaches. Flow patterns in the river are extremely
variable, ranging from 6,600 cu ft/sec1 in the upper river to 48,5000 cu
ft/sec1 in the lower reaches. Presently, its depth is controlled by a
series of high- and low-level dams and associated navigation locks at some
21 locations. The U.S. Corps of Engineers maintains a minimum 9-ft channel
Minimum 7-day in 10-year flow,
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for navigation purposes. The Ohio River receives the flow from 19 major
tributaries and over 100 reservoirs, in addition to discharges from about
295 municipalities and 200 industries. Although the flow in the Ohio River
is quite variable, it is highly regulated by the U.S. Corps of Engineers
through their manipulation of the reservoir and dam system (see map).
The change in river regime caused by its canalization and increased
pollution load has altered the species composition and relative abundance
of biological organisms from those found and recorded by the early white
settlers, and changes in dam construction have had an impact upon migratory
patterns of fish. The old wicket dams, for instance, permitted open-river
conditions for many months of the year, and fish were free to travel from
one pool to another. The installation of the fixed structure high-level
dams across the river does allow fish to travel from one pool to another,
although this does not occur in the same free manner as before. In addi-
tion, the change in variety of food organisms has been substantial and was
probably more influential ecologically than any changes in water quality.
These alterations must be taken into account when evaluating the results of
the lock-chamber studies. For example, the shift from a free-flowing
stream with a significant slope to a canalized river separated by a series
of low-level locks and dams and then to the present condition of longer,
deeper pools separated by fixed structured dams has significantly modified
biological habitats to the extent that many previously abundant species
such as the sturgeon and the paddle-fish are now reduced in number and
their distribution limited. Some species of fish (such as the deep-bodied
suckers, the gizzard shad, and perhaps some of the smaller sunfish) have
increased in abundance, for the lakes created by the dams favor them.
Carp, another species introduced into the Ohio, shows strong preference for
the quiet waters furnished by the dams.
From the standpoint of spawning and reproduction of fish, the Ohio ex-
hibits many of the traits of a large canalized river. The stream is
characterized by a gravelly bottom, a paucity of shallow water and very
few, if any, riffles or weeds suitable for nesting. Shore lines in many
localities show the effects of bank erosion caused by the large variations
in river flow and to a lesser degree the backwash of commercial and recrea-
tional boats navigating the river. As a result, the number of areas suit-
able for spawning takes place in small creeks and tributaries. Sampling
performed during the course of the ALRP, however, revealed that a large
number of species requiring shallow water, weeds, and riffles to reproduce
were in fact spawning in the small tributaries and then returning to the
main stem. The sauger, round-bodied suckers (red horses), largemouth and
smallmouth bass, and golden shiners prefer the small tributaries with
shallow gravel bars and weeds for spawning, but they are found throughout
the Ohio as both fing'erlings and mature fish.
CHEMICAL FISHING
Various chemicals (rotenone, toxaphene, creosol, copper sulfate, and
sodium cyanide) have been used in fish sampling. The most acceptable of
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these has been rotenone because of its high degradability, freedom from
such problems as precipitation and persistent toxicity, and above all, the
relative safety for the user. Rotenone, which is obtained from the derris
root (Deguelia elliptica, East Indies) and the cube root (Lonchocarpus
nicour, South America) has been used extensively since 1934 in fishery work
throughout the United States and Canada. The chemical is toxic to man and
warm-blooded animals (132 mg/kg), but has not been considered hazardous in
the concentrations used for fish eradication (0.025 - 0.050 mg/liter active
ingredient). Therefore, it has been employed in waters used for bathing
and in some instances in drinking water supplies. Activated carbon removes
rotenone very effectively, as well as the solvents, odors, and emulsifiers
present in almost all commercial rotenone formulations. The rotenone used
in lock-chamber studies is a 5% active ingredient in an emulsion base.
Best results are obtained with water temperatures above 13 C (55 F). It is
a relatively fast-acting toxicant which decomposes in 24 hr or less. The
toxicity threshold, however, differs only slightly among fish species, and
for this reason rotenone cannot be used as a selective toxicant for certain
species.
LOCK-CHAMBER APPLICATION
The lower gate of the lock chamber is left open approximately 4-6 hr
before the sampling. Personnel move into the lock chamber in boats, and
the lower gates of the chamber are closed. The rotenone emulsion is then
pumped and sprayed from a boat into the water within the lock chamber until
a concentration of 1 mg/liter is attained. The chemical is then rapidly
dispersed through the water inside the chamber by means of the ouboard
motors on the boats. The fish begin rising to the surface 5-10 min after
the rotenone has become well mixed. As the fish surface, personnel in
boats pick them up with dip nets and place them in large tubs. Because of
the size of the chambers, i.e., 100 ft by 1,000 ft, approximately five
boats and 10 men are required to conduct the work. Additional personnel
spot the fish as they surface. After all the fish have been picked up and
placed in receptacles, the lock-chamber gate is partially opened, and the
water is permitted to bleed out slowly. The fish are then taken to a
central location near the lock and are sorted, weighed, identified, and
catalogued. Appropriate species, such as the channel catfish, are frozen
in dry ice, shipped to laboratories, and analyzed for heavy metals and
pesticides. Recently, the U.S. Food and Drug Administration has cooperated
in analyzing the fish for these constitutents. The fish are also inspected
for parasites and other pathological indications possibly attributable to
adverse water quality conditions. Fish not used in the additional studies
are disposed of by burying.
SAMPLING RESULTS
A comparison of lock-chamber sampling results for the 1957-60 and
1968-70 periods reveals that a number of changes have taken place in the
composition of the fish population in the Ohio River, reflecting altera-
163
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tions in physical, chemical, and hydrological conditions such as the re-
placement of old style wicket dams, increased water pollution control, and
augmented stream flow from reservoirs during the low-flow period. The
changes appear to parallel and substantiate physical and chemical water
quality trends noted during the same period.
In the 1957-60 period the 10 most abundant species of fish in the popu-
lation samples were in descending order: the emerald shiner, gizzard shad,
freshwater drum, mimic shiner, channel catfish, silver chub, black bullhead,
threadfin shad, blue catfish, and sand shiner. The 10 species that contri-
buted the greatest total weight in the samples were gizzard shad, carp,
channel catfish, freshwater drum, emerald shiner, skipjack herring,
flathead catfish, blue catfish, black bullhead, and river carpsucker.
In comparison, the 1968-70 sampling revealed that six of the 10 most
abundant species in the population—the gizzard shad, emerald shiner, fresh-
water drum, channel catfish, bullhead, and the drum—were retained from the
1957-60 sampling period. The remaining species were replaced by the carp,
black crappie, yellow bullhead, and river carpsucker. Of the 10 species
that contributed the greatest total weight in 1957-60 the gizzard shad,
carp, channel catfish, freshwater drum, river carpsucker, flathead catfish,
and black bullhead continued in that category. The three new species were
bigmouth buffalo, white crappie, and bluegill.
The species composition also varied throughout the river as it did in
1957-60. In the upper third the most abundant species were carp, channel
catfish, gizzard shad, emerald shiner, and brown bullhead; in the middle
third, the carp, gizzard shad, channel catfish, mimic shiner, and skipjack
herring; and in the lower third, the gizzard shad, channel catfish, emerald
shiner, freshwater drum, and bigmouth buffalo.
Among the changes noted during the 10-year period was the marked in-
crease in the carp population, which had not been so predominant in the
earlier studies, and the increased abundance of species sought by both
sport and commercial fishermen - largemouth and smallmouth bass, sauger,
crappies, and sunfish. It is believed that the new deeper pools with de-
creased velocities and more lake-like settings probably account in large
measure for the increased carp population. The re-emergence of significant
numbers of the so-called sport and commercial species is probably due to
the decreased water pollution loads going to the river. Such a conclusion
is supported by ORSANCO appraisals of water quality conditions. Of the 21
water quality characteristics routinely monitored by ORSANCO, all except
four are now meeting established criterial goals for streams in the compact
area.
SUMMARY
A review of historical and recent information concerning fish in the
Ohio River during 1957-60 and 1968-70 indicates that the composition of the
fish population has changed during the period. In large measure, the
164
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changes can be attributed to the canalization of the river and increased
pollution load. Although the pollution load has been decreased in recent
years by the installation and operation of wastewater control facilities,
the lake-like setting of the Ohio River continues to influence the kinds
and numbers of fish in the river, as evidenced by the chemical fishing
studies performed in the lock chambers. Although many of the so-called
sport and commercial species have returned to the river, the fish species
desiring a lake-like setting continue to dominate the population.
REFERENCES
Clay, W.M. 1962. A field manual of Kentucky fishes. Kentucky Dep. Wildl.
Res., Frankfort, Ky.
Krumholz, L.A. 1950. Some practical considerations in the use of rotenone
in fisheries research. J. Wildl. Manag.
Ohio River Valley Water Sanitation Commission. 1962. Aquatic life re-
sources of the Ohio River. Cincinnati, Ohio.
Post, G. 1958. Time vs. water temperature in rotenone dissipation. Proc.
38th Annu. Conf. Game Fish Comm.
Trautman, M.B. 1957. The fishes of Ohio. Ohio State Univ. Press,
Columbus, Ohio.
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SECTION 16
REGISTRATION OF PESTICIDES:
CONSIDERATIONS IN CONDUCTING AQUATIC TOXICITY TESTS
Richard A. Schoettger
Reports of the U.S. Tariff Commission show that production of synthetic
organic chemicals amounts to billions of pounds per year. Pesticides alone
account for more than 1 billion pounds, of which about half are insecti-
cides and the remainder are herbicides, fungicides, and other control chemi-
cals (Fowler and Mahan, 1973). Extensive use of persistent pesticides for
over a quarter-century, often without due concern for direct and indirect
contamination of fish and wildlife habitats, has resulted in our unwitting
use of these resources as biologic indicators of contamination (Johnson,
1968; Henderson, Johnson, and Inglis, 1969; Katz et al., 1970, 1971, 1972;
Day, 1973; McKim et al., 1973, 1974). These fish and wildlife resources
are valued at more than $7 billion per year (U.S. Fish and Wildlife
Service, 1972). The lethal effects of pesticide spills, careless applica-
tions, and point and non-point discharges have been relatively obvious for
years, but not the subliminal effect of sublethal concentrations. Recent
improvements in analytical technology and nationwide sampling by national
monitoring activities have revealed a broad array of pesticide and
industrial chemical residues in various kinds of fish and wildlife and
their habitat. These findings show that trace quantities of pesticides can
be mobile and accumulative in aquatic ecosystems. With new, multidiscipli-
nary research approaches, scientists are now beginning to demonstrate what
they suspected for years—that sublethal concentrations of pesticides and
other contaminants may have subtle and adverse effects on basic life and be-
havioral processes of fish and wildlife. The scope of these processes
determines an organism's ability to cope with continuous competition and
natural stresses. Chemical contaminants are added stresses to which fish
may or may not be able to adjust, and populations may be subtly modified or
attenuated. Therefore the U.S. Fish and Wildlife Service must anticipate
to the best of its ability, through its own research and in cooperation
with other agencies and institutions, the ecological implications of known,
suspected, or potential chemical contaminants.
In view of documented effects of pesticides on fish and other aquatic
life and the apparently ubiquitous distribution of certain pesticide resi-
dues in aquatic habitats, it seems reasonable to assume that past research
requirements for pesticides have not been adequate to anticipate effects on
these resources. In 1970-71 the U.S. Fish and Wildlife Service reorganized
and integrated scientific disciplines at the Fish-Pesticide Research Labora-
166
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tory (Figure 1) so as to develop anticipatory rather than documentary infor-
mation concerning effects of pesticides on fish and fish-food organisms
(Grant and Schoettger, 1972).
High priority at the laboratory is given to research in four topical
areas:
1) Agents developed for use in fishery habitats (control of
aquatic weeds, algae, slime, mosquitos, mollusks, and fish);
2) agents intended for use on land and water adjacent to or
contiguous with fishery habitats (forest insect control,
ditch bank management, forest fire retardants);
3) agents manufactured in large volume and used widely
(selected agricultural and industrial chemicals); and
4) known contaminants of wild and propagated fish and their
food and habitat.
The investigational divisions outlined in Figure 1 are intended to re-
flect the kinds of studies that should be considered in the development and
registration of pesticides. The principal systems are ordered from rela-
tively routine and short-term studies to more complex and lengthy investi-
gations. All or parts of the framework may be used depending on the extent
and applicability of biological and chemical data already available,
intended use pattern(s), and target pest(s).
The framework is designed around fish as the primary test animal, but
it is also compatible with parallel investigations essential to antici-
pating pesticide effects on fish-food organisms. In all studies the
investigator must include sufficient test animals and replications to
estimate statistical significance of results. Sources, general physical
conditions, disease treatments, and holding conditions (such as photo-
period, diet and feeding rate, water characteristics) of test animals
should be reported. Whenever possible, test animals, diets, and holding
waters should be chemically analyzed to document pre-exposure of test ani-
mals to pesticides or other contaminants. Analytical chemistry reports
should document results for reagent blanks, limits of sensitivity and de-
tection, reproducibility, recovery efficiency for extracts, and sample
variability.
The investigational sections within the research framework are divided
into principal systems and support systems; consequently, researchers in
two or more research divisions generally integrate their efforts to achieve
common goals. Typical investigations generated by this framework include
some 11 types of studies:
(1) Acute toxicity, and variations among species and water types;
(2) teratogenicity;
167
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RESEARCH SYSTEMS
A. PRINCIPAL SYSTEMS
B.SUPPORT SYSTEMS
ACUTE TOXICITY
[Static tests}
[Flow-through tests
"
CHRONIC TOXICITY
[pro w th_ & reprod uction
llfVhpviprl
[Adaptabi|1ty_ _*****3
III
ECOSYSTEM STUDIES
ilenticl
lloticj
ANALYTICAL CHEM.
Identification
| Methods'!
MUPTAKE& EXCRETION
BIOPASSAGE
CLINICAL
Physiology |
[Pathology J
[Biochemistry!
IV
CHEMICAL MODIFICATION
[Biological!
«PhyjKal_ _Che_miaBJ__{
Figure 1. Organization of investigational divisions at the Fish-Pesticide
Research Laboratory, Columbia, Mo.
168
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(3) biological uptake, storage, and elimination;
(4) sublethal effects on growth, reproduction, and
morphogenesis;
(5) physiologic-biochemical effects and the organisms'
homeostatic ability to adapt to natural stresses;
(6) effects on energy transfer and physical, chemical, and
biological interrelations in aquatic ecosystems;
(7) methods for identifying and quantifying pesticides,
their degradation products, and other contaminants;
(8) physiochemical factors affecting molecular structure
and biologic activity;
(9) effects on behavior;
(10) methods for eliminating or deactivating chemical resi-
dues; and
(11) correlations of residues with their biologic activity.
Mount (1967) pointed out that numerous past studies on toxicological
and physiological effects of pesticides in fish have yielded few data that
can be used to correlate these effects and chemical-residue measurements
with significant damage to aquatic forms. Therefore, investigators must
keep in mind the potential interpretive value of anticipatory research on
pesticides. In-depth experiments should be designed so that they
demonstrate effects of pesticides on aquatic organisms, but they should
also include measurements of residues induced by test concentrations which
are commensurate with concentrations recommended for use of the pesticide.
Such data are essential to experimental designs for field evaluations of
pesticides and for interpreting significance of unintentional contamination
of aquatic ecosystems.
The research framework discussed above, along with a more detailed
account of guidelines for conducting toxicological research with aquatic
organisms, was published by the National Academy of Sciences in 1973.
REGISTRATION REQUIREMENTS
For the most part, pesticides must be registered in the United States
according to provisions of the Federal Insecticide, Fungicide, and Rodenti-
cide Act (FIFRA). A number of provisions in this act were most recently
amended by the Federal Environmental Pesticide Control Act of 1972. Re-
sponsibility for implementing FIFRA, as amended, is vested in the Office of
Pesticide Programs of the U.S. Environmental Protection Agency (EPA). In
general, properties of pesticides that must be researched in the registra-
169
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tion process include (1) efficacy on the target pest, (2) general and en-
vironmental chemistry of the pesticide, (3) safety to the applicator and
the consumer of treated products, and (4) effects on non-target organisms,
including those of aquatic ecosystems. More specifically with regard to
aquatic organisms, the administrator (EPA) "shall register a pesticide pro-
duct or approve amended and supplemental registration if he determines that,
when considered with any restrictions imposed, the pesticide will perform
its intended function without unreasonable adverse effects on the environ-
ment" (Environmental Protection Agency, 1975a).
If a pesticide is intended for outdoor use, or for use where it may
contaminate water, applicants for registration must submit data that will
permit evaluation of hazards to non-target animals, as given in the pro-
posed registration guidelines (Environmental Protection Agency, 1975a).
The minimum requirements include the acute toxicity (96-hr LC50) of the
technical grade pesticide for both a coldwater and warmwater species of
fish, such as rainbow trout (Salmo gairdneri) and bluegills (Lepomis
macrochirus). An acute test must also be performed on an aquatic inverte-
brate, such as a daphnid. Data reports must include calculations of the
dose-response line, the 95% confidence intervals for the LC50, and the
slope of the regression. Other studies may be required, depending on
whether acceptable research demonstrates that, under conditions of pro-
posed use, the pesticide causes no unreasonable adverse effects on plants
and animals of aquatic ecosystems. Such additional studies, considered
"conditional tests," may be judged necessary depending on other factors,
such as (1) chemical and physical properties of the pesticide, (2) amount
of pesticide applied per unit of area or time, (3) likely degree of
contamination in various environmental components according to proposed
use, (4) various species to be affected, (5) likely routes of exposure, (6)
persistence of the pesticide or its biologically active degradation pro-
ducts and transfer between environmental components, and (7) degree of
biological uptake of the pesticide or its significant degradation products.
The "conditional tests" that concern water contamination may include
(1) additional acute toxicity tests with the technical grade or formulated
pesticide against freshwater and estuarine or marine fish and invertebrates;
(2) toxicity-residue studies with bottom-feeding fish (channel catfish,
Ictalurus punctatus, or carp, Cyprinus carpio), predaceous fish (e.g.,
largemouth bass, Micropterus salmoides, bluegills, or trout), molluscs
(oysters or freshwater clams), Crustacea (Daphnia sp., Gammarus sp., or
crayfish), and insect larvae; (3) studies of chronic effects on reproduc-
tion of fish or invertebrates or both; and (4) "special studies" (actual or
simulated field studies in which proposed use patterns are tested). Other
toxicity data may be required where unusual or specific potential hazards
may be associated with a particular proposed pesticide use. At present,
when a pesticide is proposed for aquatic use, the applicant for registra-
tion is required to provide data to establish a residue tolerance in edible
tissues of fish, shellfish, or both, or obtain an exemption from the re-
quirement for a tolerance. Research guidelines for establishing tolerances
have not yet been proposed. In general, however, such research would
probably include oncogenic evaluations; chronic feeding studies in animals
170
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for the measurement of effects on the central nervous system and the hema-
topoietic system; and histo"logical changes in the liver, kidney, and male
and female reproductive systems.
Registration or re-registration of a pesticide may be questioned if the
proposed use could result in an average concentration, in water 6 inches
deep, greater than 0.5 of the LC50 for representative aquatic organisms.
Pesticides giving an average concentration in 6-inch-deep water of between
0.1 and 0.5 of the LC50 would most likely be classified for restricted use
by certified applicators. These criteria also apply to metabolites or de-
gradation products of the pesticide. In addition, a determination of un-
reasonable adverse effects on the environment must include an analysis of
the chronic effects of exposures to pesticides.
REGISTRATION GUIDELINES
The various test methods by which pesticide registration requirements
can be satisfied are included as an appendix to "Guidelines for Registering
Pesticides in United States" (Environmental Protection Agency, 1975b). In
general, methods are organized along the lines of the pesticide registra-
tion requirements - i.e., methods are given for the assessment of (1) pesti-
cide efficacy, (2) environmental chemistry, and (3) hazards to humans,
domestic animals, fish, and wildlife. The presentation of methods is not
intended to imply that they are necessarily standard, inflexible, or the
only methods that can be used. However, they are now considered acceptable
for developing data to support registration and for planning research and
are an excellent source of information. Literature citations are given
only for references that are readily available and that describe methods
that are acceptable as presented. Modifications of methods are presented
as annotated bibliographic citations, and unpublished methods are included
as full tests. Applicants for pesticide registrations are encouraged to
discuss with EPA the research methods they intend to use.
The aquatic toxicology methods include acute toxicity testing with
various marine and freshwater fish and invertebrates, chronic (complete
life cycle) and partial chronic studies (includes reproductive phase of
life cycle), accumulation tests, and field appraisal studies. The methods
classed as "routine" have been used by numerous investigators for many
years to investigate a wide variety of toxicants. Those methods classed as
"tentative" have been used by two or more toxicologists for several years,
but there is no consensus concerning detailed application of the methods,
and there are no inter!aboratory test comparisons to show consistency of
results. "Developmental" methods are those used or proposed by one or a
few investigators, and the techniques involved may not be well known and
may require that investigators have considerable experience to achieve
consistent results.
171
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CHRONIC TESTS
Except for chronic toxicity tests with fathead minnows (Pimephales
promelas), chronic or partial chronic tests with various aquatic forms--
sheepshead minnows (Cyprinodon variegatus), brook trout (Salvelinus
fontinalis), bluegills, daphnids, and midges--are considered tentative or
developmental. Because such studies may be required to appraise the hazard
of a pesticide to non-target animals, I would like to summarize the results
of two studies that we (Staff, Fish-Pesticide Research Laboratory) recently
conducted one with toxaphene (Mayer, Mehrle, and Dwyer, in press; Mehrle
and Mayer, in press) and one with 3-trifluoromethyl-4-nitrophenol (TFM)
(Foster Mayer, personal communication).
Toxaphene Investigations
Growth and reproductive effects—Although use organochlorine insecti-
cides has been reduced in recent years, some including toxaphene are still
used extensively. Between 30 and 40 million pounds of toxaphene are
currently being applied annually on crops and livestock in the United
States. Since use of DDT was curtailed, toxaphene has often been used to
replace it, alone or in combination with other insecticides. We therefore
began cooperative studies in 1972 with EPA to evaluate the effects of toxa-
phene on fishery resources.
Toxaphene is acutely toxic to fish; lethal threshold concentrations for
brook trout, bluegills, fathead minnows, and channel catfish range from 0.5
to 15.2 yg/liter. In earlier studies we found that growth of adult brook
trout was reduced during continuous exposure to 0.29 and 0.50 yg/liter toxa-
phene, and the added stress of natural spawning caused extensive deaths of
adults at these concentrations. Growth and survival of fry were affected
adversely at concentrations as low as 0.039 yg/liter, and they accumulated
toxaphene residues 5,000 - 21,000 times the water concentration.
Because toxaphene is used extensively on cotton in the southeastern
United States, we also tested it against the fathead minnow, an important
forage and bait species, and against channel catfish. Ten-day-old fathead
minnow fry were exposed continuously to concentrations of 0.06 - 1.2
yg/liter of toxaphene. The fish were reared at a constant temperature of
24 C and under a regulated photoperiod approaching natural lighting.
Growth of the fish was not affected in exposures as long as 90 days.
Between 90 and 150 days, however, the growth of all fish exposed to toxa-
phene was significantly less (P<0.05) than that of the control fish. At
this time toxaphene residues accumulated during this period exceeded 90,000
times those in the treated water. Residues in fish exposed to the highest
concentration, 1.2 yg/liter, averaged 94 yg/g.
Two-year-old channel catfish were also exposed continuously to concen-
trations of 0.023 - 0.51 yg/liter of toxaphene for 4.5 months before
spawning (Figure 2). Spawning occurred naturally through manipulation of
photoperiod and water temperature, and 85% of the fish that reached sexual
maturity spawned. Although the adults were not affected, hatchability of
172
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co
Figure 2. Multiple concentration, flow-through dlluter with controlled light and temperature
used to determine sublethal effects of toxaphene on growth and reproduction of
channel catfish.
-------
eggs from adults exposed to 0.51 yg/liter was reduced slightly, and 0.22
and 0.51 yg/liter of toxaphene increased the mortality of fry and decreased
their growth. Toxaphene residues in fry from these two exposures were 8
and 32 yg/g, respectively.
Toxaphene may also have an adverse effect on important natural fish
foods. Concentrations of 10 yg/liter or greater inhibited emergence of
midge larvae, but this is well above the concentrations affecting repro-
duction or growth in fish. However, greater resistance of midge larvae
could enable them to accumulate significant residues. Reproduction of
daphnids was halved when the organisms were exposed to 0.12 yg/liter toxa-
phene for 21 days, and the no-effect concentration was only 0.007 yg/liter.
Toxaphene concentrations of 0.04 - 0.25 yg/liter are detrimental to the
production of fish and their food; consequently contamination of waters
supporting these resources by run-off, leaching, or spraying should be
avoided. Unfortunately, these low concentrations are very difficult to
detect analytically. However, tissue residues exceeding 0.4 - 1.8 yg/g in
salmonids may be associated with reduced growth and reproductive success,
and residues over 5 yg/g may cause reduction in growth of channel catfish
fry.
Biochemical effects—Collagen is the major fibrous protein of all verte-
brates and serves as the major component in the organic matrix of connec-
tive tissue and bone. The proper ratio of collagen and minerals is
essential for rigidity and flexibility of bone, as well as overall develop-
ment and maturation.
Eggs and fry of brook trout and young fathead minnows were exposed to
toxaphene for 90 and 1500 days, respectively, at the same concentrations
(0.039 - 0.5 yg/liter and 0.06 - 1.2 yg/liter) as those tested in the
growth and reproduction studies reported above. Analyses of backbones
showed that synthesis of hydroxyproline, the major amino acid of collagen,
was inhibited during the first few weeks of exposure to toxaphene at con-
centrations of 0.039 yg/liter or higher. In older fish collagen synthesis
was reduced at all concentrations of toxaphene by the end of the exposure
period and appeared to be correlated with reduced growth. The earliest in-
hibition of collagen synthesis occurred at the highest toxaphene concentra-
tions and preceded observable reductions in growth.
In general, the net effect of toxaphene in fish was lower collagen
synthesis and greater mineralization of the backbone and whole body. We
postulated that this condition may cause the backbone of fish to be brittle
and fragile and therefore subject to breakage during times of swimming
stress. We subjected groups of toxaphene-treated and control fathead
minnows to a sublethal electrical shock (60, AC) and then examined the
backbones by x-ray. The observations (Figure 3) confirm that the backbones
of toxaphene-treated fish seem more fragile and could break while the fish
are migrating or escaping predators.
174
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Figure 3. Effects of toxaphene on backbone structure of fathead
minnows. Radiographs A and B represent fish held in
water with a low toxaphene concentration (0.055 yg/liter);
C represents the control group. Arrows show areas of
backbone affected.
175
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A condition known as "broken-back syndrome" has been reported by other
investigators in pond-reared channel catfish, as well as in natural popula-
tions. Studies with fingerling channel catfish showed that exposures for
90 days to concentrations ranging between 0.044 and 0.535 yg/liter signifi-
cantly decreased collagen and increased calcium in their backbones. Resi-
dues in the affected finger lings ranged upward from 3 yg/g. X-ray analyses
of these fish revealed aberrations in backbone structure (Figure 4).
Studies to determine the mechanism of action as well as the possibility
that other contaminants could also induce this condition are now in
progress.
TFM Investigations
The lampricide 3-trifluoromethyl-4-nitrophenol (TFM) was registered in
1964 for control of larval sea lampreys (Petromyzon marinus) in selected
tributaries of the Great Lakes. The EPA is presently renewing registration
of TFM on a year-to-year basis while research is being conducted on poten-
tial adverse effects and residues in non-target species. Brook trout are
an important and indigenous sport fish in many of the streams treated with
TFM.
Chronic exposures of adult brook trout to concentrations ranging from
0.7 to 14 mg/liter of TFM formulation (35.7% active ingredient) were begun
in 1973. The adults were exposed for 120 days before spawning, and their
offspring for 90 days. Growth of adults exposed to the highest concentra-
tion was reduced, and all died during spawning. Many of the adults exposed
to 14 mg/liter and a few of those exposed to 8 mg/liter developed blindness.
Concentrations of 3.3 mg/liter or higher reduced egg viability (as measured
by the percentage reaching the neutal keel stage) and hatchability and
growth rates of fry.
Although TFM has a significant chronic effect on brook trout at concen-
trations well below those used to control lamprey larvae, it is not likely
that use patterns for TFM would result in such long and continuous expo-
sures. Therefore, we repeated the study, but exposed two groups of adult
trout in a light- and temperature-controlled flow-through diluter, in
simulation of a typical stream treatment. Because such treatments
generally take place during the summer or early fall, one group of fish was
exposed to TFM during the summer at 15 C and the second during the fall at
9 C. Both groups were exposed to 16-18 mg/liter of TFM for 12 hr. About
19% of the adults in the first group died shortly after exposure, probably
because TFM was more toxic at the higher temperature, but those in the
second group were not affected. None of the treated adults showed signs of
blindness, and all spawned normally in November. Viability and hatch-
ability of the eggs were similar in the treated and control fish (Figure 5),
and growth of the young was not affected.
The results of these investigations serve to illustrate the utility and
versatility of chronic and partial chronic toxicity tests in estimating
potential impact or non-impact of pesticides on aquatic organisms. At pre-
sent, there are about 30,000 registered pesticide formulations that must be
176
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Figure 4. Effects of toxaphene on backbone structure of channel
catfish. Radiograph A represents a fish exposed to
0.055 yg/liter, B a fish exposed to 0.044 yg/liter,
and C a control.
177
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100 r
Viability
Hatch
o
CD
Q_
00
TFM Concentration in Water
(mg/l)
Figure 5. Comparison of viability and hatch in eggs from brook trout exposed continuously or
by simulated usage pattern to TFM.
-------
periodically re-registered. Because a significant number of new and
registered pesticides may contaminate aquatic ecosystems, chronic or
partial chronic tests provide an important intermediate measure of relative
hazard between simple acute toxicity tests and costly experimental field
trials. Within practical limits such studies can be adapted to include
simulation of various general aquatic habitats, and timing and concentra-
tion of exposures can be controlled to approximate proposed or recommended
uses. In addition, chronic test systems offer a unique opportunity to con-
duct concurrent or parallel studies of residue dynamics and physiological,
biochemical, and pathological effects that can be linked to growth and re-
productive effects in primary chronic tests.
SUMMARY
Requirements for registration, re-registration, and classification of
pesticides for general or restricted use were published recently by the EPA.
If the pesticide is intended for outdoor uses, data must generally be sub-
mitted that permit evaluation of hazards to non-target animals, including
fish and wildlife. Depth of these evaluations depends on proposed patterns
of use, environmental chemistry characteristics, and nature of the hazard
to humans, domestic animals, and non-target animals. Data to support regis-
tration can be obtained from acute and chronic or partial chronic toxicity
studies, simulated field tests, or field monitoring and observation, as des-
cribed in the extensive registration guidelines recently proposed by EPA.
Chronic testing techniques and apparatus, with controllable light and
temperature, offer versatile systems for investigating effects of pesti-
cides and other contaminants on fish according to daily and seasonal
periodicity and simulated pesticide-use patterns.
REFERENCES
Day, K. 1973. Toxicology of pesticides: Recent advances. Environ. Res.
6: 202-243.
Fowler, D.L., and J.N. Mahan. 1973. The pesticide review. Agricultural
Stabilization and Conservation Service, U.S. Dep. Agric., Washington,
D.C. 60 p.
Grant, B.F., and R.A. Schoettger. 1972. The impact of organochlorine
contaminants on physiologic functions in fish. Proc. Tech. Sess. 18th
Ann. Meeting Inst. Environ. Sci., New York City. p. 245-250.
Henderson, C., W.L. Johnson, and A. Inglis. 1969. Organochlorine
insecticide residues in fish (National Pesticide Monitoring Program).
Pest. Monit. J. 3: 145-171.
Johnson, D.W. 1968. Pesticides and fishes--a review of selected litera-
ture. Trans. Am. Fish. Soc. 97: 398-424.
179
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Katz,.M., R.S. Legore, D. Weitkamp, J.M. Cummins, D. Anderson, and D.R.
May. 1972. Water pollution: Effects on freshwater fish. J. Water
Pollut. Control Fed. 44: 1226-1250.
Katz, M., O.E. Sjolseth, D.R. Anderson, and L.R. Tyner. 1970. Water
pollution: Effects of pollution on fish life. J. Water Pollut. Control
Fed. 42: 982-1025.
Katz, M., C.H. Wahtola, R.S. Legore, D. Anderson, and S. McConnell. 1971.
Water pollution: Effects on freshwater fish. J. Water Pollut. Control
Fed. 43: 1334-1363.
Mayer, F.L., P.M. Mehrle, and W.P. Dwyer. Toxaphene effects on reproduc-
tion, growth, and mortality of brook trout. U.S. Environmental Pro-
tection Agency, Duluth, Minn. Ecol. Res. Ser. EPA-600/3-75-013.
p.
McKim, J.M., G.M. Christensen, J.H. Tucker, D.A. Benoit, and M.J. Lewis.
1974. Water pollution: Effects of pollution on freshwater fish. J.
Water Pollut. Control Fed. 46: 1540-1591.
McKim, O.M., G.M. Christensen, J.H. Tucker, and M.J. Lewis. 1973. Water
pollution: Effects of pollution on freshwater fish. J. Water Pollut.
Control Fed. 45: 1370-1406.
Mehrle, P.M., and F.L. Mayer. In press. Bone development and growth of
fish as affected by toxaphene, p. . _In_ I.H. Suffet (ed.) Fate of
pollutants in air and water environments. Wiley Intersci. Publ., New
York.
Mount, D.I. 1967. Considerations for acceptable concentrations of pesti-
cides for fish production, p. 3-6. JJT^ E.L. Cooper (ed.) A symposium
on water quality criteria to protect aquatic life. Am. Fish. Soc.
Spec. Publ. 4.
National Academy of Sciences. 1973. Water quality criteria, 1972. Appen-
dix HE. Washington, D.C. p. .
U.S. Environmental Protection Agency. 1975a. Registration, re-registration
and classification procedures. Federal Register, 40(129): 28242-28286.
U.S. Environmental Protection Agency. 1975b. Guidelines for registering
pesticides in United States. Federal Register, 40(123): 26802-26928.
U.S. Fish and Wildlife Service. 1972. National survey of fishing and
hunting, 1970. U.S. Dep. Int., Resour. Publ. 95. 108 p.
180
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SECTION 17
EXPERIMENTAL RESEARCH ON PHENOL INTOXICATION OF AQUATIC
ORGANISMS AND DESTRUCTION OF PHENOL IN MODEL COMMUNITIES
M.M. Kamshilov and B.A. Flerov1
In the first part of the investigation, some particular and general pro-
blems of aquatic toxicology have been studied on a model of phenol intoxica-
tion of aquatic organisms. Among the problems investigated were compara-
tive resistance of aquatic organisms, the role of biotic and abiotic factors
in determining, resistance, effects of different concentrations of the toxi-
cant on biological and physiological processes of aquatic organisms, and
ability of organisms to adapt. In the second part of the investigations,
destruction of phenol in different model ecosystems has been studied.
Workers from the laboratory of the physiology of lower organisms took part
in the investigation. (V.A. Alekseyev, L.A. Baronkina, P.A. Gdovskiy, N.V.
Goryacheva, B.F. Zhukov, L.I. Zakharova, V. Ya. Kostyayev, N.A. Lapteva,
G.A. Lukina, V. Ye. Matey, F.I. Mezhnin, V.R. Mikryakov, T.F. Mikryakova,
G.E. Flerova).
The principal results of the first part of the investigations are pre-
sented in the following text.
BACTERIA
Phenol, even in small concentrations (10 mg/liter) produces an inhibi-
tory effect on bacteria of the genera Bacterium, Cornybacterium, and
Micrococcus. Bacteria of the genera Pseudomonas and Micobacterium were
more resistant to the toxicant: At a concentration of 50-200 mg/liter, an
increased development took place. At a concentration of 1000 mg/liter,
phenol exerts bacteriostatic effect on these genera.
ALGAE
The most resistant algae are green algae. Retarding their growth takes
place at 30-60 mg/liter, complete inhibition at 300-600 mg/liter. Least
resistant are the chrysophyte algae. Complete inhibition of their growth
\
Institute of Biology for Inland Waters, Acad. Sci. USSR.
181
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is observed at 8-15 rug/1 Her. According to resistance, green algae and
diatoms are intermediate. In blue green algae, reproduction ceases at 100
mg/liter and in diatoms at 200 mg/liter.
Complete inhibition of photosynthesis in all algae occurs at concen-
trations from 700 to 1400 mg/liter. Resistance of algae to phenol is to a
considerable degree determined by the composition of the medium. The
richer the medium in nutrients, the more resistant are the algae grown on
it.
Chlorella has been studied in more detail. Inhibition of its growth
commences at 100 mg/liter. At 1500 mg/liter reproduction ceases. The
toxic effect of phenol is directly proportional to the light intensity.
Resistance of various strains of Chlorella is determined by their sensi-
tivity to light. Respiratory processes in Chlorella are more tolerant to
the influence of phenol than photosynthetic processes.
INVERTEBRATES
Different crustaceans, molluscs, aquatic insects and arachnids were
used as experimental organisms. Resistance of invertebrates to phenol
varied widely: At 48 hrs exposure and a temperature of 20 C, LC50s fluc-
tuate from 2 to 2000 mg/liter; however, the organisms may be divided into
three groups (Table 1) according to their resistance.
There are low resistance invertebrates—larvae of caddis flies (genus
Trichoptera), stoneflies (order Plecoptera), mayflies (order Ephemerop-
tera), beetles (order Coleoptera), damselflies (order Odonata), blackflies
(order Simuliidae) and crustaceans (suborder Cladocera). Their LC50 are in
the range of 2-50 mg/liter.
Invertebrates of intermediate resistance are larvae of culicidflies
(family Culicidae, Subfamily Orthocladiinae), order Megaloptera and image
bugs (genera Sigara, Gerris). Their LC50s are in the range of 50-300
mg/liter.
There are highly resistant invertebrates—larvae of other flies (with
the exception of the families Simuliidae, Culcidae, subfamily Ortho-
cladiinae), image bugs (with the exception of the genera Sigara, Gerris),
of beetles, molluscs, spiders and mites. Their LCbO are in the range of
400-2000 mg/liter. Aquatic invertebrates with respect to their resis-
tance to other toxic substances (pesticides) are arranged approximately in
the same order.
A comparison of the resistance of organisms indicates that Sida crystal-
lina from the Cladocera family may serve as a susceptibility test-object
for toxicological investigations. This organism, like Daphnia, is easy to
rear under laboratory conditions.
182
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TABLE
RESISTANCE OF AQUATIC INVERTEBRATES TO PHENOL
Degree of
Resistance
Species
48 hrs
LC50 mg/1
Low
Resistant
Intermediary
Resistant
TRICHOPTERA (larva)
Limnophilus flavicornis Ebr.
.Leptocerus aterrimus Steph.
Phryganea striata L.
Limnophilus stigma Curt.
EPHEMEROPTERA (larva)
Baetis sp.
Cloeon dipterum L.
Siphlonurus linnaeanus Eat.
PLECOPTERA (larva)
Nemura marginata Pict.
COLEOPTERA (larva)
Acilius sulcatus L.
Ilybius angustior Gyl 1.
Dytiscus marginal is L.
ODONATA (larva)
Platyckemis pennipes Pall.
Coenagrion pulchellum V.d.L.
Lestes dryas Kir.
Aeschna cyanea Mul1.
Sympetrum flaveolum L.
DIPTERA (larva)
Ensimul ium ex.
gr. aureum Fries.
CLADOCERA
Sida crystal!ina O.F. Muller.
Daphnia longispina O.F. Muller.
Chydorus sphaericus O.F. Muller.
Daphnia pulex De Geer,
Bosmina coregoni Baird.
Ceriodaphnia pulchella g. Sars
Lynceus brachyurus O.F- Muller.
DIPTERA (larva)
Aedes caprius Ludl.
Cryophila
Mochlonyx
lappom'ca Mart.
culiciformis De
Geer
Orthocladius sp.
Anopheles
Chaoborus
maculipennis
crystal!inus
Meig.
De Geer.
2
2
2
7
2
5
22
16
46
46
24
28
30
30
30
16
6
18
20
36
36
42
47
50
50
50
100
190
240
183.
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TABLE
Degree of
Resistance
RESISTANCE OF AQUATIC INVERTEBRATES TO PHENOL (con't.)
Species
48 hrs
LC50 mg/1
Intermediary
Resistant
MEGALOPTERA (larva)
Sialis flavilatera L.
HEMIPTERA (image)
Sigara striata L.
Gerris lacustris L.
280
165
200
Highly
Resistant
DIPTERA (larva)
Ablabesmyia monilis L.
Chironomus plumosus L.
Psectrocladius ex gr. psilopterus Kieff.
Tri goma sp.
Eri stall's sp.
HEMIPTERA (image)
Notonecta glauca L.
Naucoris^ cimicoides L.
COLEOPTERA (image)
lus flavicollis^ Sturm.
^ ^
Hydrobius fuscipes L.
Gyrinus marinus Gyl 1 .
Coel ambus novemlineatus Steph.
Ilybius angustior Gyl 1 .
Dytiscus marginal is L.
BIVALVIA
Dreissena
polymorpha Pall
Anodonta piscialis Niles.
Unio pictorum L.
Unio tumidus Phi 11.
Sphaerium corneum
ARANCINA (image)
Argyroneta aquatica Cl.
ACARIFORMES (image)
Hygrobates longipalpis Herm.
Hydrachna marita Wainst.
Limnesia undulata Mul 1
Piona nodata Mul 1.
Eyl a i s hamata Koen.
LimnesiA maculata Mul 1.
Hydrodroma despiciens Mul 1.
400
530
830
830
2000
450
500
440
860
1000
1000
1000
1800
1000
1000
1000
1000
1000
1500
440
440
660
660
660
900
1180
184
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TABLE 1. RESISTANCE OF AQUATIC INVERTEBRATES TO PHENOL (con't.)
Degree of 48 hrs.
Resistance Species LC50 mg/1
Highly Piona coccinea Koch 1500
Resistant Hydryphantes ruber De Geer. 1680
Limnochares aquatica L. 1560
Mideopsis orbicularis Mull. 1720
Arrhenurus globator Mull. 1840
185
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The difference in resistance of aquatic invertebrates is determined by
many factors. The most important of these are morpho-physiological (size
of external coverings, their permeability, peculiarities of their respira-
tory system, their surface) and behavioral (ability to avoid toxicant,
general activity in a toxic environment) characteristics.
When there are sublethal concentrations (less 5 mg/liter) on Daphnia
longispina, behavior and ^C trace food consumption is initially disrupted,
and later on embryogenesis and fertility are disrupted. Thus, physiologi-
cal indices in intertebrates are good specific tests for water toxicity.
FISH
Three stages of phenol intoxication in fish have been observed: (1)
disorderly general motor activity, (2) loss of equilibrium and, (3) cessa-
tion of motor activity and respiration. The main stages of poisoning are
the same for freshwater fish, but the degree of manifestation and duration
of intoxication varied in different species of fish.
The species characteristics of susceptibility (initial reaction to the
action of the toxicant) and resistance of freshwater fish to phenol have
been found.
Regarding susceptibility, the fish species may be arranged in the
following order of increasing sensitivity—crucian carp, blue bream, bur-
bot, bream, perch, pike, ruffe, roach, trout. Regarding resistance, the
following descending order is observed—crucian carp, roach, bream, blue
bream, pike, ruffe, perch, burbot, trout. It has been observed that high
susceptibility did not always correlate with low resistance and vice versa.
For example, roach (Rutilis rutilis L.) is highly susceptible and highly
resistant; burbot (Lota Iota L.), is on the contrary, has low suscepti-
bility and low resistance.
Fry are the most resistant, mature fishes are the least. Differences
in fish age are leveled with an increase in concentrations. Resistance to
phenol decreases with significant increase in body weight. Resistance of
fish to a toxicant is considerably less in summer than in winter-
The role of basic environmental factors in fish resistance to phenol
has been demonstrated. Resistance of fish falls with a decrease in the
dissolved 02 content and with an increase in temperature. Water hardness
and pH influence the resistance of fish to phenol practically not at all.
The adaptation of fish to the toxicant has also been investigated.
Fish preliminarily exposed to sublethal concentrations of phenol for
different time periods (from a week to more than a year) when put into a
solution with extremely toxic concentrations showed a lesser resistance to
lethal concentrations than those not exposed. This proves that a fish can
not individually adapt to phenol. The adaptation is realized by selecting
the most resistant individuals. The first generation of fish (guppies) was
186
-------
already 5 times as resistant as the initial generation. The selection is
of a nonspecific character. After the selection for resistance to phenol,
the fish were simultaneously more resistant to another toxicant, polychlor-
pinene (Figure 1).
We have also investigated the effect of sublethal concentrations on dif-
ferent vital biological characteristics of fish: Behavior (general moto ac-
tivity, feeding, defense, sexual behavior, conditioned responses) growth
rate, and reproduction. Physiological functions such as bioelectric ac-
tivity of nerve and muscle systems, neurosecretion in the hyothalamo-hypo-
physial system, immunological reactions, and blood analysis have been
included as indices of intoxication. Pathologic changes in the structure
of fish organs have been studied under lethal concentrations.
Chronic phenol influence causes considerable changes in behavior, and
first of all in conditioned reflexes and then in other functions of the or-
ganism (Figure 1). Conditioned reflex method may be recommended as a
highly sensitive test for determining water toxicity. Under a prolonged
toxicant influence, inhibition of feeding, sexual behavior, defense func-
tions, and growth rate takes place.
After Kuba's (1969) investigations, it has been established, on the
basis of electrophysiological data, that phenol acts primarily on neuromus-
cular synapses, increasing the frequency in formation of miniature terminal
disc potentials (Figure 2). Concentrations of phenol lethal for the whole
organism do not produce a noticeable effect on the total action potential
of peripheral nerves nor on the evoked responses of the fish brain (olfac-
tory bulb).
Pathomorphological changes in fish organs and hypothalamo-hypophysial
system reactions under phenol intoxication are of general and nonspecific
character. Thus, they are unlikely to be of any significance for post-
mortem diagnosis to toxicosis.
The principal results of the second part of the investigation are pre-
sented below.
DESTRUCTION OF PHENOL IN MODEL COMMUNITIES
The destruction of phenol was studied in 28 aquaria, 3-15 liter in vol-
ume, filled with river water and sand. Several components (aquatic organ-
isms, mineral fertilizers, ultraviolet radiation) were introduced into the
aquaria.
One series of experiments performed in 6 aquaria may be cited as an ex-
ample.
Aquarium 1 - without aquatic organisms
187
-------
00
00
Complete mortality
Partial mortality
Inhibition of ability to reproduce
Inhibition of sexual behavior
Retardation of conditioned reflexes
0 5 10 15 20 25 30 35
Phenol Concentration, mg/l
Fig. 1. Sequence of changes in biological indices of Lebistes reticulatus (P) in
phenol concentrations.
-------
. G k H I
^r v~rr
nr
O.I sec
M
N
L lyihhyi. in iiL JU
_ |l mv
O.I sec
Fig. 2. Effect of phenol on bioelectrical activity of peripheral, central
nerve system and neuromuscular conjunction of fisn.
A-F Influence of phenol on action potential of olfactory
nerve in pike.
A-control, B-500 mg/1, C-1000 mg/1 (10 min), D-2000
mg/1 (1 min), E-wash out by Ringer solution (5 min),
F-wash out by Ringer solution (15 min).
G-L Resistance to phenol of the potential generated by
olfactory bulb in pike (brain membrane is removed).
G-control, H-10 mg/1, 1-100 mg/1, J-500 mg/1, K-1000
mg/1, L-wash out.
M,N Influence of phenol on miniature potentials of terminal
discs in the red muscle of the pectoral fin in the carp.
M-control, N-5 min after application of mg/1 of phenol.
189
-------
2 - with filamentous algae (mougeotia genuflexa) and
oligochaets (Lumbriculus variegatus).
3 - with duckweed (Lemna trisulca).
4 - with elodea (Elodea canadensis).
5 - with elodea, duckweed and molluscs (Limnea
stagnalis and PI anorbis sp.).
6 - with the same content as in 5.
It is quite natural that some microscopic organisms were introduced to-
gether with the macrocomponents (Elodea, duckweed, molluscs, oligochaets)
and with the water and sand.
Phenol was added to the five aquaria first in doses of 1 mg/liter (1st
period - 59 days), then in doses of 5 mg/liter (2nd period - 112 days), and
then in doses of 10 mg/liter (3rd period - 194 days). Altogether, 2388 mg
of phenol per liter of medium were added to all the experimental aquaria
during the three periods (365 days). A year after the beginning of the ex-
periment, the addition of phenol was stopped (4th period - 140 days).
Aquarium #6 was the control; phenol was not added to it.
The concentration of phenol was systematically measured by pyramidon
method (Kaplin and Fesenko, 1962). The contents of nitrogen, phosphorus,
oxygen, pH and BOD5 in the medium were determined less regularly. The
bacterial population of the aquaria, especially the number of saprophytic
bacteria decomposing phenol, was constantly controlled. The numbers and
species composition of colorless flagellates, infusoria, algae and fungi
were also determined.
As seen in Figure 3 which presents data on accumulation and decomposi-
tion of phenol, the destruction of this toxicant takes place faster in the
aquarium having the most diverse composition (5) than in other aquaria.
The results of the second section of the research allows us to draw the
following conclusion.
The rate of phenol destruction is, given other similar conditions, a
function of the diversity of the community taking part in the destruction
process. The living population of a water-body is able to cope with exter-
nal disturbances, acting as a self-regulating system only if it is diverse
enough. The basis of the self-regulation is biotic circulation, i.e., the
same processes which guarantee the yearly repeated cycles of biotic produc-
tion. The presence of oxygen and nutrients (nitrogen salts and phosphorus),
being a very important factor in the destruction of phenol, is by itself
not sufficient to guarantee its effectiveness. A very energetic decomposi-
tion of the toxicant may occur at relatively low values of these factors.
190
-------
200-
50 100
i n
150 2OO 250 300 350 4OO
m
Fig. 3. Accumulation of phenol in model communities.
Abscissa - days from the beginning of the
experiment. Ordinate - concentration of
phenol in mg/1. 1-5 - numbers of aquaria.
Vertical broken lines separate the periods
differing in amount of phenol (mg) added
per liter of medium: 1-1 mg/1; 2-5 mg/1;
3-10 mg/1; 4-0 mg/1. Thick continuous lines
concentration of phenol in various aquaria.
191
-------
A large number of bacteria which destroy (decompose) phenol do not guar-
antee its active decomposition. Only the bacteria involved in biotic cir-
culation are capable of energetically destroying the toxicant (phenol) at
the minimum values of other factors (oxygen, nutrients).
When a toxic substance is added regularly in small portions, a biologi-
cal system is able to decompose it in much greater quantitities than when
the same toxicant is added at once in a large quantity.
To create highly effective detoxicating systems of living organisms, it
is necessary to account for an adaptation period during which complexes of
organism species, capable of effectively decomposing the toxicant, are
established. The duration of the adaptation period is approximately two
months.
When the toxicant is no longer introduced, the ecosystem quickly loses
its ability to decompose it.
REFERENCES
Kaplin, V.T. and Fecenko, N.G. 1962. Calorimetric determination of phenols
using dimethylamine antipyrine (pyramidon) with its content in liter
0.001 mg and higher. From "Contemporary methods of analysis of natural
waters; Moscow.
Kuba, K. 1969. The action of phenol on neuromuscular transmission in the
red muscle of fish. Jap. J. Physiol., 19, pp. 762-774.
192
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SECTION 18
HISTORY OF CHANGES IN FISH SPECIES OF THE GREAT LAKES
John F. Carr
INTRODUCTION
Changes began in the fish-species complex in the Laurentian Great Lakes
almost immediately after the first permanent settlers arrived in the basin
in the early 1800's. Changes occurred slowly at first, but accelerated
with the increased activities of man. These changes continue today and
will continue in all probability for decades or even centuries because
man's manipulations of the environment are continuing.
The Great Lakes are young; only about 10,000 years have passed since
the melting of the glaciers. Youthful lakes such as these are generally
characterized by low biological productivity, low nutrient content, and
high transparency; they are often deep and cold. So are the Laurentian
Great Lakes even today. With the exception of a few areas, the waters of
the Great Lakes are of excellent quality and can be used as potable water
without treatment. Yet man's impact on these lakes, especially on the fish
populations, has been so drastic that the Laurentian Great Lakes have been
used as worldwide symbols of accelerated aging. Some scientists have esti-
mated that the lakes, especially Erie and Ontario, have aged more in the
past 150 years than in the preceding 10,000 years. That changes of this
magnitude could occur in lakes as large as the Great Lakes was not
considered possible only a few years ago. Today, however, we are beginning
to realize the tremendous capacity we possess to change (usually to our
detriment) even the oceans and the atmosphere.
The purpose of this paper is to discuss the changes which have taken
place in the fish populations of the Great Lakes and the stresses which
have caused these changes. It has become obvious that many of the causes
of the declines are the results of deliberate actions rather than subtle un-
predictable factors.
The stresses which have been placed on the fish communities of the
Great Lakes have been sequential and reflect the progress of man's occupa-
tion of the basin and his technological development. The most obvious and
primary direct stress has been the intensive and selective exploitation of
the fish stocks. This stress began early in the 19th century and continues
to some degree today. Environmental stresses have not been as direct or as
obvious, but were present as early as 1830 and have been additive as well
193
-------
as continuous. Environmental stress on the fish of the Great Lakes has
been of five general types:
(1) Physical stress resulting from modification of the watershed
by deforestation, blockage of tributaries, and drainage of
marshes. This primarily affected anadromous species and be-
gan during the period of low human population density.
(2) Biological stress caused by the introduction and colonization
of exotic species. Introduction of new species began before
1900 and continues today.
(3) Chemical stress (first phase) in the form of oxygen-consuming
organic material dumped into tributaries and bays, and
increased plant nutrients in the inshore areas.
(4) Chemical stress (second phase) caused by toxic chemicals such
as chlorinated hydrocarbons and heavy metals.
(5) Thermal stress - more a future concern than a present concern.
The direct effects of the environmental changes on the fish popula-
tions are seldom observed and perhaps rarely occur- Indirect effects of
these changes are often cited, but only occasionally quantified. In lakes
as large as the Great Lakes, cause and effect are separated in time and
distance to an extent that only after an event can the two be linked. This
is the situation with the continuous change in the abundance of Great Lakes
fish species.
We all recognize the fact that the stresses, be they exploitation, des-
truction of spawning grounds, oxygen depletion, increased water temperature,
change in available food, or competition with introduced species, cannot
often be isolated and analyzed separately. This paper is a summary of the
changes that have occurred in the fish communities of the Laurentian Great
Lakes from the early 19th century to the present. The changing composition
of fish populations in the Great Lakes has been the subject in recent years
of many articles in scientific and popular publications. The most exhaus-
tive discussion of these changes occurred in a recent (1971) international
symposium on "Salmonid Communities in Oligotrophic Lakes" (SCOL). These
papers were published as a special issue of the Journal of Fisheries Re-
search Board of Canada in 1972. This publication contains seven papers ex-
clusively on the Great Lakes, including case histories of each of the five
Laurentian Great Lakes: Superior, Huron, Michigan, Erie, and Ontario
(Figure 1). Other comprehensive papers on changes in Great Lakes fish
species are by Smith (1964, 1968) and Christie (1974).
194
-------
ID
cn
I...
Figure 1. The St. Lawrence Great Lakes with interstate and
international boundaries.
-------
IN THE BEGINNING (TO 1850)
Fish Communities
The fish-species complex in the Great Lakes has changed drastically.
Unlike many large lakes of the world, especially the large African lakes
and Lake Baikal (USSR), the Laurentian Great Lakes have had only 10 thou-
sand years between the retreat of the glaciers and the coming of man to
produce, through evolutionary forces, a complex of species that is unique
to the system.
The Great Lakes system did produce a few unique species in this short
period, indicating that the processes were well underway to further species
diversity. The evidence for this conclusion is best illustrated by the
five endemic species (Smith, 1957; Scott and Grossman, 1-973), all of the
subfamily Coregoninae (whitefish) in the Samonidae (salmon family)-. These
five species listed in descending order of size were:
deepwater Cisco - Coregonus johannae
longjaw Cisco - Coregonus alpenae
shortnose Cisco - Coregonus reighardi
kiyi - Coregonus kiyi
bloater - Coregonus hoyi
According to Scott and Grossman (1973) all five species were found in Lakes
Huron and Michigan, four in Lake Superior, three in Lake Ontario, and one
in Lake Erie.
In addition to the five endemic species of ciscos, these wider ranging
species were also present: lake herring (Coregonus artedii); blackfin
cisco (Coregonus nigripinnus); and shortjaw Cisco (Coregonus zenithicus).
These eight species of ciscos, together with the lake whitefish (Coregonus
clupeaformis) and round whitefish (Coregonus cylindraceum), characterized
the Great Lakes fish community. Most of the species of the whitefish sub-
family, especially the ciscos, were inhabitants of deep, cold water and
therefore reached their greatest diversity in Lakes Superior, Huron,
Michigan, and Ontario (Table 1). The dramatic alteration in the species
complex of deepwater ciscos that subsequently occurred was documented by
Smith (1964) for Lake Michigan.
In addition to the Coregonines other groups and species were abundant
in the lakes. The dominant predators of the open waters present in all five
lakes were the lake trout (Salvelinus namaycush) and burbot (Lota Iota). In
the bays and nearshore areas were: lake sturgeon (Acipenser fulvescens);
northern pike (Esox lucius); suckers (primarily Catastomus catastomus and
C. commersoni); channel catfish (Ictalurus punctatus); bullheads (Ictalurus
spp.); white bass Morone chrysops; freshwater drum (Aplodinotus grunniens);
and three species of the perch family: yellow perch (Perca flavescens);
walleye (Stizostedion vitreum); and sauger (Stizostedion canadense). All
of these species have been historically of commercial significance. The
Atlanta salmon (Salmo salar) and American eel (Anguilla rostrata) were also
abundant and became commercially important only in Lake Ontario.
196
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TABLE 1. DIMENSIONS OF THE GREAT LAKES
10
Lake
Superior
Michigan
Huron
St. Clair
Erie
Ontario
Length
(miles)
350
307
206
26
241
193
Breadth
(miles)
160
118
183
24
57
53
Water
surface
(miles)
31 ,820
22,400
23,010
490
9,930
7,520
Drainage
basin
(miles)
80,000
67,860
72,620
7,430
32,490
34,800
Average surface
elevation above
mean sea
level since 1860
(ft)
602.20
580,54
580.54
574.88
572,34
246.03
Mean
discharge
(cfs)
73,300
55,000
177,900
178,000
195,800
233,900
Maximum
depth
(ft)
1,333
923
750
21
210
802
Mean
depth
(ft)
487
276
195
10
58
283
From Beeton and Chandler, 1963.
-------
Knowledge of the presence and relative abundance of the species listed
thus far is based on records of the harvest of these species by commercial
fishermen. Species not of great commercial value were, of course, also pre-
sent in the lakes. A complete list of all species known to have been
present in the lakes would be too long to include for the purposes of this
paper. Our knowledge of changes in abundance of a few of these species,
however, is sufficient to warrent their inclusion. An abundant forage
species in the deeper water of all the lakes was the fourhorn sculpin
(Myoxocephalus quadricornis). The slimy sculpin (Cottus cognatus) in-
habited intermediate depths in all lakes. The inshore waters contained a
variety of species of several families, especially the Cyprinidae. Among
the more abundant species were the emerald shiner (Notropis atherinoides)
and the spottail shiner (Notropis hudsonius).
Thus, when settlement began in the first half of the 19th century, the
lakes were occupied by a supposedly stable community of fish species which
inhabited all niches from the deepest waters of the subartic Lake Superior
to the shallow bays and marshes of Lake Erie.
Environmental Conditions
Physiochemical conditions of the lakes were not measured before the end
of the 19th century. Beeton and Edmondson (1972) used as a basis for evalu-
ating the "natural" chemical condition the limited chemical data available
about 1900 as indicative of the pristine quality of the lakes (Table 2).
TABLE 2. ESTIMATED AVERAGE CONCENTRATION OF DISSOLVED
CHEMICAL CONSTITUENTS IN THE GREAT LAKES
PRIOR TO 1900 (EXPRESSED IN MG/LITER)9
Lake
Superior
Michigan
Huron
Erie
Ontario
Total
dissolved
solids
60
128
108
142
140
Calcium
13
34
24
31
31
Sulphate
4
5
6
13
15
Chloride
2
2
4
7
7
Sodium
and
potassium
3
-
4
7
6
From Beeton, 1969.
198
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Based on these few constituents, Lakes Michigan, Erie, and Ontario were
similar in chemical composition (Table 2). Lake Superior's chemical levels
were substantially below all the others. Lake Huron receives approximately
41% of its inflow from Lake Superior, 31% from Lake Michigan, and 28% from
the Lake Huron basin. Apparently the basin is a major contributor to the
characteristics of the water quality, because the total dissolved solids
are much higher than would be expected based only on a mixture of the
waters from Lakes Superior and Michigan. The Lake Erie watershed apparently
contributed significant amounts of calcium, sulphate, chlorides, and sodium-
potassium to the waters flowing into Lake Ontario, because their chemical
characteristics are nearly identical. Although differences between lakes
were large compared to the range between other natural bodies of water, the
five Great Lakes were remarkably similar.
Analysis of nutrient concentrations for the Great Lakes has been made
only in recent years; therefore, base levels are only now being established.
Estimates based on recent values for phosphorus and nitrogen indicate that
the levels in the mid-1800's were less than 10 yg/liter for phosphorus and
usually less than 1 mg/liter for total nitrogen in the three upper lakes
(Superior, Huron, and Michigan). The Great Lakes in the 1900's would have
been classified as oligotrophic (as defined by Hutchinson, 1957) with the
probable exception of Lake Erie.
The order of the lakes, if listed from the greatest to the least
fishery productive potential in the 1800's, probably would be Lake Erie,
Lake Ontario, Lake Michigan, Lake Huron, and Lake Superior (Figure 2).
THE INITIAL IMPACT OF SETTLEMENT (1850-1900)
Changes in Fish Populations
The first environmental stresses on the Great Lakes ecosystem were pri-
marily caused by physical alterations in the basin, particularly in the
lower lakes. These alterations were deforestation of the watershed and
siltation and blockage of streams. These changes mainly affected the
tributaries and consequently the obligate anadromous species. Christie
(1972) reported the Atlantic salmon had begun to decline as early as 1830
in Lake Ontario and was extinct, or nearly so, by 1900. Documentation of
the early proliferation of mills and dams was given by Christie (1972)
based on data from Richardson (1944). On the Ganaraska watershed (one of
the larger Canadian rivers tributary to Lake Ontario) at least two sawmills,
two grist mills, and two dams had been constructed by 1800. Construction
of the mills and dams increased rapidly, reaching a maximum of 34 sawmills,
19 grist mills, 4 woolen mills, and 34 dams by 1860. In 1930, 15 dams
still remained on this single tributary. Christie (1972) considered the
elimination of the Lake Ontario Atlantic salmon stock as the best known
example of the effects of despoilation on a species habitat.
The lake sturgeon population in all the lakes was greatly reduced
during this period. Prior to 1903 the annual commercial production fluctu-
ated between 100,000 pounds and 500,000 pounds in Lake Ontario (Baldwin and
199
-------
1.0
10.0-
UJ
a:
CO
Q
O
Q.
9.0
2 ao
LJ
Q.
Q 7.0
LJ
O
8 6-°
o:
Q.
i 5.0
CO
u_
it 4.0
3.0
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Superior
I
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I
o
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-------
Saalfeld, 1962). After 1910 production never exceeded 25,000 pounds and
averaged near 15,000 pounds for many years. In Lake Erie production
dropped from 1 - 5 million pounds in the late 1800's to less than 10,000
pounds after 1910 (Baldwin and Saalfeld, 1962). Similar magnitudes of de-
cline occurred during the same period in Lake Huron, Michigan, and Superior.
The cause-and-effect relationship apparent between watershed and stream
modification and Atlantic salmon extinction was not as direct with the
sturgeon. The environmental requirements of the sturgeon were not as
narrow as those of the Atlantic salmon; however, the slow growth rate and
late maturity of the sturgeon were also factors in this species, inability
to recover from even low exploitation rates.
Christie (1972) also lists the blackfin Cisco (Coregonus nigripinnus)
as a species that became either extinct or greatly reduced in Lake Ontario
before 1900. Wells and McLain (1972) infer a sharp decline in the
abundance of this species in Lake Michigan in the early 1900's. The black-
fin was considered commercially extinct in Lake Superior by 1910 (Lawrie
and Rahrer, 1972). This species of Cisco inhabited the deep, open waters
of the lakes; consequently, environmental modification of the tributary and
watershed was not a factor in their decline. This species was the largest
of the ciscos, and selective exploitation for it was the probable cause of
its decline (Wells and McLain, 1972).
The lake trout population in Lake Erie was also decimated during this
period. Hartman (1972), in discussing this species in Lake Erie, states:
"Perhaps the decline of the lake trout population to near extinction best
illustrates the effect of essentially one stress: intensive exploitation."
Apparently, environmental stress was not a factor in the decline of lake-
dwelling species during the 1800's.
In addition to the effective loss of at least one species during the
1800's, several new species became abundant. Christie (1972) listed the
following species as becoming established in Lake Ontario before 1900:
alewife (Alosa pseudoharengus); gizzard shad (Dorosoma cepedianum); brown
trout (Salmo trutta); carp (Cyprinus carpio); and the goldfish (Carassius
auratus). Some of these species were also introduced to the other lakes
during this period. There was a flourishing fishery for carp in Lake Erie
by 1899, when over 3.5 million pounds were landed (Baldwin and Saalfeld,
1962).
Environmental Changes (Age of Physical Alterations)
Man's effect on the Great Lakes ecosystem in the last half of the 19th
century was dramatic and permanent. He removed the forest, built dams,
constructed mills, directly exploited the fish, and opened new and more
direct passage between the ocean and the lakes, as well as between Lake
Ontario and the upper lakes (Figure 3).
The effects of these physical modifications of the environment ranged
from immediate (Atlantic salmon extinction) to long-term (invasion of marine
species). The Erie Canal, which provided a connection between the Atlantic
Ocean and Lake Ontario, was opened in 1819 and extended to Lake Erie in
201
-------
5/ Lawrence
Great Lakes
ro
o
ro
swego R
Oneido L
Mohawk R
\Erie
Canal
Cflyugo L.
Seneca L
Figure 3. The St. Lawrence Great Lakes showing canals between
central Atlantic Ocean and the lakes.
-------
1825. The Welland Canal, which was opened in 1829, connected Lake Ontario
to the upper lakes. Previously the passage of fish between Lakes Ontario
and Erie was blocked by Niagara Falls. Although no biological changes were
noticed for many years after the opening of the canals, the stage was set
for dramatic and catastrophic changes to occur decades later (Aron and
Smith, 1971).
The chemical characteristics of the open waters of the lakes were
assumed to be essentially the same at the end of the 19th century as at the
beginning (Beeton and Edmondson, 1972). Although the settlement of the
Great Lakes basin had advanced rapidly in the 19th century, from a popula-
tion of a few thousand early in the century to over 10 million in 1900
(Table 3), the effects on water quality in the lakes were yet to be felt.
By 1900 man, in less than 100 years, had placed the following stresses
on the biological communities of the Great Lakes: siltation of streams;
blockage of tributaries; increase in stream temperatures; and establishment
of exotic species. In addition, he had removed barriers to migration
between the lakes; had established fisheries capable of overexploitation of
most species in all the lakes; and had begun using the lakes as the
receiver of man's domestic and industrial waste.
TABLE 3. ESTIMATED POPULATION (MILLIONS) IN GREAT LAKES
BASIN - 1900-I9609
Lake basin 1900 1925 1950 1960
Superior 0.4 (50)5 0.6 (33) 0.8 (12) 0.9
Michigan 4.0 (-20) 3.2 (50) 4.8 (23) 5.9
Huron 1.0 (20) 1.2 (25) 1.5 (33) 2.0
Erie 3.0 (93) 5.8 (48) 8.6 (17) 10.1
Ontario 2.0 (25) 2.5 (20) 3.0 (33) 4.0
Total 10.4 (28) 13.3 (41) 18.7 (22) 22.9
aFrom Beeton, 1969,
Numbers in parentheses indicate percentage change in
the ensuing time interval.
203
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ENRICHMENT AND INVASION (1900-1950)
The stage was set and the signs were present in 1900 for what was to
follow. Changes in the biological, chemical, and physical environment of
the Great Lakes became the rule and not the exception. The records of
these changes, unfortunately, are incomplete, often inaccurate, and, for
the fish populations, often not a true representation of species abundance.
The analysis of changes in the abundance of species until recently was
based on the reported catch of commercial fishermen. High prices often
maintained high catches in.the face of a decreasing abundance. Conversely,
low production often was due to low prices and lack of demand for a species
rather than low population levels. Despite these handicaps, the changing
conditions often became too obvious to be ignored.
Changes in fish-species composition, losses and gains, differ in time
between the lakes, but the sequence of species change often was similar
(Smith, unpublished manuscript). In general, the species that declined
were those most sought after by the commercial fishery. A few significant
exceptions exist to this generalization, and it is the exceptions which
clearly indicate stresses other than fishing on the biological communities
of the Great Lakes.
Changes which occurred in native fish species of commercial interest
between 1900 and 1971 are summarized in Table 4. Detailed discussion of
these declines by species and lakes appear in Smith (1968), the papers of
the SCOL Symposium (1972), and Christie (1974). The data presented in
Table 4 refer to production trends in the total lake and, therefore, are
not descriptive of events in the unique ecological areas of each lake such
as the Bay of Quinte in Lake Ontario, the western basin of Lake Erie,
Saginaw Bay of Lake Huron, or Green Bay of Lake Michigan (Figure 1). These
geographic areas were, and are, more shallow and productive and warmer than
the open portions of the lake to which they are connected. The fish-species
complex here was also more diverse than in the open lake, containing many
warmwater species, especially the centrarchids and percids.
Ecological and Cultural Changes, 1900-1925
During the first quarter of the 20th century, the northern pike fishery
was reduced to a fraction of former production; lake whitefish in Lake
Ontario and lake herring in Lake Erie began declining; the first sea
lamprey was reported in Lake Erie; and the first rainbow smelt were found
in Lakes Michigan and Huron. The gains and losses in these and other
species were to be repeated many times in the next 50 years in the other
lakes.
The introduction of the smelt into Crystal Lake in the drainage basin
of Lake Michigan was deliberate, but its establishment in Lake Michigan was
not contemplated, nor was its rapid spread to other Great Lakes. The sea
lamprey reached Lake Erie nearly 100 years after the Wei land Canal was
opened and established itself in the upper lakes. The beginning of the
declines in lake whitefish and lake herring were, of course, undetected at
the time and thus alarmed no one.
204
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TABLE 4. SUMMARY OF FISH SPECIES DECLINE IN THE GREAT LAKES
BY YEAR, LAKE, AND CURRENT COMMERCIAL PRODUCTION
Species
and
lake
Lake trout
Ontario
Erie
Huron
Michigan
Superior
Lake whitefish
Ontario
Erie
Cisco (chubs)
Ontario
Huronb
Michigan
Superior
Lake herring
Ontario
Erie fa
Huron
Michigan
Superior
Northern pike
Ontario
Erie
Burbot
Ontario
Erie
Michigan
Blue pike
Ontario
Erie
Sauger
Erie
Huron
Beginning
of
decline
(year)
1928
?
1935
1943
1950
1924
1953
1941
1961
1970
1965
1941
1924
1939
1952
1961
1933
1914
j
1930°
1947
1948d
1952
1955
1945
1935
Product first
below 100,000
pounds
(year)
1942
1900
1945
1950
c
1966
1960
1950
1970
c
c
1953
1958
1957
1963
c
1938
1924
1934®
1961e
1959
1955
1959
1955
1937e
1974
production
(1 ,000's pounds)
1
0
1
37
526
16
1
0
50
3,267
1,926
32
0
2
6
2,186
21
15
0
°f
230T
0
0
0
0
for
H
L
H
H
H
L
L
L
L
L
L
L
L
L
L
M
L
L
L
L
H
L
L -
M
L
Potential
recovery (H,M,L)a
and reason
stocking and sea
lamprey control (s.l.c.)
environmental
stocking & s.l.c.
stocking & s.l.c.
stocking & s.l.c.
environmental
environmental
environmental
environmental
environmental
environmental
environmental
environmental
environmental
environmental
stocking
environmental
environmental
environmental
environmental
s.l.c.
environmental
environmental
stocking
environmental
H, M, and L indicate high, medium, and low potential for recovery, respectively.
Excludes Georgian Bay and North Channel.
cRemains above 100,000 pounds.
Production normally less than 100,000 pounds.
Production less than 10,000 pounds.
First exceeded 100,000 pounds in 1973.
205
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Environmental changes, except in streams and bays near centers of high
human population density, were nearly undetectable in 1925 in the lakes
proper. Changes had occurred, however, in Lakes Michigan, Erie, and Ontario
in the few chemical constituents for which data are available (Table 5).
The absolute concentrations of these constituents, even the highest levels,
were well below levels of ecological or toxicological concern. The rates
of change, especially in total dissolved solids, sulphur, and chloride,
however, are staggering considering the tremendous volume of water which
had been changed as much as 160% in only 25 years.
TABLE 5. ESTIMATED AVERAGE CONCENTRATION OF DISSOLVED CHEMICAL
CONSTITUENTS IN THE GREAT LAKES IN 1925 (EXPRESSED IN
MG/LITER)a WITH PERCENTAGE CHANGE AFTER 1900 IN PAREN-
THESES
Lake
Superior
Michigan
Huron
Erie
Ontario
Total
dissolved
solids
58 (-3)
143 (+12)
108 (0)
146 (+3)
149 (+6)
Calcium
13 (0)
34 (0)
24 (0)
33 (+6)
34 (+10)
Sulphate
4 (0)
13 (+160)
9 (+50)
16 (+23)
18 (+20)
Chloride
2 (0)
4 (+100)
4 (0)
11 (57)
11 (57)
Sodium
and
potassium
3 (0)
-
4 (0)
7 (0)
7 (17)
aFrom Beeton, 1969.
The increase in population growth in the basins of the Great Lakes was
approximately 28% between 1900 and 1925 (Table 3). Growth in numbers was
greatest in the basins of Lakes Erie and Ontario. The population in the
Erie basin increased 93% to 5.8 million, and in the Lake Ontario basin the
increase was 25% to 2.5 million. Lake Michigan's basin effectively lost
0.8 million when the Chicago Sanitary Canal, which diverted the waste from
the city to the Mississippi River drainage, was completed in 1900. The in-
creasing urbanization and industrial expansion was the probable cause of
the increase in the chemically conservative ions. Undoubtedly, concentra-
tions of other chemical components also increased, especially the plant
nutrients phosphorus and nitrogen. The load of oxygen-demanding organic
compounds can also be assumed to have increased.
The stresses causing the decreases in some native fish species in the
lower lakes by 1925 were man caused, principally by heavy exploitation.
The role of environmental change, especially water-chemistry change, in re-
duction in fish populations apparently was minor except in the tributaries
206
-------
and bays. The drainage of marshes, however, may have been a significant
factor in the loss of the northern pike as an important commercial species
in Lake Erie. The establishment of the two marine species, smelt and
lamprey, was too recent to have had a measurable impact on other fish
species by 1925.
Environmental and Cultural Changes, 1925-1950
Several catastrophic events affecting the Great Lakes fish stocks
occurred during this period. The most damaging event was the invasion by
the parasitic sea lamprey of all the upper Great Lakes. After at least 50
years in Lake Ontario, the lamprey made its way into Lake Erie in 1923
where it did not flourish because of the lack of suitable spawning streams
and limited deepwater environment. In 1932 the first sea lamprey was re-
ported in Lake Huron; 4 years later the lamprey was in Lake Michigan; and
by 1946 the first report was made of a sea lamprey in Lake Superior (Table
6). Three years after the first sea lamprey was reported in Lake Huron,
the production of lake trout started to decline (1935), and by 1946 (Table
4) the commercial fishery for this species in Lake Huron proper was
finished, although the fishery in Georgian Bay lasted another 9 years.
Lake trout production began to decline in Lake Michigan in 1943 (7 years
after the first lamprey was reported); by 1950 production dropped below 0.1
million pounds (Table 4), and the species was virtually extinct 3 years
later. Only 18 years passed from the time the first sea lamprey was re-
ported in Lake Huron (1932) until the species was commercially extinct in
Lakes Huron and Michigan.
The demise of the lake trout population in Lake Ontario and the role of
the sea lamprey is more complicated than in the upper lakes. Whether the
sea lamprey was endemic to Lake Ontario (Christie, 1972), or became
established after the opening of the Erie Canal (Smith, 1974), at least 75
years passed before the lake trout production began its final decline
(1928). A substantial fishery continued, however, for another 10-12
years. The species was last reported in the commercial catch statistics as
late as 1964 (Baldwin and Saalfeld, 1962, with supplement). That the sea
lamprey was a strong factor in the loss of lake trout in Lake Ontario is un-
disputed; the reasons why the struggle lasted so long remain a subject of
speculation.
The sea lamprey's favored prey was the lake trout, but other species as
well were victims of this marine invader. Larger individuals of lake white-
fish, ciscos, lake herring, suckers, and burbot were attacked by the sea
lamprey. The production of burbot (never a prime commercial species) began
to decline in Lake Ontario in 1930, in Lake Erie in 1947, and in Lakes
Huron and Michigan by 1948 (Table 4). The burbot population became commer-
cially extinct in these lakes about 1960.
Other species also began declining during this period, although the de-
clines were not related to the sea lamprey. The sauger began declining in
Lake Huron (primarily Saginaw Bay) in 1935; 2 years later the species was
commercially extinct. Beeton (1969) gave the reason for the decline as the
development of an environment not suitable for the sauger or the Saginaw
207
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TABLE 6. FISH-SPECIES INTRODUCTIONS IN THE GREAT LAKES BY YEAR
AND LAKE, FIRST YEAR OF COMMERCIAL SIGNIFICANCE, AND
CURRENT PRODUCTION
Species
Sea lamprey
Al ewi f e
Gizzard shad
Coho salmon
Chinook salmon
Rainbow smelt
Carp
Goldfish
White perch
First reported in
First recorded commercial catch
Year
1800's
1921
1936
1946
1873
1931
1933
1949
1953
1900
1968
1968
1967
1966
1966
1969
1970
1967
1967
1967
1923
1925
1930
1931
1935
1885-1895
1880-1895
1880's
1880's
7
7
1950C
Lake Year Lake
Ontario
Huron
Michigan
Superior
Ontario ?
Erie ?
Huron ?
Michigan 1956 Michigan
Superior
Ontario
Ontario
Erie
Huron
Michigan
Superior
Ontario
Erie
Huron
Michigan
Superior
Michigan 1946 Ontario
Huron 1946 Erie
Superior 1935 Huron
Ontario 1931 Ontario
Erie 1935 Superior
Qntarip 1899 Ontario
Erie 1892 Erie
Huron 1899 Huron
Michigan 1893 Michigan
Superior
1929 Erie
Ontario 1955 Ontario
Reached
commercial ,
significance
Year Lake
1920's
1957
1952
1952
1950
1933
1956
1910
1893
1908
1903
1933
1964
Ontario
Michigan
Ontario
Erie
Huron
Michigan
Superior
Ontario
Erie
Huron
Michigan
Erie
Ontario
1974 Production
1 ,000 pounds
1,332
45,508
110
15,766
215
1,748
2,853
411
3,152
739
3,244
54
371
lake
Erie
Michigan
Ontario
Erie
Huron
Michigan
Superior
Ontario
Erie
Huron
Michigan
Erie
Ontario
Production exceeded 100,000 pounds.
bYear of record from Smith, 1972.
cYear of record from Christie, 1972.
208
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TABLE 7. FISH-SPECIES IN THE GREAT LAKES THAT HAVE
EXPERIENCED SEVERE DECLINES, LAKE AFFECTED,
AND SUSPECTED CAUSE OF DECLINE
Species
Principal
lakes affected
Primary cause(s)
of decline
Atlantic salmon Ontario
Sturgeon
Lake trout
Northern pike
Lake herring
Burbot
Cisco (chubs)
Sauger
Lake trout
Walleye
Blue pike
Whitefish
All lakes
Erie
Erie, Ontario
Huron
All lakes
All lakes
All lakes
Huron, Erie
All lakes
(except Erie)
All lakes
Erie and Ontario
All lakes
Yellow perch
Fourhorn sculpin Ontario, Erie
Erie, Huron,
Michigan
Emerald shiner
Michigan
Deterioration and blockage of
streams, exploitation
Exploitation, destruction of spawning
streams
Exploitation
Destruction of spawning areas, ex-
ploitation
Exploitation, environmental changes,
competition with introduced species
Sea lamprey, environmental change
Exploitation, competition with intro-
duced species, sea lamprey
Environmental change, exploitation
Sea lamprey, exploitation
Environmental changes, exploitation,
destruction of spawning streams
Environmental changes, exploitation
Environmental changes, exploitation,
sea lamprey
Competition with introduced species,
exploitation, environmental changes
Competition with introduced species,
environmental change
Competition with introduced species,
environmental change
209
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Bay populations of walleye and whitefish. In Lake Erie the sauger produc-
tion fell below 0.5 million pounds in 1946 (Baldwin and Saalfeld, 1962) for
the first time after nearly 70 years of production between 1 and 6 million
pounds. Environmental changes, plus heavy exploitation, were believed to
be the causes (Table 7).
The decline of the lake herring, historically the most productive
species in the Great Lakes (Smith, 1968), began in Lake Erie in 1925, and
by 1963 this fish had become commercially extinct in all the lakes except
Superior. Heavy exploitation was undoubtedly a factor in the decline of
the lake herring. The role and impact on this decline of introduced ale-
wife and smelt and of environmental factors, however, have not been iso-
lated. The collapse of the lake herring stocks in the mid-1920's was the
event most responsible for stimulating interest and concern in the welfare
of the Great Lakes aquatic environment. This concern was primarily res-
ponsible for identifying the rapid deterioration in the water quality of
Lake Erie, which is discussed in a following section.
Water Quality and Population Changes, 1925-1950
Changes in dissolved chemical constituents continued to accelerate
after 1925 in all the lakes except Superior (Table 8). The absolute values
of these "indicator" chemical parameters are of no toxicological concern,
but again the rate of change indicated substantial inputs from cultural and
industrial sources. Concentrations of these and other chemical compounds
must have been substantial in the receiving waters near the pollution
source. The loss of whitefish, lake herring, sauger, and other species
from the inner portions of Saginaw and Green Bays due to water quality
changes would be expected.
TABLE 8. ESTIMATED AVERAGE CONCENTRATIONS OF DISSOLVED CHEMICAL
CONSTITUENTS IN THE GREAT LAKES (EXPRESSED IN MG/LITER)3
1950 WITH PERCENTAGE CHANGE SINCE 1925 IN PARENTHESES
Lake
Superior
Michigan
Huron
Erie
Ontario
Total
dissolved
solids
56 (-3)
150 (5)
110 (2)
170 (16)
172 (15)
Calcium
13 (0)
34 (0)
24 (0)
38 (15)
38 (12)
Sulphate
4 (0)
17 (31)
13 (44)
23 (44)
25 (39)
Chloride
2 (0)
5 (25)
6 (50)
19 (73)
19 (73)
Sodium
and
potassium
3 (0)
-
4 (0)
9 (29)
10 (43)
From Beeton, 1969.
210
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Population Increases, 1925-1950
The population in the Great Lakes basin had exceeded 18 million by 1950
(Beeton, 1969), an increase of approximately 40% in 25 years (Table 3).
Again, the greatest numerical growth was in the Lake Erie basin with an in-
crease from 5.8 to 8.6 million (48%). The population in Lake Erie's basin
was 46% of the total population and, combined with the Lake Ontario popula-
tion, accounted for 62% of the total. The population in the Lake Michigan
basin was nearly 5 million in 1950. This continued concentration of people
in Michigan, Erie, and Ontario lake basins, with the associated municipal,
industrial, and agricultural wastes, was the primary cause of the accele-
rated rate of increases in dissolved chemical constituents in these lakes.
Lake Erie—Demise is Heralded
The sudden collapse of the Lake Erie lake herring fishery in 1925
awakened the public to the need for scientific investigations into the
causes of the precipitous decline. The magnitude of the decline in lake
herring production was from an average of 26 million pounds per year in the
previous decade to 6 million pounds in 1925, to less than a million pounds
in 1929.
Since environmental factors were thought to be the cause of the lake
herring decline, two intensive limnological studies (Wright, 1955; Fish,
1960) were initiated in 1928. Wright (1955) found unfavorable conditions
in rivers and estuaries, but concluded that environmental changes in the
open waters of the western basin of Lake Erie in 1928-30 had no adverse
effect on the decline of fish stocks. Fish et al. (1960) also found no
environmental basis for the decline of lake herring in the central and
eastern basins in 1928-30. Although neither investigator found measurable
environmental degradation in the open lake, their studies for the first
time established a scientific base line of data on benthic organisms, plank-
ton, and dissolved oxygen. The base line has subsequently been invaluable
in measuring environmental changes in Lake Erie.
The effects of the sea lamprey on the lake trout stocks (previously
discussed) were recognized in the 1940's, and attempts to control the
lamprey began in 1946. Another decade passed, however, before an organized
and substantial program was developed to control this destructive parasite.
CHANGE AND REHABILITATION (1950-1975)
Fish Stocks
Changes in abundance of fish stocks are continuing in 1975; however,
many changes are now deliberate and controlled. Uncontrollable changes in
native species (usually decreasing in numbers) and in introduced species
(usually increasing in numbers) frequently occurred during the past 25
years.
211
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1950's
Lake trout production reached zero in Lake Michigan and began to
decline in Lake Superior in 1950. Walleye production started to decline in
1950 in Lake Michigan. Cisco (chub) production dropped below 100,000
pounds in Lake Ontario, the first of the lakes to lose its chub population.
In 1952 production of lake herring in Lake Michigan and blue pike in
Lake Ontario began their "terminal" decline. Lake whitefish, blue pike,
and walleye production began declining in Lake Erie by 1956. By 1959 pro-
duction had fallen below 100,UOO pounds for lake herring in Lake Ontario,
Huron, and Erie; sauger in Lake Erie; and blue pike in Lakes Ontario and
Erie. Within a few years the blue pike had become virtually extinct in
Lakes Ontario and Erie. The emerald shiner, once exceedingly abundant,
became extremely scarce in Lakes Michigan and Huron. -
Increases also occurred in the 1950's. Smelt production exceeded
200,000 pounds in Lakes Ontario and Huron; 800,000 pounds in Lake Superior;
6 million pounds in Lake Erie; and 9 million pounds in Lake Michigan. The
rainbow smelt had become a significant species in the commercial catch of
all the lakes in less than 30 years after its introduction in Lake Michigan.
The first alewife was reported in Lake Michigan in 1949, and the
species was first reported in the commercial catch in Lake Michigan in 1956.
By 19b7 production exceeded 100,000 pounds, and by 1958 (9 years after it
was first reported) over 1 million pounds of alewives were produced in Lake
Michigan. Similar rapid colonization of the white perch occurred in Lake
Ontario; only 5 years elapsed between the first record of its presence
(1950) and the first report of it in the commercial catch (1955).
1960's
During this decade species of the whitefish family continued to decline
in the Great Lakes. Lake whitefish production fell below 100,000 pounds in
Lakes Ontario and Erie; lake herring began declining in Lake Superior and
fell below 100,000 pounds in Lake Michigan; deepwater Cisco (chub) produc-
tion began declining in Lakes Huron and Superior. Shallow water species
also declined: walleyes in Lakes Michigan and Ontario; yellow perch in
Lake Michigan; and northern pike in Lake Huron. The fourhorn sculpin, once
abundant in Lake Ontario, was extremely rare in the 1960's. A major decline
of the fourhorn sculpin during this period was also noted in Lake Michigan
(Wells and McLain, 1972).
Rehabilitation of the Fish Stocks
Biologists recognized that rehabilitation of fish stocks, principally
lake trout, could not begin until the sea lamprey was brought under con-
trol. A special agency, the Great Lakes Fishery Commission, was created in
19b6 by a treaty between Canada and the United States to fund and coordi-
nate existing efforts to control the sea lamprey. Initial control methods
attempted to block spawning migrations into streams by means of mechanical
and electrical barriers. This method proved ineffective. In 1957, after
212
-------
several years of research and testing thousands of chemicals, one was found
which was toxic to the lamprey but not lethal to other fish. Treatment of
lamprey-infested streams with the chemical 3-trifluoromethyl-4-nitropheno1
(TFM) began in 1958. By 1962, 2 years after all known lamprey nursery
streams had been treated in Lake Superior, success was verified when the
number of adult lampreys at assessment barriers was reduced nearly 85%
(Baldwin, 1964). The incidence of lamprey wounds on lake trout dropped
sharply, and survival of lake trout increased dramatically in Lake Superior.
A method of control had been found just in time to save the last natural
population of lake trout in the Great Lakes. The first complete treatment
of all Lake Michigan lamprey-infested streams was completed in 1963, Lake
Huron in 1970, and Lake Ontario in 1972. Chemical treatment of streams at
intervals of 2-4 years must continue, however, if the rehabilitation of
lake trout and other species is to become permanent.
The introduction of Pacific salmon in the Great Lakes had been
attempted many times, but had produced limited results until the successful
introduction of the coho salmon (Oncorhynchus kisutch) in Lake Michigan in
1966. By 1969 coho and Chinook salmon (Oncorhynchus tshawytscha) had been
introduced into all the upper lakes. The purpose of stocking coho and
chinook salmon in the Great Lakes was to increase the sport fishing
potential and not to establish self-sustaining populations. "Successful"
introduction, therefore, relates to rapid growth and high survival rates.
The pink salmon (Oncorhynchus gorbuscha), however, was an "unplanned" plant
in Lake Superior, where it succeeded in establishing spawning runs in 1959
and by 1975 had become established in Lakes Huron and Michigan.
Supplemental plantings of lake trout, following lamprey control, have
been made since 1958 in Lake Superior. The stocks have been built up to
near pre-lamprey levels. Reproduction of hatchery-reared fish has been dis-
appointing, however. Only in the last 2 or 3 years has the outlook im-
proved, when increasing numbers of young native trout have been reported.
The re introduction of lake trout in Lake Michigan, beginning in 1965,
has proved extremely successful in terms of survival and growth. No evi-
dence of reproduction, however, has been reported. Lake trout are now
being stocked in Lakes Huron and Ontario. Biologists continue to be
optimistic about the reestablishment of self-sustaining populations of lake
trout in all the Great Lakes, except Erie.
Salmonids other than lake trout and Pacific salmon have been stocked in
the Great Lakes since the lamprey has been controlled. Steel head trout
(rainbow trout), brown trout, brook trout, and Atlantic salmon are now
stocked in the lakes. Over 20 million salmonids annually are stocked in
the Great Lakes. In 1974 the first experimental plant of hatchery-reared
saugers was stocked in Lake Erie.
Environmental Changes, 1950-1975
Scientific investigations of environmental conditions of the Great
Lakes have increased exponentially during the past 25 years. Changes in
fish populations, benthic organisms, plankton, and water quality are now
213
-------
measured with greatly improved accuracy and frequency. The ability to con-
trol environmental conditions and to understand ecological interactions,
however, remains a goal of the future.
Chemical^changes--The increase in major ions continued in all the lakes
except Superior during the last 25 years. The rates of increase in all the
ions except calcium remain high for Lakes Ontario and Erie (Table 9).
Population increases also were substantial in the basins of Lakes Ontario
and Erie (Table 3) and probably account for the chemical changes.
TABLE 9. ESTIMATED AVERAGE CONCENTRATIONS OF DISSOLVED CHEMICAL
CONSTITUENTS IN THE GREAT LAKES IN 1970 (EXPRESSED IN
MG/LITER)9 WITH PERCENTAGE CHANGE SINCE 1950 IN PAREN-
THESES
Lake
Superior
Michigan
Huron
Erie
Ontario
Total
dissolved
solids
55 (-2)
155 (3)
115 (3)
206 (21)
210 (22)
Calcium
13 (0)
34 (0)
27 (12)
38 (0)
40 (5)
Sulphate
4 (0)
20 (43)
17 (31)
27 (17)
30 (20)
Chloride
2 (0)
7 (40)
7 (17)
21 (42)
29 (53)
Sodium
and
potassium
2 (-33)
5 (0)
4 (0)
14 (56)
15 (50)
aFrom Weiler and Chawla, 1969.
Critically low dissolved oxygen (DO) concentrations had not been re-
ported in the open waters of the Great Lakes until 1953. In that year
Britt (1955) measured DO concentrations less than 1 mg/liter in the western
basin of Lake Erie. Although the low DO levels lasted only a few days, it
caused a substantial mortality in the burrowing mayfly (Hexagenia) popula-
tion. In some areas of the western basin the entire population was killed,
where more than 1,000 mayfly nymphs per square meter had previously been
found (Britt, 1955). The first extensive zone of low DO (less than 1 ppm)
was measured in 1959, in the western portion of the central basin of Lake
Erie. An interagency synoptic survey of this basin in 1959 found an area
of approximately 1,400 square miles which contained less than 1 ppm of DO
in the hypolimnion. These conditions of low DO undoubtedly had occurred be-
fore 1959 (Carr, 1962).
Dissolved oxygen levels of less than 1 mg/liter occur annually in the
bottom waters of the central basin of Lake Erie and by 1974 covered several
thousand square miles. Oxygen depletion also has been reported in southern
214
-------
Green Bay (Lake Michigan) and the Bay of Quinte (Lake Ontario). Low DO
levels in the open waters of the other lakes have not been reported. The
virtual extinction of the sauger and blue pike and the decline of the
walleye population in Lake Erie are thought to be partially caused by the
low DO conditions (Smith, 1974).
Toxic Substances in Fish
Chemical contaminants in Great Lakes fish have been measured with in-
creasing frequency in the past decade (1965-75). Measurements were first
made in 1965 of the residues of the insecticides DDT and dieldrin in Great
Lakes fish. All 28 species for which DDT and dieldrin analysis has been
made contained measurable levels. Several species (chubs, lake trout, lake
herring) from Lake Michigan exceeded the U.S. Food and Drug Administra-
tion's (FDA) tolerance level of 5 yg/g in fish used for human consumption
(Reinert, 1970). Since the use of DDT was banned in 1972, the level in
Lake Michigan fish has decreased rapidly, from an average of 10 yg/g in
bloater chubs before 1972 to less than 3 yg/g in 1974.
During the same period in which DDT levels were decreasing in Great
Lakes fish, polychlorinated biphenol (PCB) levels were increasing. Again,
the species containing the highest levels were lake trout and bloater chubs
in Lake Michigan. The average concentration of PCB in Lake Michigan lake
trout above 24 inches exceeded 20 yg/g in 1974. Concentrations above the
FDA's 5 yg/g tolerance level have been reported in fish from Lake Ontario
and Lake Huron, as well as Lake Michigan.
In 1969 mercury levels in excess of the FDA's tolerance level of 0.5
yg/g were discovered in several species of fish (including walleye and
white bass) from Lakes St. Clair and Erie. Mercury levels above 0.5 yg/g
were also reported from Lakes Superior and Ontario. Two years following
curtailment of the source of mercury pollution to Lake St. Clair, the
levels in fish began to decrease. In two instances (DDT and mercury) stop-
ping the sources of chemical contaminants resulted in the rapid decline of
the toxicants in the environment. This success should give support to con-
tinued efforts to solve problems by eliminating the direct cause.
The contamination of Great Lakes fish with levels of DDT, PCB, and mer-
cury exceeding the FDA tolerance level has resulted in great financial hard-
ship to the commercial fishing industry. Direct or even indirect adverse
effects on the fish populations of the Great Lakes have not been detected.
Apparently, DDT and PCB in the low nannogram-per-liter levels in the open
lake waters have, through biomagnification, reached the microgram-per-gram
level in fish tissue.
Changes in Benthos and Plankton
Changes in bottom-dwelling organisms have been documented by several
investigators within the past 20 years: Britt (1955) and Carr and Hiltunen
(1965) for Lake Erie; Schneider, Hooper, and Beeton (1969) for Saginaw Bay;
Hiltunen (1967) for Lake Michigan. In all of these studies the changes
have been from the more "pollution-intolerant" organisms (mayflies,
215
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caddisflies, amphipods) to "pollution-tolerant" forms (primarily oligo-
chaetes and midge larvae) and from greater to lesser species diversity.
The geographic areas affected by loss of intolerant organisms are being
extended further into the lakes from the pollution sources (primarily river
mouths). The effects of these changes on fish populations will remain
speculative until the interactions can be quantitatively assessed.
Phytoplankton and zooplankton populations have also changed markedly in
many areas of the Great Lakes (Beeton, 1969). The changes in phytoplankton
have been from dominance by multispecies diatom communities to species of
green and blue-green algae more tolerant of eutrophic conditions. Zoo-
plankton communities have reacted similarly (Beeton, 1969), resulting in a
loss of species diversity and increases in species associated with
eutrophic environments. Again, the relation of these changes in fish
populations is incompletely understood.
CONCLUSIONS
It is obvious that the environment in many areas of the Great Lakes has
deteriorated. Assigning direct cause and effect to changes in specific
populations or species of Great Lakes fish is difficult and controversial.
Heavy exploitation of many stocks is undoubtedly a factor in the decline of
many species. Changes in water chemistry, plankton, bottom fauna, and un-
exploited fish species, however, clearly show that factors other than fish-
ing have drastically changed the characteristics of the Great Lakes.
Now that we know our capabilities, how can we avoid past mistakes and
stop, or perhaps even reverse, the trend toward environmental chaos? One
possibility is to understand the forces that operated in the past to pro-
duce present conditions. Scientists and administrators with responsibility
for protecting the aquatic environment can learn much from the perturba-
tions foisted on the Great Lakes. For example, early recognition of the
effects of unmanaged commercial fishing could have prevented, or at least
delayed, the decimation of many fish populations. Wise management of the
uses of tributary streams would have saved many stocks of anadromous
species. It is difficult, however, to blame these errors of omission on
our predecessors, for they did not have the advantage of hindsight to im-
prove their foresight. Our generation has no such excuse. Opportunities
missed in the past to protect the aquatic communities are gone, but oppor-
tunities remain to save and rehabilitate our aquatic environment.
Recognition of environmental degradation in the Great Lakes has led
Canada and the United States to a firm commitment to halt and reverse this
trend. Evidence of success in this endeavor is already apparent. Rehabil-
itation of many tributaries has permitted the establishment of spawning
runs by anadromous species. Levels of DDT in fish tissue have decreased as
much as 80% after the use of the insecticide was banned. More comprehen-
sive and better treatment of municipal and industrial waste has resulted in
noticeable improvements in the quality of receiving waters.
216
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Let us hope that by the year 2000 the history of changes in fish
species of the Great Lakes will show only increases in native species bet-
ween 1975 and the new century.
REFERENCES
Aron, W.I., and S.H. Smith. 1971. Ship canals and aquatic ecosystems.
Science 174: 13-20.
Baldwin, N.S., and R.W. Saalfeld. 1962 (plus 1970 supplement). Commercial
fish production in the Great Lakes, 1867-1960 (supplement 1961-68).
Great Lakes Fish. Comm., Tech. Rep. 3. 166 p.
Baldwin, M.S. 1964. Sea lamprey in the Great Lakes. Can. Audubon Mag.,
November-December, Chap. 19: 535-558.
Beeton, A.M. 1969. Changes in the environment and biota of the Great
Lakes, p. 150-187. lr\_ Eutrophication: causes, consequences, correc-
tives. National Academy of Sciences, Washington, D.C.
Beeton, A.M., and D.C. Chandler. 1963. The St. Lawrence Great Lakes, p.
. Jji Limnology in North America. Univ. Wisconsin Press,
Madison, Wis.
Beeton, A.M., and W.T. Edmondson. 1972. The eutrophication problem. 0.
Fish. Res. Board Can. 29: 673-682.
Britt, N. Wilson. 1955. Stratification in western Lake Erie in summer of
1953; effects on the Hexagenia (Ephemeroptera) population. Ecology
35: 239-244.
Carr, John F. 1962. Dissolved oxygen in Lake Erie, past and present.
Proc. 5th Conf. Great Lakes Res., Univ. Michigan, Great Lakes Res.
Div., Publ. 9. p. 1-14.
Carr, John F., and Carl K. Hiltunen. 1965. Changes in the bottom fauna
of western Lake Erie from 1930 to 1961. Limnol. Oceanogr. 10: 551-569.
Christie, W.J. 1972. Lake Ontario: Effects of exploitation, introduc-
tions, and eutrophication on the salmonid community. J. Fish. Res.
Board Can. 29: 913-929.
1974. Changes in the fish species composition of the Great
Lakes. J. Fish. Res. Board Can. 31: 827-854.
Fish, C.J., and Associates. 1960. Limnological survey of eastern and cen-
tral Lake Erie, 1928-1929. U.S. Fish Wildl. Serv., Spec. Sci. Rep.
Fish. 334. 198 p.
217
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Hartman, VI.L. 1972. Lake Erie: Effects of exploitation, environmental
changes, and new species on the fishery resources. J. Fish. Res. Board
Can. 29: 899-912.
Hiltunen, Jarl K. 1967. Some Oligochaetes from Lake Michigan. Trans. Am.
Microsc. Soc. 86: 433-454.
Hutchinson, G.E. 1957. A treatise on limnology, Volume I: Geography,
physics, and chemistry. John Wiley and Sons, New York. 1015 p.
Lawrie, A.M., and J.F. Rahrer. 1972. Lake Superior: Effects of exploita-
tion and introductions on the salmonid community. J. Fish. Res. Board
Can. 29: 765-776.
Reinert, Robert E. 1970. Pesticide concentrations ir> Great Lakes fish.
Pest. Mon. J. 3: 233-240.
Schneider, O.C., F.F.. Hooper, and A.M. Beeton. 1969. The distribution and
abundance of benthic fauna in Saginaw Bay, Lake Huron. Proc. Conf.
Great Lakes Res. 12: 1-20.
Scott, W.B., and E.J. Grossman. 1973. Freshwater fishes of Canada. Fish.
Res. Board Can., Bull. 184. 966 p.
Smith, S.H. 1957. Evolution and distribution of the Coregonids. 0. Fish.
Res. Board Can. 14: 599-604.
1964. Status of the deepwater cisco population of Lake Michi-
gan. Trans. Am. Fish. Soc. 93: 155-163.
. 1968. Species succession and fishery exploitation in the Great
Lakes. J. Fish. Res. Board Can. 25: 667-693.
. 1972. Factors of ecological succession in oligotrophic fish
communities of the Laurentian Great Lakes. J. Fish. Res. Board Can.
29: 717-730.
. 1974. Responses of fish communities to early ecologic changes
in the Laurentian Great Lakes, and their relation to the invasion and
establishment of the alewife and sea lamprey. MS in preparation.
Weiler, R.R., and V.K. Chawla. 1969. Dissolved mineral quality of Great
Lakes waters. Proc. Conf. Great Lakes Res. 12: 801-818.
Wells, L., and A. McLain. 1972. Lake Michigan: Effects of exploitation,
introductions, and eutrophication on the salmonid community. J. Fish.
Res. Board Can. 29: 889-898.
Wright, S. 1955. Limnological survey of western Lake Erie. U.S. Fish
Wildl. Serv., Spec. Sci. Rep., Fish. 139. 341 p.
218
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SECTION 19
SEA LAMPREY (PETROMYZON MARINUS LINNAEUS) IN THE
SAINT LAWRENCE GREAT LAKES OF NORTH AMERICA:
EFFECTS, CONTROL, RESULTS
Carlos M. Fetterolf, Jr.
The sea lamprey (Petromyzon marinus Linnaeus 1758) is in the class
Agnatha, subclass Cyclostomata (Marsipobranchii), order Petromyzontiforms
(Hyperoartia), and family Petromyzontidae. It is anadromous over most of
its range (Figure 1), spending its parasitic adult life in the sea, but is
landlocked in the Great Lakes and a few lakes in New York State.
Figure 1. Range of the sea lamprey. Modified from Leim
and Scott (1966).
Sea lampreys ascend freshwater tributaries and prefer a stony, gravelly
bottom for spawning (Figure 2). Adults excavate depressions about 15 cm
deep and 0.6-1 m in diameter in which the landlocked Great Lakes females
deposit about 50,000 - 70,000 eggs each. Adults die after spawning. Eggs
hatch into larvae (ammocetes), blind and toothless with a flexible, fleshy
hood overhanging the mouth. Ammocetes live 3-14 years as filter feeders in
burrows constructed in soft sediments of tributaries. Transformation
(metamorphosis) involves disappearance of the hood and development of teeth
on the tongue and a buccal funnel with teeth radiating in all directions
from the mouth (Figure 3). Spawning Great Lakes sea lampreys are 30-60 cm
long. Ammocetes reach about 12-16 cm in length before transformation.
219
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-12-20 MONTHS
Figure 2. Life cycle of the sea lamprey (from Crowe, 1975)
220
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"Transformers" (metamorphosed larvae) move downstream to lake or sea where
they feed by attaching themselves to fishes, rasping a hole through the
body covering and sucking body juices. Depending in part on hunger and
size of the adult lampreys and size of their prey, one feeding can be fatal
or the prey can withstand several attacks.
Figure 3. The mouth of the sea lamprey, which is lined with horny
teeth surrounding a rasping tongue in the center.
Passage of the sea lamprey to the upper Great Lakes was blocked by
Niagara Falls (about 50 m high) between Lakes Ontario and Erie. Completion
of the We I land Canal for shipping in 1829 enabled the sea lamprey to bypass
Niagara Falls. Moving up the Great Lakes, the lamprey was recorded in Lake
Erie in 1921. The mid-1930's the animal had reached Lakes Huron and
Michigan and by the 1940's it was firmly established in Lake Superior. By
the mid-1940's fish stocks in Lakes Huron and Michigan had been severely
damaged, and similar damage was predicted correctly for Lake Superior. The
catastrophic decline in lake trout (Salvelinus namaycush) in relation to
sea lamprey invasion is best traced in Lake Superior (Figure 4).
221
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0)
z
o
/\.
IOO
80
SO
40
20
630
1935
I94O
1945
1950 1955 I960 1965 1970
Figure 4. Production metric tons) of lake trout, 1930-66
(broken line), and number of sea lampreys caught
in index streams in Lake Superior, 1953-69 (solid
line). From Smith, 1971.
Fisheries agencies recognized the urgent need for a control program,
and because the Great Lakes are bordered by two countries, United States
and Canada, including one Canadian province and eight U.S. states, the need
for international and interstate cooperation was imperative (Figure 5). In
1948 a committee representing these jurisdictions was established. Initial
control efforts were experimental and uncoordinated because so many
agencies were involved, and funding was not assured. In 1955 a Convention
on Great Lakes Fisheries was ratified by the United States and Canada which
established the Great Lakes Fishery Commission. The commission's program
is divided into two major segments: (1) sea lamprey control and (2) coordi-
nation of fisheries research and management. The commission has no regula-
tory authority, but provides the forum in which mutually beneficial courses
of action are developed. Funding is through the Department of External
Affairs in Canada and by legislative appropriation to the Department of
State in the United States. The early programs of the study of sea lamprey
life history and distribution, development and testing of barriers in
streams, and screening of chemicals that would selectively destroy larvae
were continued, coordinated, refined, and expanded under commission
auspices after 1955. The control programs are carried out by agents of the
commission, the U.S. Fish and Wildlife Service, and the Canadian Department
of the Environment. Research is funded by, and has been done mostly by,
the U.S. Fish and Wildlife Service.
Sea lampreys are most vulnerable to current control methods when
concentrated in streams as adults on upstream migrations, as larvae in the
streams, or as transformers moving downstream. While tests were carried on
to find a selective chemical, a control program by means of mechanical and
222
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electromechanical barriers was operated. At its peak in 1959 the program
included about 135 barriers in the United States and Canada, but 401 of the
5,747 tributaries to the Great Lakes are known to produce sea lamprey. The
effectiveness of barriers as a control method was never adequately
determined, but there is no doubt that barriers are effective in killing
large numbers of adult sea lampreys. The barrier program was discontinued
as the major control method after the discovery of a selective chemical.
Electrical barriers are still used at selected sites as a means of
assessing program effectiveness and to provide control over experimental
areas.
,2»-
PORT
ARTHUR
ILLINOIS
CHICAGO
Figure 5. The Great Lakes. Numbers igdicate streams on which sea
lamprey counting weirs were installed. From Smith, 1971.
Some 6,000 chemicals were screened through laboratory bioassay over a
7-year period (Applegate et al., 1957). Promising toxicants were field
tested in 1957 and 1958. These successful tests led the commission in 1958
to adopt use of two chemicals, 3-trifluoromethyl-4-nitrophenol (TFM) and
2', 5-dichloro-4'nitrosalicylanilide (Bayer 73), as the major sea lamprey
223
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control technique. Routine stream treatments are carried out with TFM or
with TFM plus a small amount (1-4%) of powdered Bayer 73. The addition of
Bayer 73 reduces up to one-half the amount of TFM required and greatly
reduces the cost of treatments. A granular form of Bayer 73, which settles
to the bottom before chemical release, is also used in difficult-to-reach
areas during treatment with TFM, but is more frequently used as a collect-
ing tool in surveys. At proper concentrations the chemicals destroy sea
lamprey larvae without significantly affecting other fauna and flora.
A defined range of concentrations, dependent on alkalinity, pH, and
temperature, must be maintained for several hours throughout the treatment
area. Field bioassays, conducted in mobile units, identify the lowest con-
centration of TFM that kills 100% of sea lamprey larvae in 9 hr or less and
the highest concentration that does not kill more than 25% of the test
species (usually rainbow trout, Salmo gairdnerii) in 18-24 hr (Kanayama,
1963). These criteria provide safety factors at both extremes. Stream con-
centrations are maintained between these limits by controlled applications
at several stations on the stream. The defined range can vary between 1.0
and 2.3 mg/liter toxicant on low alkalinity streams (10-20 mg/liter as
CaC03) and between 6.7 and 17.0 mg/liter toxicant on high alkalinity
streams (163 mg/liter as CaC03).
Sea lamprey control with lampricides was initiated in Lake Superior in
1958 and expanded to Lake Michigan in 1960, to Lake Huron in 1966, and Lake
Ontario in 1971. The first "round"1 of treatments was completed in Lake
Superior in 1961; in Lake Michigan in 1966; in Lake Huron in 1970; and in
Lake Ontario in 1972. The total number of treatments since 1958 exceeds
1,000.
In Lake Superior, where the control program has been in operation for
the greatest number of years and where its effectiveness has been most care-
fully evaluated, sea lamprey abundance has been reduced by about 90%. A
quantitative measure of sea lamprey abundance has been obtained from counts
of mature (spawning run) sea lampreys reaching assessment barriers.
Numbers of mature sea lampreys in the Lake Superior spawning runs declined
sharply in 1962, the year after the first round of stream treatments had
been completed (Figure 6). The decrease was accompanied by a marked de-
cline in the incidence of fresh sea lamprey wounds on lake trout and later
by an improved survival of lake trout to older age and larger size. Equiva-
lent quantitative data are not available for Lakes Huron and Michigan, but
the responses of sea lamprey and fish populations to control efforts have
been similar to those in Lake Superior.
The Great Lakes Fishery Commission is concerned that the control pro-
gram is singularly dependent on chemicals, primarily TFM. Only one chemical
manufacturer submits bids. Costs have risen sharply to $13.18 kg, and we
use over 45,360 kg a year. Early in the program it was necessary to treat
each stream only once every 4 years. However, the average ammocete is now
A "round" denotes that all known sea-lamprey-producing streams tributary
to that lake have received one chemical treatment.
224
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transforming in a shorter period, presumably because of reduced competition
for food and space. This requires more thorough surveys and more frequent
treatments and emphasizes the need for alternative controls. The commis-
sion is developing an integrated control program including permanent
barriers on selected streams and is sponsoring research into chemical
attractants and repellants as well as chemosterilants.
ALL MAJOR LAKE SUPERIOR
STREAMS TREATED
1958 1959 1960 1961 1962 1963 1964 1965 1966 1967 1968 1969 1970 1971 1972 1973 1974
Figure 6. Sea lamprey catch from eight streams tributary to
Lake Superior. Modified from Crowe, 1975.
In 1971 comprehensive studies of the immediate and long-term effects of
lampricides in the environment were initiated. Results suggest that the
effects are very small, that the chemical control program can continue, and
that registration of the lampricides by the Environmental Protection Agency
will be forthcoming upon completion of the required studies. About $1.2
million were allocated to do this registration-oriented research in 1971-74.
The annual budget of the commission is about $4 million. Without con-
trol of sea lamprey the sport and commercial fisheries would be limited.
To date, with fish stocks only in the process of rehabilitation, the value
of the Great Lakes sport fishery is estimated at over $350 million. The
commercial fishery is valued at $19 million at the dock and approaches $100
million at the market. There is an excellent return from the money
invested in sea lamprey control.
225
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ACKNOWLEDGEMENTS
I have been associated with the Great Lakes Fishery Commission since
July 1975. As a recent executive secretary I take no credit for the pro-
gress and programs reported above. The credit is deserved by the pioneer
sea lamprey control workers in the U.S. Fish and Wildlife Service, the
Canadian Department of the Environment, the Great Lakes Fishery Commission,
all cooperators, and the administrations that supported them.
REFERENCES
Applegate, V.C., J.H. Howe11, A.E. Hall, Jr., and M.A. Smith. 1957. Toxi-
city of 4,346 chemicals to larval lampreys and fishes. U.S. Fish, and
Wild!. Serv., Spec. Sci. Rep. Fish. 207. 157 p.
Crowe, W.R. 1975. Great Lakes fishery commission: history, program, and
progress. Great Lakes Fish. Comrn., Ann Arbor, Mich. 23 p.
Kanayama, R.K. 1963. The use of alkalinity and conductivity measurements
to estimate concentrations of 3-trifluoromethyl-4-nitrophenol required
for treating lamprey streams. Great Lakes Fish. Comm. Tech. Rep. Ser.
7,
Ann Arbor, Mich. 10 p.
Leim, A.H., and W.B. Scott. 1966. Fishes of the Atlantic coast of Canada.
Fish. Res. Board Can., Bull. 155. 485 p.
Smith, B.R. 1971. Sea lampreys in the Great Lakes of North America, p.
Ir± M.W. Hardisty and I.C. Potter (ed.) The biology of lampreys, Vol. 1.
Academic Press, New York.
226
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VOLUME II
PROCEEDINGS OF THE SECOND USA-USSR
SYMPOSIA ON THE
EFFECTS OF POLLUTANTS UPON AQUATIC ECOSYSTEMS
June 22-26, 1976
Borok, Jaroslavl Oblast
USSR
Edited by
Wayland R. Swain
and
Nina K. Ivanikiw
ENVIRONMENTAL RESEARCH LABORATORY-DULUTH
OFFICE OF RESEARCH AND DEVELOPMENT
U.S. ENVIRONMENTAL PROTECTION AGENCY
DULUTH, MINNESOTA 55804
-------
DISCLAIMER
This report has been reviewed by the Environmental Research Labora-
tory-Duluth, U.S. Environmental Protection Agency, and approved for publi-
cation. Mention of trade names or commercial products does not constitute
endorsement or recommendation for use.
n
-------
FOREWORD TO VOLUME II
These Proceedings result from the second symposium held by Project
02.02-1.3 of the US-USSR Joint Agreement in the Field of Environmental
Protection, established in May, 1972.
Both broad review and narrowly specific papers were presented by
participants from both countries in an effort to continue the joint pro-
cedural, technological and methodological exchange and familiarization
begun at the Duluth Symposium in October of 1975. Learning does not occur
and subsequent understanding and application must be based on a
foundation of fact. The atmosphere of mutual interest, candor and respect
which surrounded this symposium enabled another series of steps in the
learning process. Perhaps the philosophy underlying this symposium, and
the project itself is best expressed by an old saying, which trans-
literated from the Russian approximates: Vyek zhee-vee, Vyek oo-chee,
Live a lifetime, learn a lifetime.
-------
PREFACE TO VOLUME II
This volume contains fifteen papers presented at the Second US-USSR
Symposium on the Effects of Pollutants on Aquatic Ecosystems. All of the
papers were presented in English or Russian with simultaneous
translations into the corresponding language at Borok, Jaroslavl Oblast,
USSR during June 22-26, 1976 at the Institute for the Biology of Inland
Waters of the USSR Academy of Sciences.
Professor N.V. Butorin, Director of the Institute and Project Leader
for the Soviet side* served as official host for the American delegation
and has assumed the responsibility for the publication of these pro-
ceedings in the Russian language. This joint bilingual publication re-
presents a reaffirmation of the continuing commitment pledged by both
countries to cooperative environmental activities.
IV
-------
INTRODUCTION
The Joint US-USSR Agreement on Cooperation in the Field of Environ-
mental Protection was established in May of 1972. These proceedings re-
sult from one of the projects, Project 0.2.02-1.3, Effects of Pollutants
Upon Aquatic Ecosystems and Permissible Levels of Pollution.
As knowledge related to fate and transport of pollutants has grown,
it has become increasingly apparent that local and even national
approaches to solving pollution problems are insufficient. Not only are
the problems themselves frequently international, but an understanding of
attentive methodological approaches to the problem can avoid needless
duplication of efforts. This expansion of interest from local and
national represents a logical and natural maturation from the provincial
to a global concern for the environment.
In general, mankind is faced with very similar environmental problems
regardless of the national or political boundaries which we have erected.
While the problems may vary slightly in type or degree, the fundamental
and underlying factors are remarkably similar. It is not surprising,
therefore, that the interests and concerns of environmental scientists
the world over are also quite similar. In this larger sense, we are our
brother's brother, and have the ability to understand our fellowman and
his dilemma, if we but take the trouble to do so. It is this singular
idea of concerned scientists exchanging views with colleagues that pro-
vides the basic strength for this project. While our methods may vary,
our goals are identical, and therein lies the value of such a coopera-
tive effort.
Wayland R. Swain, Ph.D.
Project Officer, U.S. Side
-------
FIGURES
Section
Page
5 Number of fry produced per 100 grams of female fathead
minnow in HCN 42
5 Number of fertilized eggs produced per female brook trout
in HCN 48
5 Growth of brook trout in various concentrations of HCN . . 50
6 The scheme of microbiological indices of water quality . . 56
6 Scheme of the order of test tubes in determination of
living bacteria by titer method 58
6 Position of the test-tube (N 2) with ethyl alcohol in
the row when determining the spores of bacteria in mud
deposits 61
6 Effect of Ag ions on bacteria. Analysis using the hetero-
trophic assimilation of C02 70
6 Determination of the reserves of phenol in the water of
the Kanskoe Reservoir by the method of Right and Hobbit . . 71
9 Hydroxylative enzyme systems in liver and bone that
require vitamin C 98
12 Changes in the conditioned reflex activity in the common
guppy (Leb-atei Ae£ccu£o£o6) under the influence of sub-
lethal concentrations of phenol 124
12 Symptoms of intoxication of medicine leech in solutions
of polychlorpinene (1-4), chlorophos (5-8), and phenol
(9-12) 126
14 Correlation between Woodiwiss' biotic index and BOD5 ... 145
14 Correlation between Woodiwiss1 biotic index and bi-
chromate oxygen consumption 146
-------
Section Page
15 Distribution of seston in the River Svisloch in August
of 1973 152
15 Relative chlorophyll content in seston of the River
Svisloch in June of 1973 153
15 Specific Oxygen Consumption (SOC) by seston in three
representative lakes 154
15 Relationship of photosynthesis ($) and distraction (D)
in the Neman River in August of 1975 155
15 The ration of $/D and $/R on the Pripyat River 156
VII
-------
TABLES
Section Page
3 Acute toxicity of six candidate forest insecticides to
brook trout in water of different temperature, hardness,
and pH 15
3 Acute toxicity (96-h LCso and 95% confidence intervals,
mg/1) of six candidate forest insecticides to mature scuds
and stonefly naiads tested at 18 °C and 16 °C 20
3 Effect of PCB (Aroclor 1254) residues on sensitivity of
brook trout to candidate forest insecticides 23
3 Effect of DDT residues on sensitivity of atlantic salmon
to a mixture of guthion and dylox 24
3 Effect of PCB (Aroclor 1254) residues (total body) on
sensitivity of brook trout to four forest insecticides . . 25
5 Analysis of well water used in bioassays 40
5 Survival and weight of first-generation fathead minnow
after 28 days, 56 days, and 84 days of exposure to
hydrogen cyanide 43
5 Egg production, egg survival and terminal weights of
first-generation adult fathead minnows exposed to
various concentrations of cyanide 44
5 Survival, length and weight of fathead minnow after 28
days and 56 days of exposure to hydrogen cyanide 46
5 Egg production of adult brook trout exposed to HCN 144
days before start of spawning 47
5 Percentage survival of brook trout eggs and alevins ex-
posed to various levels of HCN 49
5 Growth of brook trout alevins from hatch to 90 days ex-
posed to various levels of HCN 51
viii
-------
Section Page
5 Ninety-six hour LC$Q and threshold concentrations of HCN
to fathead minnows and brook trout juveniles (yg/liter) . . 53
6 Contents of bacteria in waters of varying trophic degree . 57
6 Ratio of the number of saprophytic bacteria to their
total number as an index of the cleanliness of water ... 60
6 Oxygen consumption for respiration of organisms in
water-bodes of various types (summer values) 66
6 Mean mid-summer values of heterotrophic assimilation
of C02 in water bodies of various types 67
7 Percent un-ionized ammonia in aqueous ammonia solutions . . 76
8 Changes in boundary concentrations of 3 pollutants
influencing biomass in Vapknia. magna 85
8 The influence of factors on the character of action of
the pollutants on aquatic organisms 88
8 Distribution of organisms according to their relative
sensitivity to toxic substances (toxobity) 89
9 Summary of experimental conditions and sampling periods
used during continuous exposure of brook trout, fathead
minnows, and channel catfish to toxaphene 94
9 Relation between backbone development and weight in fish
exposed to toxaphene 95
9 Statistical significance of the effects of toxaphene on
growth and hydroxyproline concentrations concentration
in fish 96
9 Mean concentrations of vitamin C (yg/g of wet tissue) in
liver and backbone and collagen (mg/g of dried bone) in the
backbone of channel catfish fed a diet low in vitamin C,
after 90- and 150-day exposures to different concentrations
of toxaphene 97
10 Results of the toxicity of a copper-ammonium solution on
test culture of &LcAoc.yt>&A heAu-g-inoAa (laboratory strain,
5-day experiment) 103
10 Primary production and destruction in phytoplankton samples
(blue-green + diatom) under the action of heavy metals.
Values indicated as percent of the control 105
IX
-------
Section
Page
10 Gross photosynthesis in a water-body treated with diurone
at the time of a blue-green algal bloom 106
10 Tests on DDT content in storage tissues of fish 108
10 Increase in sensitivity of biological tests on survival
of aquatic organisms at 30 °C 109
11 Analysis of well water values expressed in mg/1 115
11 96-hour and threshold LCso of hydrogen sulfide (mg/1)
for brook trout, bluegill, fathead minnow and goldfish . . 116
11 Chronic effect of sublethal concentrations of "hydrogen
sulfide (mg/1) on trout, bluegill, fathead minnow and
goldfish 117
11 Increment in weight and survival of bluegill started as
young-of-the-year and exposed to varied concentrations
of hydrogen sulfide for 826 days at 03 of 6.0 mg/liter
and mean temperature of 17.8-18.5 °C varied seasonally . . 118
11 Weight and reproduction of adult bluegills exposed to
varied concentrations of hydrogen sulfide for 97 days
at 23.5-23.9 °C 118
11 Growth of fathead minnow in 112-day exposure to hydrogen
sulfide at 23 °C, pH 7.7, 02 6.4-7.3 mg/liter - expressed
as mean weight in grams of three replications at the end
of succeeding 28-day periods 119
12 Symptoms of intoxication in carp exposed to the short-
term action of toxic substances 125
12 Toxicity of some substances for medicine leech and their
threshold concentrations (mg/1) producing avoidance
reaction 128
13 Classification of asbestos 134
15 Relative chlorophyll content of seston in clean and
polluted parts of rivers 151
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ACKNOWLEDGEMENTS
In any project of the scope and complexity of this effort, the Project
Officers become increasingly indebted to a large number of individuals
who contribute their time and effort with no thought of personal gain.
Unfortunately, the list of persons who materially aided the effort is too
extensive to allow a complete discussion. However, while those persons
who made outstanding contributions to the success of this project are
acknowledged below, the editors also wish to thank all those others, both
Soviet and American, whose efforts and assistance smoothed the way to a
satisfactory completion of this phase of the project.
Sincere thanks are extended to Ms. Elaine Fitzback and Mr. Igor Kozak
whose assistance with the translations have made possible the publication
of these proceedings. The substantial contributions and tireless efforts
of Ms. Virginia Shannon, Ms. Debbie Caudill, and Ms. Dawn Armatis to the
preparation of the proceedings are acknowledged with deep appreciation.
-------
CONTENTS
Foreword iii
Preface iv
Introduction v
Figures vi
Tables viii
Acknowledgements xi
1. Toxicity Tests in the Regulation of Waste Discharges in
the United States
Peter Doudoroff 1
2. Toxicological Control of Pollution of Freshwaters
N.S. Stroganov 8
3. Toxicity of Experimental Forest Insecticides to Fish and
Aquatic Invertebrates
Richard A. Schoettger and Wilbur L. Mauck 11
4. Principles and Methods of Biological Establishment of the
Norms of Chemical Substances and Evaluation of the Level
of Pollution in Water-Bodies
V.I. Lukyanenko 28
5. Chronic Effects of Low Levels of Hydrogen Cyanide on
Freshwater Fish
Lloyd L. Smith, Jr 39
6. Microbiological Indices of the Quality of Water and Methods
of Their Determination
V.I. Romanenko 55
7. Ammonia and Nitrite Toxicity to Fishes
Rosemarie C. Russo and Robert V. Thomann 75
8. A Research System for Developing Fisheries Standards for
Water Quality, Considering the Peculiarities of Transferring
Experimental Data to Natural Water Bodies
L.A. Lesnikov 83
9. Collagen and Hydroxyproline in Toxicological Studies
With Fishes
Foster L. Mayer and Paul M. Mehrle 92
Xll
-------
10. Experimental Testing of Toxicity of Water Media and
Increasing of the Sensitivity of Biological Tests
L.P. Braginski, V.D. Bersa, T.I. Biger, I.L. Burtnaya,
F.Ya. Komarovski, A.Ya. Malyarevskaya, E.P. Shcherban . . 102
11. Chronic Effects of Low Levels of Hydrogen Sulfide on
Freshwater Fish
Lloyd L. Smith, Jr 113
12. The Behavioral Aspects of Aquatic Toxicology
B.A. Flerov 122
13. Geologic Pollution Problems of Lake Superior
Albert B. Dickas, Ph.D 133
14. Experimental Application of Various Systems of Biological
Indication of Water Pollution
G.G. Winberg 141
15. Structural and Functional Characteristics of Seston as
Indices of Water Pollution
A.P. Ostapenya 150
xm
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SECTION 1
TOXICITY TESTS IN THE REGULATION OF WASTE DISCHARGES
IN THE UNITED STATES
Peter Doudoroff
I have recently undertaken a cursory examination of much of the Soviet
literature in the field of aquatic toxicology. I have found in that lit-
erature no evidence to indicate that biological tests for toxicity of
wastewaters are often required or routinely performed in regulating waste
discharges in the U.S.S.R. Neither the standardization of toxicity bio-
assay methods, nor the formulation or systemization of procedures for the
application of bioassay results in the control of waste disposal seems to
have received nearly as much attention in the U.S.S.R. as in the United
States. Some of the pertinent ideas and current practices of American
workers that are briefly and incompletely reviewed here, and my own
thoughts concerning their relative merits, may be interesting and useful,
therefore, to those Soviet scientists who must deal with regulatory pro-
blems. I know that many industrial effluents and their individual com-
ponents have been tested by Soviet investigators for acute and chronic
toxicity to fish and other aquatic organisms. But it is not that scienti-
fic research into the toxicity of these various water pollutants to which
I refer. I am speaking of regular biological testing of wastewaters by
technicians to verify compliance with, or to detect violations of, some
specific regulatory requirements limiting the discharge of toxic wastes.
Toxicology studies reported in the Soviet literature often have been
directed toward the determination of maximum acceptable concentrations of
paricular toxic substances to actual concentrations of which can be deter-
mined by chemical analysis is receiving waters. In the United States,
also, much research of this kind has been done and continues. Defi-
ciencies in the chemical criteria of water or wastewater quality so
developed, and a need for more reliance on biological tests of effluents
in pollution control have long been apparent to many American workers.
We realized that many of the toxic components of industrial wastes had
not been, and would not soon be identified, or could not be reliably mea-
sured for lack of suitable analytical methods. Further, the toxicities
of even identified and measurable compounds to important species of fish
(and the interactions of these toxicants with natural components of the
various receiving waters and among themselves in complex mixtures) were
mostly unknown and unpredictable. Largely for these reasons, my
colleagues and I long ago undertook the standardization of toxicity bio-
assay methods suitable for routine application in industrial laboratories,
1
-------
using locally important fish as test animals and waters receiving the
tested wastes as diluents (Hart, Doudoroff, and Greenbank, 1945;
Doudoroff, ^t at., 1951). The methods developed and recommended have
been widely adopted by other investigators, and by regulatory agencies
and industrial organizations in the United States and elsewhere. They
have appeared repeatedly, with some minor modifications or refinements,
in manuals of currently approved, standard practice, such as the eleventh
edition of "Standard Methods for the Examination of Water and Wastewater"
(American Public Health Association &t al., 1960) and the subsequent edi-
tions. Similar methods for evaluation of the toxicity of water pollutants
to organisms other than fish have been developed, but their use in the
United States outside of research laboratories has not yet been extensive.
Chemical criteria of water quality can be very useful, and cannot be
entirely ignored in controlling pollution for the protection of aquatic
life. However, deficiencies in this area, even after the great technolog-
ical advances of recent years, are still very real and apparent. Thus,
frequent reliance on toxicity bioassays of effluents regulating the dis-
charge of toxic wastes is still necessary. Several ways in which these
toxicity tests can be used in controlling water pollution have been pro-
posed by biologists and tried by regulatory agencies.
The maximum safe or harmless concentration of an industrial waste or
other toxicant in a receiving water cannot be directly determined, by per-
forming a toxicity test of short duration, e.g., a 96 hour test. Much
longer and more difficult tests can be successfully undertaken only in re-
search laboratories, and cannot be frequently repeated. When only the
acute toxicity of an effluent is known, its highest permissible concentra-
tion in the receiving water must be computed by some prescribed formula
that has been judged to be appropriate. For example, the concentration
found by experiment to be fatal in 48 hours to just 50 percent of test
animals, termed the 48-hour "median tolerance limit" or "median lethal
concentration" (LC50), may be simply multiplied by a fractional "appli-
cation factor", e.g., 0.10, to obtain the permissible or presumably safe
concentration. This formula, with the application factor of 0.10, was
first recommended tentatively by the Aquatic Life Advisory Committee of
the Ohio River Valley Water Sanitation Commission (1955) as one believed
to be sufficiently but not unreasonably restrictive. That committee
noted, however, that smaller or larger application factors may be often
fully justifiable. The recent trend has been to reduce the permissible
concentrations, substituting the 96-hour LC50 for the 48-hour value, and
using smaller application factors in the formula. More complicated
formulas have been proposed (Hart, Doudoroff, and Greenbank, 1945;
Doudoroff vt &£., 1951; footnote No. 7). For some time they attracted
much favorable attention, but regrettably they have not proved
sufficiently useful for wide-spread adoption by regulatory agencies.
For each of a large variety of toxic substances, an individual appli-
cation factor has been recently proposed (National Technical Advisory
Committee on Water Quality Criteria, 1968; Committee on Water Quality
Criteria, National Academy of Sciences and National Academy of
Engineering, 1972/1973). The recommended values are based on the results
2
-------
of laboratory studies in which median lethal concentrations of the toxi-
cants were related to the highest levels that were apparently harmless to
the test animals in experiments of long duration. The usefulness of
these widely ranging application factors in dealing with industrial wastes
that are complex mixtures of variable, and often incompletely known com-
position, is questionable. Application factors most appropriate to
different kinds of industrial wastewaters differ greatly. Prescribed ap-
plication factors also vary with the value of commercial or recreational
fisheries that are to be protected, some of which merit a high degree or
level of protection than others. Economic and social considerations can-
not be overlooked in deciding how much risk of impairment of fish produc-
tion by waste discharges is to be deemed acceptable (Warren, 1971, pp.
15-23, 375-386).
After the test animal to be used, a suitable test temperature and ex-
posure period, and an appropriate application factor has been carefully
chosen, an announced enforcement of this regulation next must be under-
taken. To determine whether or not the concentration of an acutely toxic
effluent that has been judged permissible in a receiving body of water is
outside an allowable dispersion or mixing zone, the amount of dilution of
the effluent within the mixing zone must be known. Only a concentration
of the effluent (percent by volume) equal to the product of its concentra-
tion in the receiving water (at the boundary of the mixing zone) and the
reciprocal of the prescribed application factor needs to be tested for
toxicity. If, at any time, this concentration is found to kill more than
50 percent of the test animals in the specified exposure period, either
the toxicity of the effluent or the rate of its discharge may be regarded
as excessive. If 50 percent or more of the test animals survive in such
tests, the permissible concentration is not exceeded, and hence, the reg-
ulatory requirement is not being violated.
The above procedure is applicable to all effluents that are toxic
enough to kill at least half of the test animals in the prescribed test
period when they are not diluted. If an effluent is of low toxicity, but
the product of its concentration (percent by volume) and the reciprocal
of the prescribed application factor is greater than 100 percent, one can-
not reasonably conclude from the result of the acute toxicity test that
aquatic organisms are not being endangered. The possibility of serious
chronic or sublethal toxicity of the effluent cannot be judged negligible
in the absence of better evidence. Tests for toxicity of such effluents
that have no readily measurable acute toxicity must be required to pro-
tect aquatic life adequately.
Reliance on acute toxicity tests alone in regulating discharges of
any toxic wastes can occasionally lead to serious error. The regulatory
practices considered above are based on certain assumptions which cer-
tainly cannot be always valid. The chronic or sublethal toxicity of a
complex industrial waste at low concentrations can be quite independent
of its acute toxicity at much higher concentrations, since the causative
agents of these variations in toxicity can be entirely different com-
ponents of the wastes. Further, environmental variables, such as natural
water quality and temperature, can influence toxicity in very different
-------
ways, even when the same toxicant is the active agent at low and high con-
centrations. A species of fish that is more resistant than another to
the lethal action of a toxicant can be more susceptible than the other to
sublethal injury by the same compound.
The value of acute toxicity tests in the assessment and control of
pollution can be often reasonably questioned. For entirely different rea-
sons, chemical data may also be very misleading. Until analytical
methods are perfected, and the problems of interpretation of the chemical
information are solved, continued heavy reliance on toxicity tests of
short duration appears to be warranted.
For some purposes, the acute toxicity test is clearly the best
possible test. Even when dilution of toxic effluents is sufficient to
prevent damage to aquatic life outside the mixing zone, fish may be
killed when they enter the mixing zones if the dilution is not very rapid.
To avoid fish kills in the immediate vicinity of wastewater outfalls, it
may be necessary to limit the toxicity of the effluent without regard to
the amount of further dilution. It is evident that the toxicity must be
sufficiently low so that the exposed fish will be overcome rapidly by the
effluent, unless they are known to be attracted by the effluent or likely
to remain in it for long periods for other reasons. The safe level of
acute toxicity varies with the area of the mixing zone as well as the
ability of fish to avoid high effluent concentrations. Acute toxicity
tests and limits are most appropriate for the regulation of some infre-
quent discharges of toxic wastes that are of sufficiently short duration,
that there is no need to protect organisms against chronic toxicity of
the wastes. The toxicity of intermittantly discharged wastewater can be
relatively high without causing serious damage, but the amounts or rates
of their dilution, and the duration of the discharges must be considered
prior to setting limits.
Some regulatory authorities have favored much more stringent and uni-
form toxicity limits independent of the amounts of dilution of the ef-
fluents, the frequency, and the duration of the discharges. For example,
96-hour survival has been required of an average of at least 90 percent,
or even of no less than 80 or 90 percent at any given time (in every
test), of prescribed test animals (fish) in undiluted effluents of various
kinds. Specifically, application has been made to pulp and paper mill
effluents diluted to 65 percent of their full strength, without regard to
plant location. Various arguments have been advanced in support of such
uniform requirements unrelated to the assimilative capacities of waters
receiving the wastes. The enforcement of this type of regulation is
simplier than the limitation of toxic waste concentrations in the
receiving waters, and some people believe that this solution more equit-
able than the latter restrictions. However, when the dilution factor
related to wastewaters is great, the requirements in question can be
quite unnecessarily restrictive, necessitating costly waste treatment or
other expensive measures that result in no real benefits. On the other
hand, when the dilution is slight, the same requirements can be quite
inadequate for protection of aquatic life against chronic or sublethal
toxicity of the wastes.
4
-------
For the above reasons and others beyond the scope of this paper, I
strongly disapprove of such pollution-control measures or requirements.
They tend to discourage the selection of sites for industrial plants
where effects on the environment will be minimized, because they do not
permit reduction of waste disposal costs by choosing the more favorable
locations. They also discourage the conservation of water by industry,
because more frugal use of water that results in great reductions in
volume of effluents, usually results also in some increases of the con-
centrations of toxicants in the effluents.
Instead of arbitrarily limiting the concentrations of toxicants in ef-
fluents for protection of aquatic life outside the mixing zones, it is
surely more reasonable to limit the amounts of these harmful substances
discharged per unit of time. These amounts can decrease while the toxi-
cant concentration increase, if the volume of the effluents discharged
per unit of time is simultaneously reduced. The measured toxicity of a
wastewater is a function of the concentration of the toxicant or mixture
of toxicants present. If all poisons were equally toxic, it would be a
measure of their total concentration whenever several are present. The
toxicity of any solution expressed in toxicity units (t.u.), sometimes
called the "toxicity concentration". This is equal to the reciprocal of
its median lethal concentration (LC5Q) expressed as the decimal fraction
by volume (percent by volume -f 100]. For example, if the 1059 is 0.2 (or
20 percent) by volume, the toxicity is 1/0.2 or 5.0 t.u. An application
factor prescribed for an effluent is equal numerically to its permissible
concentration at the boundary of the mixing zone expressed in appropriate
(corresponding) toxicity units. The expression of toxicity levels in
such units has been shown by experiments to be useful in the estimation
or prediction (by summation) of the toxicities of various mixtures of
poisons whose individual toxicities have been determined (Brown, 1968;
Warren, 1971, p. 210).
An approach to the regulatory problems that has recently been gaining
favor in the United States and Canada is to express the output of toxi-
cants from each important source as a value to which the name "toxicity
emission rate" (T.E.R.) has been given (California State Water Resources
Control Board, 1972). This value can be computed by dividing the rate of
flow of the effluent by the determined median lethal concentration ex-
pressed as a decimal function by volume. For example if the determined
96-hour median tolerance limit or LC50 of an effluent is 0.20 (20 percent)
by volume, and its flow rate is 2.0 m3/min.:
3 3
T F p _ 2 _ in m ^^ _ ,n t.u. • m
I.L.K. - 0-2 1U 96-hr. LCKn . min. IUmin. '
t)U
where t.u. represents toxicity units based on results of 96-hour tests.
Uniform requirements that limit the T.E.R. per unit of industrial pro-
duction of a given kind (for example, per ton of cellulose pulp produced
per day by a sulfate-process pulp and paper mill) would be much more rea-
sonable than are those that simply limit the toxicity of the effluents
-------
without regard to their discharge rate or volume. However, for reasons
already noted, I cannot fully approve any waste-disposal regulations that
are entirely independent of local conditions and needs, the manner of dis-
charge of the wastes, and the ability of receiving waters to assimilate
them without impairment of any beneficial uses of the waters. I believe
that toxicity emission rates can best be limited so as always to permit
reasonable utilization of the assimilative capacity of the receiving
waters with no undue risk of injury to aquatic life by toxicants, the
calculation and control of these rates may prove especially useful in pro-
tecting valuable organisms in waters that receive toxic wastes in con-
siderable amounts from several sources.
Procedures for equitable apportionment of the assimilative capacities
of waters for toxic pollutants among multiple sources of these wastes
have not yet been thoroughly developed, and will not be discussed in de-
tail here. Much attention recently has been given to this difficult
regulatory problem in the United States, especially in California.
Various proposals for its solution have been advanced and studied, but
none, as yet, have been shown to be entirely sound nor widely accepted.
When there are several important sources of toxic pollution of a water
body, evaluation and consideration of the persistence, as well as the
dilution, of the discharged toxicants in the receiving water may be
necessary. The acute toxicity of combined, fresh effluents from different
sources not separated by large distances can often be easily determined
experimentally. Perhaps the combined toxicity can also be estimated with
sufficient accuracy for regulatory purposes by measuring the toxicity of
each effluent, and assuming an additive interaction of the poisons in the
mixture. This assumption is implied by the proposed summation of toxicity
emission rates. But when effluents from several plants are discharged
into a body of water at points that are remote from each other, the non-
persistent toxicants from the different sources do not occur together in
the receiving water. Unless most of the toxicants involved are very per-
sistent, the safe discharge rates for the effluents can be much greater
than they would be if all the effluents were discharged together at one
point into a common mixing zone. The proper allowances to be made for
natural self-purification of waters are not easily determined, and very
little is known about the interaction of poisons at sublethal concentra-
tion levels. Measurements of rates of loss of toxicity under appropriate
conditions in the laboratory can be useful in estimating levels of resi-
dual toxicity of polluted waters in nature. Without some such correction
for natural self-purification, a value for the "toxicity concentration"
(in t.u.) at a given point in a stream, computed by dividing the sum of
the toxicity emission rates from all upstream sources by the stream flow
(in m3/min., for example), is not meaningful. But a correction for loss
of toxicity can introduce an error in the opposite direction when the re-
sults of the calculation are applied in regulating the waste discharges.
Because slowly acting, accumulative poisons tend to be also highly per-
sistent, the appropriate application factor for the acute toxicity of a
solution of nonpersistent and persistent poisons is likely to decrease as
the solution ages.
-------
REFERENCES
American Public Health Association, American Water Works Association, and
Water Pollution Control Federation. 1960. Standard methods for the
examination of water and wastewater. Eleventh edition. American
Public Health Association, Inc., New York.
Aquatic Life Advisory Committee of the Ohio River Valley Water Sanitation
Commission. 1955. Aquatic life water quality criteria -- first
progress report. Sewage and Industrial Wastes, 27: 321-331.
Brown, V.M. 1968. The calculation of the acute toxicity of mixtures of
poisons to rainbow trout. Water Research, 2: 723-733.
California State Water Resources Control Board. 1972. Water quality con-
trol plan for ocean waters of California. State Water Resources
Control Board, Resolution No. 72-45. Sacramento, California.
Committee on Water Quality Criteria, National Academy of Sciences and
National Academy of Engineering. 1972/1973 (both dates given).
Water quality criteria 1972. U.S. Environmental Protection Agency,
Washington, D.C., Ecological Research Series EPA-R3-73-033.
Doudoroff, P., B.G. Anderson, G.E. Burdick, P.S. Galtsoff, W.B. Hart,
R. Patrick, E.R. Strong, E.W. Surber, and W.M. Van Horn. 1951.
Bioassay methods for evaluation of acute toxicity of industrial
wastes to fish. Sewage and Industrial Wastes, 23: 1380-1397.
Hart, W.B., P. Doudoroff, and J. Greenbank. 1945. The evaluation of the
toxicity of industrial wastes, chemicals and other substances to
freshwater fishes. Waste Control Laboratory, Atlantic Refining
Company, Philadelphia, Pa.
National Technical Advisory Committee on Water Quality Criteria. 1968.
Water quality criteria -- report of the National Technical Advisory
Committee to the Secretary of the Interior. U.S. Federal Water
Pollution Control Administration, Washington, D.C.
Warren, C.E. (with P. Doudoroff). 1971. Biology and water pollution con-
trol. W.B. Saunders Company, Philadelphia, Pa.
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SECTION 2
TOXICOLOGICAL CONTROL OF POLLUTION OF FRESHWATERS
N.S. Stroganov
The methods of chemical, microbiological and sanitary/hydrobiological
analysis have been well established for the evaluation of water quality.
Each of the methods allows the characterization of water in any single
aspect, i.e., chemical, epidemiological or sanitary significance.
At the present time, various toxic substances such as petroleum and
its associated products, pesticides, heavy metals, detergents, metal
organic compounds and many others are discharged into surface waters in
ever growing amounts. The toxic substances impart an altered quality to
the receiving waters, to which the aquatic organisms react with extreme
sensitivity, but which is not evaluated by the methodology noted earlier.
Thus, toxicology evaluations of water quality are required.
As industry develops, and as chemistry advances in various fields
related to the economy, the increasing potential for pollution of the
water requires the organization of toxicologic control of waste waters to
assure continued quality of all large water-bodies and to provide more
reliable protection for these surface waters. Using integrated models,
toxicological control must characterize the quality of water in relation
to the suitability for the life of aquatic organisms, it is expedient to
exercise control of the toxicity of waters in two ways:
1. Evaluation of the toxicity of waste waters discharged
from industrial sources, sewage plants, and points of
population concentration, and
2. Evaluation of the toxicity and associated hazards of
waters for aquatic organisms.
These varieties of control characterize both the quality of waters
entering a water-body and the waters of the reservoir itself in which the
self-purification has already begun or become pronounced. The essence of
such toxicological control may be divided into two categories as follows:
The First Variant
Three organisms having varying degrees of sensitivity are suggested
for testing for acute toxicity: waterfleas (Vaphnia magna, Straus) as a
-------
sensitive organism; the large pond snail, Lonnea t>ta.QnaJiij> as an organism
of intermediate sensitivity; and the guppy, L&biAtu /j.e£ccu£a£o6, a fish
of least sensitivity. The duration of the test is five days at a tempera-
ture of 17-22 °C. Survival and general condition (behavior) are the
criteria of toxicity to the organism. Toxicity is estimated using a five
degree system:
1. First Degree. Very acutely toxic. All organisms of
the three species die within the first day.
2. Second Degree. Acutely toxic. All organisms of the
organisms of the three species die within five days.
3. Third Degree. Toxic. From 70 to 100 percent of the
Vaphyiia die; not more than 20 percent of the pond snails
expire; and all guppies survive.
4. Fourth Degree. Slightly toxic. The mortality of Vaphnia.
does not exceed 30 percent; and all of the pond snails and
guppies survive.
5. Fifth Degree. Conventionally non-toxic. All of the or-
ganisms of the three species survive, and by outward ap-
pearance and behavior do not differ from control organisms.
When the variability in the degree of toxic waste water is considered,
control of periodically diluted waste water is recommended. In this re-
gard, waste water is diluted in the following order: 0 (initial), 5, 10,
100, 500 times with clean water from a river or brook containing no toxic
substances. The scheme of the tests is identical to the initial waste
water. The results obtained are expressed graphically by plotting the
degree of dilution against the degree of toxicity. The resultant slope
of the curve suggests either the rate of increase characterizing the toxi-
cological danger of the waste water, or the rate of loss of the toxicant
with dilution.
The first variety of control of toxicity may be used for differential
determination of the degree of toxicity of waste waters discharged from
different sources within a single industry. This enables the detection
of the most dangerous sources and allows these waters to be directed to
special treatment.
The Second Variant
The waters of large rivers, reservoirs and lakes do not possess the
acute toxicity of waste waters, although they receive these waters. How-
ever, even in low concentrations, a prolonged influence of toxic
substances on aquatic organisms leads to death or diminution of the num-
bers of the most sensitive of these organisms. These conditions transform
individual aquatic communities and the ecosystem as a whole. Therefore,
toxicological control is also necessary for the waters of rivers, reser-
voirs, and lakes which undergo anthropogenic influence.
-------
The toxicological control of the second variety is assayed by pouring
the water to be tested into 10 beakers, 100 ml in each. Clean water in
the same quantity serves as a control. To each of the beakers one-two
day old Vaphnia magna. are added. The duration of the experiment is 30
days at a temperature of 17-22 °C. The Vapknia. are fed with CklpnMa. 4p .
and the water is changed every 3-4 days. The criteria of toxicity in-
clude survival, rate of reproduction (fecundity), time of maturation, and
frequency of moulting.
The toxicity or harmfulness of the water for the organism is expressed
by three degrees:
1. First Degree. Water is toxic. 50 percent or less of
live less than 30 days.
2. Second Degree. Water is harmful. 80 percent or more
of the Va.phru.0. survive for 30 days, hut fecundity is
reduced by 25 percent or more as compared with control
populations.
3. Third Degree. Water is clean. All the Va.ph.vua. survive,
and the fecundity is not reduced by more than 25 percent
in comparison with the control.
By simultaneously conducting toxicologic control of the first variety
every 5-7 days while running the longer experiment of the second variety,
the occurance of late results is minimized. Compensation is achieved by
exerting simultaneous control of the discharge water and the receiving
water of the water body. If periodic discharge occurs, detection is en-
abled by the toxicological control of the first variety.
Toxicological control does not indicate the nature of the toxicant,
but does show the danger of the water in question for many aquatic or-
ganisms.
10
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SECTION 3
TOXICITY OF EXPERIMENTAL FOREST INSECTICIDES TO
FISH AND AQUATIC INVERTEBRATES
Richard A. Schoettger and Wilbur L. Mauck
INTRODUCTION
In 1952, the U.S. Forest Service began aerial spraying of insecticides
to control outbreaks of the spruce breakdown ( Cho'uAtone.uAa fiam^&iand).
The first forest insecticide to be applied was the organochlorine com-
pound, DDT, which was commonly applied at a rate of 1.12 kg per 9.3 liters
per hectare (1 pound DDT per 1 U.S. gallon per acre). Applications of DDT
continued for nearly 20 years before its further use was prohibited in
1969 by the U.S. Environmental Protection Agency. However, this com-
pound was banned in the State of Maine for use as a forest insecticide as
early as 1967.
DDT was of particular concern to environmental scientists because of
its persistence, toxicity and bioaccumulation in non-target organisms. As
early as 1963, DDT was recognized as an imminent hazard to terrestrial,
avian, and aquatic fauna by the President's Science Advisory Committee.
This committee recommended a reduction in the use of persistent pesti-
cides. Certain uses of DDT were prohibited by the U.S. Department of Agri-
culture (USDA) in 1969 and an informal review of the remaining uses con-
tinued through 1970. On the basis of some 10,000 pages of testimony by
more than 50 scientific experts, the Administrator of the Environmental
Protection Agency issued an opinion and order, published July 7, 1972 in
the Federal Register, cancelling or suspending all uses of DDT, except
those related to human health (USDA, 1973).
After the banning of DDT, government officials and environmentalists
urged the development and use of insecticides that were highly specific in
their action, were more readily biodegradable, and did not bioaccumu-
late. Because research on radically new methods of control might consume
excessive time or be unproductive, the U.S. Forest Service decided that
investigations of DDT replacements should be directed only to chemicals
already in production, or experimental compounds near the production stage
(Schmiege oJL at., 1970). Also, the compounds selected should be applied
from the air, by spraying procedures that had been developed for DDT.
Three conditions were to be met: (1) the insecticide should be more toxic
to the western spruce budworm than to other organisms; (2) the insecticide
and its degradation productions must not accumulate in plants or animals
11
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found in forest ecosystems; and (3) the insecticide spray should be
directed efficiently to the target insect (Schmiege aJi al., 1970).
Currently, the U.S. Forest Service is evaluating the efficacy of
several organophosphate, carbamate insecticides, and insect growth regula-
tors for the control of three major insect pests: the gypsy moth
(?ofctk&&iia dlbpasi), Douglas-fir tussock moth (H
-------
to maintain the recreational and aesthetic values of these forest areas
(USDA, 1975). However, streams in forest areas also support important
fisheries, particularly trout. Some watersheds of the northeast are
drained by major spawning and nursery streams for brook trout (SalveLinuA
fiontinatlb) and Atlantic salmon (Salmo koJLoJi]. Therefore, the Fish-
Pesticide Research Laboratory of the U.S. Fish and Wildlife Service is
cooperating with the U.S. Forest Service in investigating the potential
toxicological effects of candidate forest insecticides on fish and
aquatic invertebrates. The investigations were designed to evaluate
changes in toxicity due to various water qualities associated with bio-
geographic regions. The investigations also included the toxicity of the
candidate insecticides to aquatic invertebrates, to early life stages of
brook trout, the possible toxic interaction of insecticide combinations,
and the susceptibility of fish containing residues of DDT, or a poly-
chlorinated biphenyl (PCB; Aroclor 1254). This paper reports progress in
our cooperative research.
MATERIALS AND METHODS
Six experimental forest insecticides,3umithion , carbaryl, Dylox ,
Matacil®, Dimilin© (TH-6040), and Orthenedvwere provided by various
chemical companies. Stock solutions of the candidate insecticides
(technical grade) in reagent grade acetone were prepared immediately be-
fore each static test. Stocks of field formulations were prepared by
dilution with distilled water.
Test waters of different chemical characteristics were formulated from
deionized water of at least 1 million ohms resistivity by adding reagent
grade salts (Marking, 1969). Mineral acids and bases were used to buffer,
adjust, and maintain pH (Marking and Dawson, 1972). Various test tempera-
tures were controlled by water baths.
Fish obtained from Federal and State hatcheries were maintained for 2
weeks under standard fish cultural care (Brauhn and Schoettger, 1975).
They were acclimated to test conditions of temperature and water quality
before the experiments and subsequently transferred to test containers
about 24 hours before addition of the toxicant (Committee on Methods for
Toxicity Tests with Aquatic Organisms, 1975). Fish used in this investi-
gation were brook trout and Atlantic salmon. Mature scuds
pAe.udoLunnae.iu>) and late instar naiads of a stonefly
c.aLL^ofLnic.0.) were used in the toxicity tests for invertebrates. The in-
vertebrates were obtained from wild populations in streams and maintained
in the laboratory (Committee on Methods for Toxicity Tests with Aquatic
Organisms, 1975).
Toxicity data were analyzed with the statistical method described by
Litchfield and Wilcoxon (1949) to determine the LC5Q (concentration pro-
ducing 50% mortality) and 95% confidence interval. Estimations of toxic
chemical interactions were made with a modification of the methods
developed by Marking and Dawson (1975). With this method, toxicities of
paired insecticide mixtures were determined in the manner of the indi-
13
-------
vidual chemicals, except that the two chemicals were added as fractions
of their activity (96-h LC^o's), and in a 1 to 1 ratio of their inde-
pendent toxicities. Toxicity of the mixture of chemicals was computed by
summing the LCso ratios of the combined to the independent toxicities of
each chemical in the mixture, and calculating an interaction index (see
footnote, Table 3). Evaluations of potential additive activity, and
activity greater or less than additive activity were made by computing a
range of index values for the 95% confidence intervals around LCso values.
RESULTS
Acute Toxicity
Fish—
Toxicities of the candidate forest insecticides ranged between 0.5 to
11 mg/1 (96-h LCso's), with exception of Orthene and the field formula-
tion of Matacil (Table 1). Orthene was the least toxic; up to 100 mg/1
did not kill fish within 96 h. The toxicities of Sumithion and Orthene
(both organophosphate insecticides) were apparently unaffected by changes
in temperature (7-17 °C), water hardness (40-320 mg/1, as CaCOs), or ph
(6.5-9.0). Further, the toxicities of these two field formulations were
similar to those for their technical forms. In contrast, the toxicity of
technical grade carbaryl, a carbamate insecticide, was influenced by
water quality. It was 3 times more toxic to brook trout at 17 °C than at
7 °C; and was most toxic in hard, alkaline water, e.g., it was 4 times
more toxic at pH 9.0 than at pH 6.5 (Table 1). The field formulation of
carbaryl was tested for toxicity. Because it is not soluble in acetone
or water, there were no mortalities, and its toxicity could not be rea-
sonably established with these techniques.
Matacil is also a carbamate insecticide, and, like carbaryl, its toxi-
city was appreciably influenced by temperature and pH, but not by water
hardness (Table 1). It was over 3 times more toxic at 17 °C than at
12 °C, and over 6 times more toxic at pH 9.0 than at pH 7.5. However,
the most significant finding from these tests was the discovery that the
field formulation of Matacil (17% active ingredient) was as much as 70
times more toxic than its technical form and could pose a serious hazard
to trout.
The toxicity of the organophosphate insecticide Dylox to brook trout
was influenced by temperature, water hardness, and pH (Table 1). The
toxicity was 18 times greater at 17 °C than at 7 °C; almost 10 times
greater at pH 8.5 than at pH 6.5; and about 4 times more toxic in very
hard water (320 mg/1) than in soft water (40 mg/1). The toxicity of the
field formulation of Dylox (40% active ingredient) was similar to that of
the technical grade material.
Dimilin is a comparatively new growth regulator that inhibits chitin
synthesis during the molting of immature insects. No extensive toxicolog-
ical tests have been conducted with this compound, but preliminary data
indicate that it is relatively non-toxic to fish.
14
-------
TABLE 1. ACUTE TOXICITY1 OF SIX CANDIDATE FOREST INSECTICIDES TO BROOK TROUT IN WATER OF
DIFFERENT TEMPERATURE, HARDNESS, AND pH.
CJl
Compound
(% active
ingredient
Carbaryl
(Technical, 99.5%)
(Field formulation,
49%, Sevin-4-Oil®)
Temp.
(C)
7
12
17
12
12
12
12
12
12
12
Test Water
Water
hardness
(mg/1 as CaCOs)
40
40
40
40
320
40
40
40
40
PH
7.5
7.5
7.5
8.0
8.0
6.5
7.5
8.5
9.0
40 7.5
(continued)
96h-LCt^Q and 95% confidence interval
(mg/1)
3.0
2.0-4.5
2.1
1.7-2.6
0.9
0.7-1.3
5.4
4.5-6.5
2.5
1.8-3.5
4.6
3.5-6.0
3.7
2.0-6.7
2.1
1.7-2.6
1.1
0.8-1.6
___ 2/
-------
TABLE 1 (Continued)
Compound
(% active
ingredient
Metacil
(Technical, 98%)
(Field formulation,
17%)
Temp.
(C)
7
12
17
12
12
12
12
12
12
12
Test Water
Water
hardness
(mg/1 as CaCCh)
40
40
40
40
320
40
40
40
40
40
PH
7.5
7.5
7.5
8.0
8.0
6.5
7.5
8.5
9.0
7.5
96h-LC5o and 95% confidence interval
(mg/1)
9.4
6.9-13
9.0
6.2-13
2.6
2.1-3.2
8.0
5.8-11
11
8.7-13
8.6
5.0-15
9.0
6.2-13
7.8
5.2-12
1.4
0.9-2.1
0.13
0.10-0.17
(Continued)
-------
TABLE 1 (Continued)
Test Water
Compound
(% active
ingredient
Dylox
(Technical, 99%)
(Field formulation,
40.5%)
Orthene
(Technical, 94%)
Temp.
(C)
7
12
17
12
12
12
12
12
12
7
12
12
Water pH
hardness
(mg/1 as CaCOs)
40
40
40
40
320
40
40
40
40
40
40
40
7.5
7.5
7.5
8.0
8.0
6.5
7.5
8.5
7.5
7.5
7.5
7.5
(Continued)
96h-LC5Q and 95% confidence interval
(mg/1)
9.4
7.9-11
2.5
2.2-2.9
0.50
0.38-0.65
2.4
2.1-2.8
0.62
0.49-0.79
9.2
7.8-11
4.4
2.9-6.5
0.96
0.78-1.2
5.5
4.6-6.58
>100
>100
>100
-------
CO
Compound
(% active
ingredient
Orthene
(Technical, 94%)
(Field formulation,
75%)
Sumithion
(Technical, 95%)
(Field formulation,
40%)
Temp.
(C)
12
12
12
12
12
12
7
12
17
12
12
TABLE '
Test Water
Water
hardness
(mg/1 as CaCQg)
40
320
40
40
40
40
40
40
40
40
320
1 (Continued)
PH
7.5
7.5
6.5
9.0
7.5
7.5
7.5
7.5
7.5
8.0
8.0
96h-LCso and 95% confidence interval
(mg/1)
>100
>100
>100
>100
>100
1.7
1.4-2.1
2.1
2.0-2.5
2.2
2.0-2.5
1.8
1.5-2.0
2.0
1.7-2.4
1.4
1.1-1.7
(Continued)
-------
TABLE 1 (Continued)
Compound
(% active
ingredient
Sumithion
(Field formulation,
Dimilin
(Field formulation,
Temp.
(C)
12
12
12
40%)
12
10
25%)
Test Water
Water
hardness
(mg/1 as CaCO^}
40
40
40
40
40
PH
6.5
7.5
8.5
9.0
7.5
96H-LC50 and 95% confidence interval
(mg/1)
1.7
1.4-1.9
1.7
1.3-2.1
1.5
1.1-2.1
1.6
1.2-2.0
>100
1
Toxicity based on active ingredient in static tests.
?
"Not determined.
-------
TABLE 2. ACUTE TOXICITY1 (96-h LCso AND 95% CONFIDENCE INTERVALS, mg/1)
OF SIX CANDIDATE FOREST INSECTICIDES TO MATURE SCUDS AND
STONEFLY NAIADS TESTED AT 18 °C and 16 °C
Invertebrate and Test Temperature
Compound Scud Stonefly
(% active
ingredient) 18 °C 16 °C
Sumithion >0.01 0.004
(95%) 0.002-0.006
Matacil 0.012
(98%) 0.008-0.018
Dylox 0.040 0.035
(99%) 0.026-0.060 0.022-0.055
Carbaryl 0.016 0.002
(99.5%) 0.012-0.019 0.001-0.003
Orthene >100
(94%)
Dimilin 0.030
(25%) 0.020-0.050
Toxicity based on active ingredient in static tests.
2A11 chemicals tested were technical grade material except the field
formulation of Dimilin.
20
-------
Invertebrates--
The 96-h LCso's for scuds and stoneflies were in the range of 0.002 to
0.04 mg/1 of the insecticides tested, with the exception of Orthene (Table
2). In general, invertebrates appear to be about 100 times more suscep-
tible to these chemicals than are brook trout. Although Dimilin was not
toxic to brook trout, it was highly toxic to scud. Of all six compounds,
Orthene was by far the least toxic to both brook trout and the aquatic in-
vertebrate tested.
Toxicity to Eggs and Sac-Fry
To date, studies of the effects of insecticides on life stages of
brook trout have been completed only for Sumithion. Eyed eggs (2 days be-
fore hatch) were exposed to concentrations of Sumithion similar to those
detected in Maine streams after experimental aerial applications of the
compound (Marancik, 1976). These tests were conducted in a flow-through
system described by Mount and Brungs (1967), and test concentrations were
reduced by one-half for 4 consecutive days and then stopped. For example,
with the highest concentration tested, eggs were exposed to 0.1 mg/1 on
day 1, 0.05 mg/1 on day 2, and the sac fry to 0.025 mg/1 on day 3, and
0.012 mg/1 on day 4. Treatment was then stopped, but the fry were ob-
served for an additional 30 days to monitor anomalies or delayed morta-
lity. The Sumithion concentration of 0.1 mg/1, which was several times
greater than the highest post-treatment concentration reported by Maranick
in streams, did not significantly (p=0.05) affect survival or development
of brook trout sac fry.
Chemical Interactions
The U.S. Forest Service anticipates that several of the candidate
forest insecticides may be applied to forests at about the same time, con-
currently for different pests, or that applications of different compounds
to infested areas may overlap. In addition, fish such as Atlantic salmon
and brook trout containing background residues of DDT and PCB's may also
be exposed to Guthion (an organophosphate insecticide), which is used in
blueberry culture. Therefore, the likelihood of salmon and trout receiving
multiple chemical exposures is high, and potential toxic interactions must
be explored.
Insecticide Interactions--
Brook trout and Atlantic salmon were exposed to paired mixtures of
candidate forest insecticides, or to a combination of Dylox and Guthion.
The toxicities of all insecticide combinations were simply additive, ex-
cept for the Dylox and Guthion mixture which was synergistic in brook trout
and Atlantic salmon (Tables 3 and 4). This synergistic action of Dylox
and Guthion was also observed by Marking and Mauck (1975) in studies with
rainbow trout (SaJtmo gcuAdnvi).
Insecticide, DDT, and PCB Interactions--
The carbamate insecticides carbaryl and Matacil were about twice as
toxic to brook trout containing Aroclor 1254 (PCB) residues of 2.3 yg/g as
compared with trout containing lower residues (Table 5). However, there
21
-------
appeared to be no interaction of the organophosphate insecticides Dylox
and Sumithion with Aroclor 1254 residues in this species.
Mixtures of Dylox and Guthion were essentially as synergistic to
Atlantic salmon as they were to brook trout, but young salmon containing
total body residues of 1.5 yg/g DDT (and its analogs) were no more suscep-
tible to the mixture than were those without residues (Table 4). Aroclor
1254 residues of 2.3 yg/g in brook trout did not appreciably alter the re-
lative synergism of the Dylox with Guthion, or the additive toxicity of
any of the other forest insecticide mixtures (Table 3).
DISCUSSION
Sumithion, carbaryl, Dylox and Matacil are all relatively much more
toxic to young brook trout and Atlantic salmon than are Orthene and
Dimilin. However, barring accidental spills or excessive or overlapping
applications, the toxic concentrations determined (with one exception) are
well above those measured in streams after experimental aerial applica-
tions. The exception, the liquid field formulation of Matacil, was 10 to
70 times more toxic to brook trout than its technical grade form, and con-
centrations exceeding the 96-h LCso of 0.13 mg/1 could be expected in
streams after aerial applications. The hazard connected with use of this
formulation would be increased in alkaline waters. Alkaline pH and high
water temperatures increased the toxicity of several of the candidate
forest insecticides to fish, but for most of these chemicals, even this
elevated toxicity does not appear sufficient to pose a significant toxicity
hazard.
Although some of the candidate forest insecticides appear to be syner-
gized by other insecticides, this reaction is not likely to be an imminent
hazard to brook trout and Atlantic salmon, unless streams receive far
greater doses of the chemicals than have thus far been measured in experi-
mental aerial applications. A similar relationship exists in fish contain-
ing Aroclor 1254 residues, but the relatively toxic field formulation of
Matacil could be even more toxic to fish containing significant Aroclor re-
sidues. Chemical interactions in invertebrates were not tested. However,
considering the high toxicity of all of the compounds except Orthene to
these organisms, synergistic interactions could have a dramatic effect.
The high susceptibility of scuds and stonefly naiads to Sumithion,
carbaryl, Dylox, Matacil, and Dimilin suggests that aquatic invertebrates
are much more sensitive to these compounds than are fish. In addition,
the LCso values appear to be well within the concentrations measured in
streams after experimental aerial applications. Insecticide concentra-
tions of 0.24 mg/1 (Haugen, 1976) to 0.013 mg/1 (Marancik, 1976) have been
measured at 20 minutes and 24 hours, respectively, after application.
However, Orthene should not pose a significant toxicity hazard to fish or
invertebrates.
22
-------
TABLE 3. EFFECT OF PCB (AROCLOR 1254) RESIDUES ON SENSITIVITY1 OF BROOK TROUT TO
CANDIDATE FOREST INSECTICIDES
ro
CO
PCB residues in trout and
insecticide mixture tested
Aroclor 1254 residues,
0.08 yg/g\ or less
Carbaryl , Dylox
Carbaryl , Sumithion
Sumithion, Dylox
Guthion, Dylox
Guthion, Dylox
Aroclor 1254 residues,
2.3 yg/g"
Carbaryl , Dylox
Carbaryl, Sumithion
Sumithion, Dylox
Guthion, Dylox
Guthion, Dylox
Interaction Range3
index2 -3.0 -2.0 -1.0 0 +1.0 +2.0
III 1 1 1 1 1 1 1 1 1 1 1 1
-0.68 |
-0.30 j
+0.33
+0.65
+0.93
-0.88
-0.29
+0.10
+0.67
+0.70 .
1 1 1 1 1 1 1 1 1 1 1 1
1
i
i
i
i i
i i
I i
1 i
j
i
i
i
i
i
i
Tests conducted in standard reconstituted water at 12 °C.
Interaction Index (I.I.) = ft? iL^50! + n™ K \ where Ai and Bi are independent LC50's of insecticides A and B,
MI iu-50) BI IU.50J and Am and Bm are LCso's for A and B calculated from tests of
the mixture.
Ranges that cross the zero line indicate addivitive toxicity; ranges on the positive side indicate greater than
additive toxicity; and ranges on the negative side indicate less than additive toxicity.
Mean total body residues.
-------
TABLE 4. EFFECT OF DDT RESIDUES ON SENSITIVITY1 OF ATLANTIC SALMON TO A MIXTURE OF GUTHION AND DYLOX
3
Interaction Range
DDT residues in salmon index^ -1.0 0 +1.0 +2.0
1 1 1 1 1 1 1 1 1
DDT residues, less
than 0.01 jjg/g
mixture +0.56
DDT residues, 1.5 yg/g
mixture 0.95
1 I I I 1 1 1 1 1 | 1 1 1 1 1 1 1 1 1 | I 1 I I 1 1
I i
f )
1 C
I f
Tests conducted in standard reconstituted water at 12 °C.
Interaction index (I.I.) = -" !~50! + IJm
where Ai and Bi are independent LCso's of insecticides A and B,
at50j tn ILL50; and Am and Bm gre LC50.S for A and B caicuiated from tests of
the mixture.
Ranges that cross the zero line indicate additive toxicity; ranges on the positive side indicate greater than
additive toxicity, and ranges on the negative side indicate less than additive toxicity.
Mean total body residue of DOT, including DDT, DDE, and ODD.
-------
TABLE 5. EFFECT OF PCB (AROCLOR 1254) RESIDUES (TOTAL BODY) ON
SENSITIVITY OF BROOK TROUT TO FOUR FOREST INSECTICIDES
Residues of 96-h LCso and 95%
Aroclor 1254 in fish Confidence Interval
Insecticide (yg/1) (mg/1)
Carbaryl ^2 ^
5.1-6.9
0.4 5.0
4.1-6.0
2.3 2.5
1.7-3.7
Matacil Q>082 n
8.8-13
0.4 10
7.7-14
2.3 6.5
4.5-9.0
DylOX 0.82 3.8
3.1-4.6
0.4 4.6
3.6-5.8
2.3 3.9
3.0-5.1
Sumithion Q>082 K6
1.4-1.7
0.4 1.5
1.3-1.7
2.3 1.7
1.5-2.0
Toxicity based on active ingredient; tests conducted with 0.3-g fish at
12 «C.
2
Background Aroclor 1254 residues.
25
-------
In summary, aerial application of the six potential DDT substitute
forest insecticides, with the exception of the field formulation of
Matacil, should not have a major toxic effect on brook trout and Atlantic
salmon. Marancik (1976) reported some inhibition of brain cholinesterase
in fish after aerial applications, but enzyme activity returned to normal
within 48 h. Nevertheless, all of the candidate insecticides, except
Orthene, may kill aquatic invertebrates. Results published by others
(e.g., Burdick e£aZ.t 1960; Elson eX aJL., 1973; Flannagan, 1973) showed
that populations of stream invertebrates were markedly reduced after
aerial applications with forest insecticides other than DDT. Therefore,
field investigations are needed to determine whether any of the present
insecticides significantly depress aquatic invertebrate populations, and,
if so, the duration of the population depressions. The timing of such
effects on invertebrates may be critical with respect to adequate food
supplies for young salmon and trout.
REFERENCES
Batzer, H.O. 1973. Net effect of spruce budworm defoliation on mortality
and growth of balsam fir. J. For. 73: 34-37.
Brauhn, J.L., and R.A. Schoettger. 1975. Acquisition and culture of re-
search fish: Rainbow Trout, Fathead Minnows, Channel Catfish, and
Bluegills. Ecol. Res. Series No. EPA-660/3-75-011. U.S. Environ-
mental Protection Agency, Corvallis, Oregon. 54 p.
Burdick, G.E., H.J. Dean, and E.J. Harris. 1960. The effect of sevin
upon the aquatic environment. N.Y. Fish Game J. 7: 14.
Committee on Methods for Toxicity Tests with Aquatic Organisms. 1975.
Methods for acute toxicity tests with fish, macroinvertebrates, and
amphibians. Ecol. Res. Series No. EPA-660/3-75-009. U.S. Environ-
mental Protection Agency, Corvallis, Oregon. 61 p.
Craighead, F.C. 1923. Relation between mortality of trees attacked by
the Spruce Budworm and previous growth. J. Agric. Res. 30(6):
541-555.
Elson, P.F., A.L. Meister, J.W. Saunders, R.L. Saunders, J.B. Sprague, and
V. Zitko. 1973. Impact of chemical pollution on Atlantic Salmon in
North America, p. 83-110. In International Atlantic Salmon Sympo-
sium, 1973. International Atlantic Salmon Foundation, St. Andrews,
N.B. Canada.
Flannagan, J.F. 1973. Field and laboratory studies on the effect of expo-
sure to fenitrothion on freshwater aquatic invertebrates. Manit.
Entomol. 7: 15-25.
Haugen, G.N. 1976. Effects on Sevin-4-Oil and Dylox on small trout
streams of southwest Montana. U.S. Forest Service, Bozeman, Montana.
(Unpubl. Rep.).
26
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Litchfield, J.T., Jr., and F. Wilcoxon. 1949. A simplified method of
evaluating dose-effect experiments. J. Pharmacol. Exp. Ther. 96:
99-113.
Marancik, G. 1976. Monitoring of fish--1975 pesticide application for
Spruce Budworm control. U.S. Fish and Wildlife Service, Laconia, N.H.
36 p. (Unpubl. Rep.).
Marking, L.L. 1969. lexicological assays with fish. Bull. Wildl. Dis.
Assoc. 5: 291-294.
Marking, L.L., and V.K. Dawson. 1972. The half-life of biological
activity of antimycin determined by fish bioassay. Trans. Am. Fish.
Soc. 101: 100-105.
Marking, L.L., and V.K. Dawson. 19/5. Method for assessment of toxicity
or efficacy of mixtures of chemicals. U.S. Fish. Wildl. Serv.,
Invest. Fish Control No. 6/ (Circ. 185). 7 p.
Marking, L.L., and W.L. Mauck. 1975. Toxicity of paired mixtures of
candidate forest insecticides to Rainbow Trout. Bull. Environ.
Contam. Toxicol. 13: 518-523.
Mount, D.I., and W.A. Brungs. 1967. A simplified dosing apparatus for
fish toxicological studies. Water Res. 1: 21-29.
Schmiege, O.C., C.E. Crisp, R.L. Lyon, R.P- Miskus, R.B. Roberts and
P.J. Shea. 1970 Evaluation report on Zectran®. U.S. Forest Service,
Berkeley, Calif. 35 p. (Unpubl. Rep.).
U.S. Department of Agriculture. 1973a. Environmental statement on the
cooperative Douglas-Fir Tussock Moth pest management plan—Oregon and
Washington. U.S. Forest Serv. Service, Portland, Oregon. 223 p.
(Unpubl. Rep.).
U.S. Department of Agriculture. 19/3b. Final environmental statement on
the cooperative Gypsy Moth suppression and regulatory program. U.S.
Forest Service, Upper Darby, Pa. 257 p. (Unpubl. Rep.).
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27
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SECTION 4
PRINCIPLES AND METHODS OF BIOLOGICAL ESTABLISHMENT OF THE NORMS OF
CHEMICAL SUBSTANCES AND EVALUATION OF THE LEVEL OF POLLUTION
IN WATER-BODIES
V.I. Lukyanenko
Among the most important of the present problems is the question of
"clean water", i.e., the protection of waters from chemical pollution in
order to preserve the biological processes associated with a high quality
of water. This problem is extremely acute, complex, and enormous by its
scale. In this regard, it is sufficient to recall that from 450 to 700 km3
of waste water is discharged annually, the greater fraction of which under-
goes either no or only partial treatment. Reliance is placed jointly upon
the dilution of the waste waters with clean river waters and the process
of self-purification. However, to neutralize even 450 km3 of waste water,
provided at least a half of it undergoes treatment, a total of 6,000 km3,
or almost 40% of the so-called stable river discharge of the globe will be
required. For this reason, the prediction of the Institute of Geography
of the Academy of Sciences of the USSR seem to be quite real. This
Institute suggests that by the year 2000, all the water of rivers will
have to be used for neutralization of waste waters, even if the sewage be
treated by more perfect methods.
It is necessary to account for the fact that the pollution of waters
accelerates, but their self-purification capacity declines. Therefore,
one can not plan to increase the "toxicologic load" on rivers. On the con-
trary, the general way to solve the problem of "clean water" is to reduce
this load by building sewage treatment plants and raising their efficiency.
This will enable a decrease in the amount of clean water required for dilu-
tion of wastes, and provide for optimum functioning of the ecosystems re-
sponsible for "self-purification" of waters. Thus, exhaustion of water
resources in the nearest future can be avoided.
The urgency of this task of improving treatment of industrial wastes
and increasing the efficiency of sewage treatment plants is determined not
only by the high toxicity of many hundreds of chemical substances contained
in waste waters. The fact is that the volume of discharge from arable
lands containing various pesticides has greatly increased for the last two
decades. This discharge enters the same waters which receive domestic and
industrial wastes. Consequently, the "toxic" load on natural waters be-
comes greatly increased. The principal way of preventing pollution of
waters by toxic industrial wastes is the treatment of the sewage and limit-
28
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ation of its discharge into receiving waters. This way is not applicable
to the diffuse discharge from the agricultural lands. Here the method is
to determine the toxicity for aquatic organisms of each of the respective
toxicants, and prohibit or restrict the use of pesticides characterized by
a very high toxicity to fishes and aquatic invertebrates. The list of
pesticides used in modern agriculture is extensive, and the search for new
ones is so rapid that thousands of herbicides alone are patented every
year. It is not difficult to realize how labourous the task of preliminary
determination of toxicity of new pesticides is, especially if one accounts
for the present empirical approach of the water toxicologists to its solu-
tion.
Currently, aquatic toxicology does not enjoy the general theory of the
action of pesticides on a living cell and the organism as a whole. In
both domestic and foreign literature, there are still very few studies de-
voted to the mechanisms of action of toxicants upon the cell, subcell, and
molecullar structures. Without an understanding of these mechanisms, it
is impossible to comprehend the development of toxicologic processes
brought about by different groups of toxicants. Nevertheless, it is this
understanding of toxicologic processes that must be the basis for choosing
the methods of evaluation of the toxicity of a substance, or a group of
structurally related substances. This understanding is also essential for
the determination of the maximum permissable concentrations (MPC) of these
substances in natural waters. The accepted practice in the USSR
demonstrates that one of the most efficient means of providing protection
of waters from pollution is hygienic and fisheries standardization, i.e.,
the establishment of maximum permissible concentrations (MPC) for toxic
substances entering water bodies. It is not merely coincidence that both
the medical profession and biologists have arrived at this solution.
Ichthyologists, hydrobiologists, sanitary and hygiene medical personnel
face the same problem, i.e., assuring clean natural waters, preventing
such a degree of pollution as to cause poisoning of animals and human
beings, and alteration of the normal course of biological processes
determining productivity of the waters, and their "self-purification"
which render water drinkable. It is quite understandable that the degree
of toxicity of a substance for animals, fish, and aquatic invertebrates
can be established only under experimental conditions. Harmless
concentration of a given substance for a given organism may be found in
the same way. The sensitivity and resistance of various land animals to
toxic substances is not the same as for aquatic organisms, thus, the
sanitary-hygienic and fisheries requirements for quality of water related
to toxic substances are different.
Experimental data accumulated to date clearly show that the values of
the MPC of many substances for aquatic organisms, especially for fishes
(fisheries MPC), are lower, i.e., more "stringent" than for warmblooded
animals and humans (Sanitary-hygienic MPC). For example, the fisheries
MPC of copper (0.01 mg/1), nickel (0.01 mg/1) and zinc (0.01 mg/1) are
only one hundredth as high as the sanitary-hygienic MPC of the same metals
(1 mg/1). Similary, the toxicity of many organic substances, for fishes
and aquatic invertebrates (especially pesticides) is hundreds of times as
high as that of the warmblooded animals. The cause of these differences
29
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is quite evident. The warmblooded animals undergo a short-term contact
with the polluted water which enters them in relatively small portions,
while the water is the permanent home of the aquatic organism.
Does it mean that the criteria for toxicity and methods of estimation
of toxic effect of various chemical substances elaborated by general and
sanitary toxicology are not applicable to one of the classes of verte-
brates - fishes, and to aquatic invertebrates? Of course it does not.
The author (1973) has previously emphasized that toxicology of fish is a
part of comparative and general toxicology. In connection with this, many
ideas and methods may be used in solving the practical tasks of aquatic
toxicology associated with protection of waters from chemical pollution.
The last ten years of impetuous development of ichthyo-toxicological
investigations both in our country and abroad have yielded confirmation of
fruitfulness of this point of view. Today we can -take the next step on
the way to consolidation of the efforts of medical man and biologists in
solution of the problem of "clean water".
The unity of aims of both the sanitary-hygienic and fisheries per-
sonnel in establishment of standards for chemical substances discharged
into natural water (i.e., preservation of clean water in rivers and
reservoirs) conditions the necessity for creative analysis of the
principles of establishment of standards and elaboration of the universal
system of the MPC. This system must enable protection of water bodies
from ecological sanitary-hygienic and fisheries point of view. In
essence, from a biological approach. After all, the cleanliness of water
depends upon the biological processes of production and distruction; the
dynamic equilibrium which determines the high quality of "living" water.
Within the foundation of the biological establishment of standards for
the MPC of chemical substances must lie the main elements of sanitary-
hygienic and fisheries principals established at the present time. This
includes the evaluation of the effect of chemical substances on organo-
leptic properties (taste and smell) of water and aquatic organisms, the
sanitary condition of the water-body (processes of mineralization of
organic pollution), the toxic action of the incoming substances upon
aquatic organisms of different levels of organization, and the effect upon
laboratory animals which are used by medical personnel for determining
sanitary-toxicological harmful ness.
When expansion of the unit biological standard of the MPC is con-
sidered, the generally accepted methodical scheme of the sanitary-hygienic
setting establishment of standards of the MPC will not suffer any
essential change, since its validity is established. It is only necessary
to expand the genetic aspects of investigations, since many chemical
substances entering the waters with sewage are genetically active, i.e.,
capable of bringing about both mutations and modifications in concentra-
tions that are significantly lower than those established for the hygienic
MPC (Rapoport, 1972). Genetic activity of toxicants (induction of genetic
mutations; aberration of chromosomes) is manifested at such a level that
is is impossible to evaluate these changes with the common physiological
and biochemical tests. The main object of investigations of the mutagenic
30
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and morphogenic activity of toxicants must become, in the view of I.A.
Rapoport (19/2), more intensively studied by genetics, e.g., dsiot>opkila.
&p. being a likely choice since it possesses, like humans, the nucleo-
protein genom, but the number of genes is only one tenth - one twentieth
as great. There is good cause to agree with I.A. Rapoport (1972) when he
states, "genetic experiments on cbioAopklta. provide a unique possibility to
determine the ability of the chemical agents to induce mutation in genes,
injure chromosomes, as well as assess the influence of pollutants upon
moving apart the chromosomes, the latter being a very important parameter
of the genetic danger of chemical pollutants in the environment".
The scheme of the fisheries MFC must be reviewed and modernized in two
directions. First, it is necessary to pay more attention to the biologi-
cal aspects of setting of the standards of harmful substances entering
water-bodies, and to evaluation of the efficiency of this process. It
must be emphasized that here two different "ecological" aspects are
addressed: (1) the ecological foundations of the biological establishment
of the standards of the MPC, and (2) the ecological foundations of
evaluation of the efficiency of this process, directly on water-bodies.
The first of the two aspects has been considered by the author (1967)
in detail at an earlier date. On the basis of that experimental data, and
data from the literature, a number of propositions about the importance of
the role of ecological (abiotic) factors of aquatic environment for deter-
mining sensitivity and resistance of aquatic organisms to toxic agents
have been formulated. The propositions include the necessity of consider-
ing this dependence when the MPC is established. Physical and chemical
properties of the water medium influence the latent period, dynamics of
intoxication, and the threshold of resistance of fishes and other organisms
to poisons. In other words, the actual toxicity of some poisons (ions of
heavy metals, acids, alkalis, and organic poisons) may be either reduced
or intensified depending upon the environmental background.
There are two principal routes of influence of physical and chemical
parameters of the water medium upon the toxic-resistance of aquatic or-
ganisms: (1) direct, and (2) indirect. The first is a direct influence
upon the living organism, by changing the level of metabolism which leads
to an increase in the toxic-resistance. Changes in the normal regime of
functioning of different physiological systems, particularly the onset of
extreme conditions (high temperatures, rapid fluctuations in temperature,
oxygen deficiency, etc.) lead: (1) to easier penetration and accumula-
tion of toxicants in an organism; (2) to destruction or weakening of the
detoxification mechanisms and the processes of releasing of the toxicants
from the organism; (3) to increase in sensitivity of some functional
systems (target functions) to toxic substances; and (4) to a decrease in
resistance. Any combination of these changes, or an individual change
alone, can reduce the total resistance of the organism, and thus lead to a
greater toxic effect for a given chemical agent or combination of sub-
stances.
The second route of influence of the abiotic factors of environment
upon the resistance of the aquatic organism to poisons is the indirect
31
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mechanism, a factor which often decreases the actual toxicity of a sub-
stance. By this is meant a change in the toxicity of the substance owing
to decrease of its actual concentration in solution, or its physical and
chemical transformation. The decrease in the toxicity of many heavy metal
ions in hard water and in sea water due to formation of precipitates is
illustrative of this mechanism. A change in the toxicity of various
metals as a result of a complete or partial hydrolysis, formation of poorly
soluble carbonates, and precipitation from solutions having the pH value
far from neutral is also well known.
Experimental data on the dependence of the degree of toxicity of
various substances and resistance of aquatic organisms (mostly fish) upon
ecological factors are long known, although these data are still not used
in establishing the MPC. However, a skillful use of existing information
when determining the MPC of a given substance would allow not only a
reduction of the duration of experiments by conducting them under extreme
conditions (high temperatures, low oxygen content, etc.), but would also
invest the proposed MPC with an "ecological factor of safety", i.e., with
due consideration of the range of fluctuations of physical and chemical
factors of the environment.
Such an "ecological MPC" would guarantee the relative well-being of
the ecosystem as a whole, as well as its separate components. It would
protect the components from poisoning even under conditions of deteriora-
tion of the main abiotic factors which usually leads to a decrease in the
toxic-resistance of aquatic organisms.
Another aspect of the so-called ecological setting of standards for
harmful substances in water became an object of special discussions only
recently. By this is meant the so-called ecological MPC designed to
secure cleanness or "health" of a water body as a whole, i.e., preserva-
tion of natural ecosystem of the water-body and not only important com-
mercial organisms. However, the attempts made to clarify this concept
lead to the conclusion that one must speak not of "ecological MPC", but of
ecological foundations of establishing the MPC as noted by M.M. Kamshilov,
"Determination of concentrations of foreign substances not disturbing
natural biological circulation in aquatic ecosystem". The fisheries MPC
are, in essence the "ecological MPC", since they must secure protection
from toxicants of not only fish, but also the ecosystem as a whole.
It is quite a different matter, when we speak of search and standardi-
zation of the indices of "ecological well being" which are very important
for the estimation of the efficiency of biological standard setting rela-
tive to toxicants discharged into waters. The investigation of the
polluted water bodies of the effect of domestic and industrial wastes on
the ecosystems of these water-bodies has been performed by sanitary hydro-
biology. Aquatic toxicology, including ichthyo-toxicology was born as an
outgrowth of sanitary hydrobiology.
The development of these disciplines became possible as a result of
the realization of the fact that no investigation and no description of
the changes occurring in the communities residing in polluted waters,
32
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regardless of how thorough they might be, was able to address the issues
of which components were involved, and what degree of removal would be
required. These questions could only be answered with a help of experi-
mental methods of investigation and establishment of the MPC.
The impetuous development of aquatic toxicology and ichthyo-toxicology
during the last two decades has lead to a notable decline in sanitary-
hydrobiological investigations. However, the main object of these
investigations remains the water-bodies polluted with organic matter.
Herein lies one of the reasons behind the development of the interest
of some toxicologists in ecological aspects of aquatic toxicology which
must actually be dealt with by sanitary hydrobiology. A distinct demarka-
tion of the tasks and methods of sanitary hydrobiology and aquatic toxi-
cology is needed not only for successful solution of specific problems
faced by each of the sciences, but also for establishment of fruitful
contacts when solving the problems of protection of waters from pollution.
It is to be emphasized that the evaluation of the effectiveness of biologi-
cal setting of standards for substances discharged into waters can be exer-
cised only from the criteria and methodology of sanitary hydrobiology,
which has as its object of study the water-body as a whole and its living
communities.
In this regard, the indices of ecological health imposed by M.M.
Kamshilov, e.g., the oxygen concentration in the water; ratio of produc-
tion to destruction; character of benthic communities; and the distribu-
tion of indicator species; deserve deep attention. The organoleptic
symptom of harmful ness must also be included. This symptom implies the
influence of pollution on organoleptic properties of not only the water,
but also of the aquatic organisms, including fish. Unfortunately, this
aspect of fisheries investigation attracts scant attention in comparison
with the sanitary-hygienic setting of standards. It is enough to recall
that out of 420 established sanitary-hygienic MPC standards of harmful sub-
stances, more than a half (216) are limited by the organoleptic index, 147
by sanitary-toxicologic considerations, and 57 substances by the general
sanitary index of harmfulness. But within the fisheries MPC, only 15
substances of a total of /O are limited by the organoleptic index. The
main reason for the poor use of the organoleptic index in the fisheries
MPC is an insufficient elaboration of the methods of objectively evaluating
the reaction of aquatic organisms, including fish, to changes in taste and
smell of the water caused by toxic substances. The methods of investiga-
tion of these reactions exist, in the form of conditioned reflexes which
still awaits wide application in ichthyological investigations.
Nevertheless, the sensitivity of fish to the odors of many chemical
substances greatly exceeds that of a human being. Thus, Hasler and Wisby
(1951) discovered in fish the ability to detect phenol in concentrations
0.01 mg/1, or even 0.005 mg/1, using a conditioned reflex. These concen-
trations are considerably lower than the threshold for humans. These data
agree with the results of Neurath (1949) who reported that fish detect the
the smell of phenethyl alcohol at the concentrations 250 times lower than
humans. An even greater sensitivity of the eel to beta-phenethyl alcohol
33
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was demonstrated by Teichmann (1957). The eel showed a reaction to this
substance at concentrations as low as 3 x 10-20 mg/1, i.e., when only 2-3
molecules of the substance could be present in the olfactory bulb.
The fish discern perfectly well the smells of many aquatic plants (Walker
and Hasler, 1949), as well as the smells of fish and other vertebrates
(Von Frisch, 1941; Sliultz, 1956; Walker and Hasler, 1949). In most cases
the sensitivity of fish to the smells of substances excreted by closely
related species is greater than to those of taxonomically remote species.
The repellent effect on fish may be produced by substances excreted from
the skin of other classes of vertebrates. Thus, buffotoxin extracted from
the skin of adult toads is preceived by fish, and produces a repellent
effect even at a dilution of l:2.4x!06. These are but a few of the
reported studies noting the extremely high sensitivity of fish to changes
in smell and taste of water which they inhabit.
Such a high sensitivity of fish to the organoleptic properties of
water can not but affect the distribution of fish in a water-body. It is
not difficult to imagine that tens and hundreds of substances, mostly of
an organic nature, entering natural waters with sewage may produce a re-
pellent or an attracting effect on fish changing feeding, wintering, and
spawning conditions; causing unnaturally high concentrations of fish in a
limited area, driving fish away from food, and thus making it difficult to
use the nutrient base, and reducing the bio-productivity of a whole water-
body. All of these complicated and intricate manifestations of the "ecolo-
gical ill-health" of a water body receiving organic substances, even in
strict conformity with the established standards, may be properly con-
sidered only on the basis of the knowledge of the reactions of avoidance of
the poisons, which change taste and smell of the water and food organisms.
Therefore, one of the main tasks in the field of experimental aquatic toxi-
cology is the thorough study of behavioral reactions. This study must in-
clude the reactions of detection and avoidance of chemical agents, the
study of mechanisms and the character of these reactions (repellent or
attracting), and the resolving power of the olfactory and gustatory organs
in fish. In summary, the idea is to establish physiological foundations
for a wide application in biological standards which have been well esta-
blished in the sanitary-hygienic standards related to chemical pollution
of water-bodies. At the same time, the study of avoidance reactions in
fish will enable an understanding of the peculiarities of distribution of
fish in water-bodies which are "loaded" with waste waters in accordance
with the biological standards. It should be noted that the distribution
of indicator organisms, fish included, may serve as one of the indices of
"ecological health" of a water-body. Thus, reference is made to the
"ecological establishment of the standards" or, to be more correct, to the
ecological principles of evaluation of the efficiency of biological
setting of standards.
So, the ecological principle is very important for the evaluation of
the effectiveness of biological setting of standards for substances
entering waters. Even the most thorough observations and detailed descrip-
tions of the changes in aquatic ecosystems are not able to reveal harmless
concentrations of chemical substances discharged into a water-body, or to
establish what substance or a group of substances cause the noticed changes
34
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in the ecosystems. Similarly, the study of the most important physiologi-
cal and biochemical parameters of various aquatic organisms dwelling in a
polluted water-body and the detection of essential disturbances in func-
tioning of living systems do not reveal harmless concentrations of chemi-
cal agents. Therefore, the main task of the biological setting of stan-
dards, i.e., establishment of the MPC, may be solved only under experi-
mental conditions on the most sensitive test-objects or representative or-
ganisms of any natural ecosystem. Of course, such a scheme of experiments,
i.e., the separate use of the most sensitive elements of the ecosystems,
leads to certain simplifications of the real situation in the water-body.
It would be, however, a naive assumption to expect that these simplifica-
tions might be avoided by experimenting in ponds or canals rather than in
aquaria, or complex "natural ecosystem" as opposed to individually sensi-
tive test objects. This delusion comes from confusing the main tasks of
the principles and methods of sanitary hydrobiology and aquatic toxicology.
One should not be embarrassed by the experimentally unavoidable "simplifi-
cations" of a real situation, just as medical science is not embarrassed
when sanitary-hygienic standards for chemical substance are established.
The sanitary-hygienic MFCs are meant to secure the safety of human beings,
but they are established in experiments using small rodents or other
larger mammals (rabbit, dog).
Similarly, the genetic investigations designed to determine the MPC
will most likely be performed on dsioAopkJJa 4p., a classical test animal
in genetic investigations, having a number of advantages over laboratory
mammals. In this regard, the position of aquatic toxicologists is easier
since the possibility of studying the MPC directly exists. This direct
application also allows the selection of the most sensitive species. The
use of the most sensitive and least resistance components of natural
ecosystems, namely fish and aquatic invertebrates, makes the experimenting
on more complex ecosystems for the establishment of the MPC unnecessary.
The concept of "natural ecosystems" itself has no single meaning, and
will be essentially different for each experimental water body, to say
nothing of those natural waters which serve as "receivers" of waste
waters. Figuratively speaking, "the natural ecosystem" of the experi-
mental small pond differs from "the natural ecosystem" of a reservoir or a
lake to much greater extent than do small laboratory rodents from man.
Consequently the appeals to change the experiments on test-objects in
aquaria for experiments on "natural ecosystems" in small ponds are lacking
serious scientific foundations, and do not take into consideration the
needs of today's life, i.e., to establish in the shortest time the MTCs of
hundreds of substances entering waters in connection with the appearance
of new branches of industry, modernization of technological processes and
the advances of chemistry in agriculture.
It is well known that the metazoans, which constitute the basis of
grazer circulation, are more sensitive to toxicants than are the unicell-
ular organisms. Among the metazoans, the vertebrates are more sensitive
to various toxicants, specifically, organic compounds, than are the
invertebrate forms. This fact served as a basis for wide use of various
species of fish when establishing the MPC values with the help of the
35
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method of the so-called fish-test both in the USSR and abroad, including
the USA. It should be noted that the time factor, i.e., determination of
the toxicity of a substance in a shortest possible time, becomes decisive
today. Therefore, the problem of rapid establishment of biological MPC of
chemical substances is the number one problem in both the scientific and
the commercial considerations, since the number of toxic substances
discharged into the waters grows at a frightening rate.
The approaches of Soviet and American water toxicologists to solution
of this problem differ primarily in the importance attached to long-term
(chronic) and short-term (acute) experiments. In the USA there is a
method of estimation of the conventional harmless concentration of a toxi-
cant assumed to be 0.1 x TL (tolerance limit). The TL value is found in
short-term experiments (exposure of 24-96 hours). Such an approach
enables quick answers to the question of the degree of toxicity of a sub-
stance and its conventional harmless concentration. An essential disad-
vantage of this method is that the toxicity of many substances is dis-
played in prolonged (chronic) experiments at the concentrations much lower
than 0.1 x TL.
The method of estimation of the MPC in the Soviet Union allows acquisi-
tion of more confident data about the toxicity of substances owing to pro-
longed observations of the survival of various aquatic organisms (from
fish to microbes), but is is very laborous and time consuming. This makes
it necessary to find a reliable, but time saving method of determination
of MPC. The main direction of the search is the experimental elaboration
of the transition from acutely toxic and threshold concentrations, to
maximum permissible ones. Here cooperative Soviet-American investigations
are needed with the application of the methods of physiological, biochemi-
cal and biophysical analysis for the quickest possible detection of the
symptoms of intoxication of varying aquatic organisms.
The physiological and biochemical foundations of the determination of
the MPC (Lukyanenko, 1965, 1967, and 1973) allowed the development of the
method of physiological/biochemical indicators, the resolving capacity of
which is tens of times higher than that of the method of the "fish-test".
The method of the indicators allows detection of toxic effects of a sub-
stance by an understanding of the condition of one or another functional
system of the organism. This can be done in a much shorter time with the
indicator experiments than with the "fish-test", since the disturbances in
functions are observed long before the lethal effect. The choice of the
method of determination of the MPC of the investigated substance must be
based on the knowledge of toxicological dynamics of this substance and the
mechanism of its action, i.e., a clear idea of the most susceptible func-
tion or "target function" (Lukyanenko, 1973).
Determination of the MPC of chemical substances entering water-bodies
is an important function, but not the only task of aquatic toxicology.
The diversity and complexity of the problems faced by aquatic toxicology
and ichthyo-toxicology implies a wide application of many modern methods,
primarily physiological and biochemical evaluation of the toxicity of in-
vestigated substances.
36
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Thus, it is necessary to establish and accept a unified scheme for
conditioning ichthyo-toxicological investigations in order to obtain com-
parable data on the toxicity of the pollutants of water. It is no secret
that the Mont Blanc of experimental data on toxicity accumulated in the
world literature is of little value. The reason for this is simple: data
from different authors are hardly comparable because of the absence of
standardization in performance of the experiment and because of the lack
of uniform expression of the results. Today when the agenda is inter-
national cooperation in the field of aquatic toxicology and ichthyo-toxi-
cology, the absence of unified standard scheme for conducting toxicity
experiments and determining MFC values is especially distressing.
Therefore, it is necessary to consider again the standard scheme of
ichthyo-toxicological investigations (Lukyanenko, 1967, 1968) which
includes acute, subchronic and chronic experiments. The acute experiment
is performed for preliminary evaluation of the degree of toxicity of a
substance using the "fish-test" method. The indicator of toxicity is the
death of the experimental fish. It is reasonable to conduct the experi-
ment at relatively high temperatures, and relatively low oxygen content,
taking into account the range of fluctuations in natural water body.
Water hardness and pH value are selected in such a manner as to show the
maximum toxicity of the substance tested. The least resistant fish
species of ichthyofauna tested should be used as the test object. In this
case, it is important to keep in mind the characteristic of resistance for
the given species at various stages of ontogenesis, selecting the least
resistant from them. Taking into consideration the relation of lethal
effect and test duration, the duration of the acute test should be limited
to 24 hours.
A subacute test is carried out to show the path of the toxicant's
effect on fish and function development mechanisms, with the most sensi-
tive methods for determining the threshold concentration demonstrated in
the chronic test. Concentrations of the substance which possess an ex-
pressed toxicity effect on the organism are used. They are usually found
within a range of 1/2-1/10 of the lethal concentration. The test is con-
ducted on the least reistant species of fish. Duration of the tests is
from 10 days to 1 month. Since the toxicodynamics of the majority of
substances discharged into water bodies is unknown (applicable to fish),
the most complete set of indicators possible which allow us to evaluate
the functional condition of the organism's various systems is necessary.
Together with indicators integrally reflecting the organism's condition,
such as increase in live weight, level of feeding excitability, and
intensity of oxygen consumption, finer (or more delicate) physiological
and biochemical indicators (activity of various enzymes, hemotological
indicators, humoral and cell factors of inborn immunity, behavioral re-
actions, and electro-physiological tests, which characterize the condition
of various functional systems should be used.
The chronic test - is the final stage of ichthyotoxicological research.
Its task is to demonstrate the threshold concentration, toxic effect zone
and maximum inactive concentration. It is expedient to test 3-5 concentra-
tions with a five-fold interval. The initial concentration and range of
37
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concentrations are selected from data obtained in the acute and subacute
tests. The duration of the chronic test is from 1 to 3 months, but no
longer. To shorten the duration of the tests, it is expedient to use the
functional loading method, and on this basis to determine the condition of
the indicator function (selectivity imitated), established on the basis of
data from the subacute test.
Special emphasis was given to ichthyotoxicological tests when setting
biological standards for MPC, since the toxicological indication of harm
is more dangerous than all components of the natural ecosystem, and conse-
quently, for the "health" of the water body. However, this does not mean
that in biological standard setting, it is always necessary to be ruled
solely by this sign. That principle is composed when actually setting hy-
gienic standards, and with whose agreement the MPC of the substance tested
is established according to the limiting factor of harm (general sanitary,
organoleptic or toxicological). For example, the standard may be estab-
lished according to the least concentration of the substance which demon-
strates an unfavorable influence on the water body. This criteria must be
preserved to the end that setting biological standards be fully self
justifying.
Considerations of principles and methods for biological standard
setting for chemical substances stated in this report, and evaluations of
the level of pollution in water bodies require further development.
REFERENCES
Frisch, K. 1941. von. Z. vergl. Physiol. B. 29.
Hasler, A.D. and W.I. Wisby. 1951. Trans. Am. Fish. Soc. V. 79.
Lukiyanenko, V.I. 1965. In Coll.: Problems of Hydrobiology M.
Lukiyanenko, V.I. 1967. Toxicology of fishes M.
Lukiyanenko, V.I. 1968. Tez. dokl. na vsescuzn. nauchn. Konf. po vopr.
vodn. toksikol. M.
Lukiyanenko, V.I. 1973. In. Coll.: Experimental aquatic toxicology.
Riga, issue 4.
Neurath, H. 1949. Z. vergl. Physiol. B. 31.
Rapoport, I.A. 1972. In Coll.: Scientific foundations of establishing
MTC in water medium and self-purification of surface waters M.
Shultz, F. 1956. Z. vergl. Physiol. B. 38.
Teichmann, H. 1957. Naturwissenschaften. B. 44.
Walker, T.J. and A.D. Hasler. 1949. Physiol., zool. V. 22.
38
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SECTION 5
CHRONIC EFFECTS OF LOW LEVELS OF HYDROGEN CYANIDE
ON FRESHWATER FISH
Lloyd L. Smith, Jr.
INTRODUCTION
The toxicity of cyanides to fish has long been recognized. Compounds
containing the cyanide radical are frequently present in effluents of
many industries including electroplating plants, steel mills, petroleum
refineries and gas works. In aqueous solution the cyanide radical from
simple alkali cyanides such as NaCN hydrolyzes to form free cyanide (CN
ion and molecular HCN). The molecular (un-ionized) component predominates
at pH values found in most natural waters, with less than 4 percent of
free cyanide occurring in the ionic form below pH 8 at 25 °C. As the pH
of aqueous simple cyanide solutions is increased, the percentage of free
cyanide present as the CN~ ion is increased to satisfy the equilibrium
reaction of HCN * H+ + CN~.
Little information is available on the long-term effects of hydrogen
cyanide on fish. Neil (1957) and Broderius (1970) found that free cya-
nide concentrations of 10 yg/liter, expressed as CN, impaired the swim-
ming performance of salmonid fishes. Leduc (1966) measured the growth of
juvenile cichlids ( C-ccA&uoma bxjna.ca£atan1).
The work reported here was designed to determine the effect of low
levels of HCN on the fathead minnow, V-umphal^> piome&u , from egg
through the juvenile period of the second generation, and on brook trout,
Sa£veJttno6 fiontznatlb, adults through egg maturation to spawning and
development of the second generation to advanced juvenile stage.
MATERIALS AND METHODS
Experiments with fathead minnows were started with eggs from labora-
tory stock originating at the Duluth laboratory of the U.S. Environmental
Protection Agency. The brook trout adults utilized were from the State
of Wisconsin hatchery at Osceola, Wisconsin. Water for experiments was
taken from the laboratory well (Table 1).
Eighty fathead minnow larvae were placed in each of 15 20-liter glass
tanks, and sodium cyanide solution was introduced to the chambers with a
39
-------
TABLE 1. ANALYSIS OF WELL WATER USED IN BIOASSAYS*.
Concn
Item (mg/liter)
Total hardness as CaCO 3 220
Calcium as CaCO3 140
Magnesium as CaCO3 70
Iron 0.02
Manganese 0.04
Chloride <1.0
Sulfate <5
Fluoride 0.22
Total phosphorus 0.03
Sodium 6
Potassium 2
Ammonia nitrogen 0.20
Organic nitrogen 0.20
Phenols <0.005
Cu, Cd, Zn, Ni, Pb, Hg <0.01
aWater taken from well head after iron removal and before
aeration and heating.
40
-------
toxicant dispensing system designed by Mount and Warner (1965). This
system cycled every 3 min to deliver 1 liter of water and a measured
amount of toxicant. Twelve levels of HCN were maintained from 5 to 100
lag/liter at pH 8.06-8.09, 24.8-25.1 °C and dissolved oxygen of 5.9-6.3
mg/liter. Cyanide concentrations were analyzed by the Epstein coloro-
metric method (APHA, 1971) from samples taken in the test chambers 3
times per week. HCN was calculated from dissociation constants of Izatt,
&t &£., (1962). Three controls with well water were run simultaneously
with the cyanide treatments.
After 106 days minnows were transferred to 20 treatment and 5 control
chambers, each containing 35 liters of water and 20 fish. In 149 days or
when first spawning occurred, four mature males and three females were
selected and left in each treatment tank.
After 192 days total exposure of fry and spawning adults, 50 eggs
from each treatment were placed in plastic cylinders with a screen on one
end. Cylinders were suspended in the same test water as adults and os-
cillated until the eggs hatched. After 227 days from the start of expo-
sure of the parent generation, a growth and survival experiment was
started with second-generation fry. Length was determined photographi-
cally after 28 days of fry exposure, and length and weight were measured
directly at termination after 56 days.
Ten 19-month-old adult brook trout were placed in 340 liter tanks
with eight cyanide treatments and two controls. Treatments ranged from
5.7-75.3 yg/liter HCN at 7.95 pH, 9.0-15 °C and 6.5-7.9 mg/liter Q2- Tne
HCN metering apparatus utilized was the same as for fathead minnows.
Temperature, pH and cyanide analyses were made 3 times per week. Exposure
began on May 5 and continued through spawning 196 days later. After 143
days, exposure spawning boxes were placed in the tanks and the number of
fish in each tank reduced to two males and four females. Spawned eggs
were removed from each box each day.
From each spawning of 100 viable eggs, 50 were randomly selected and
placed in oscillating cups to hatch at 9 °C. Viability was determined at
12 days by development of the neural keel. Twenty-one days after hatch,
25 alevins (larvae) 15-19 mm long, depending on previous treatment, were
placed in 20-liter glass tanks where they were held at 9 °C. There were
3 replications in each of 8 treatments ranging from 5.6 to 77.5 yg/liter
HCN and 3 controls. Fish were measured photographically at the end of
each 30-day period, and at 90 days they were weighed. After 60 days,
fish numbers were reduced to 20. Alevins were fed an unrestricted diet
of dry trout food in pelletized form.
RESULTS
Fathead - First Generation
After 28 days, survival in the first-generation experiment averaged
64 percent in the 3 controls and ranged from 80 to 11 percent in treat-
41
-------
ments. Between 28 and 56 days, survival was 80 percent or greater in all
chambers. Survival was 95 percent or greater in all chambers between 56
and 84 days (Table 2). Mortality rate in these two periods was not
significantly correlated with HCN concentration at the 0.05 level. After
56 days exposure, the mean weight of fish in the treatments ranged from
130 to 65 percent of that in controls, and from 128 to 86 percent of mean
weight in controls after 84 days. The length of time from the start of
HCN exposure to the onset of spawning averaged 156 days in controls, and
ranged from 148 to 206 days in treatment chambers. In no treatments did
spawning begin significantly earlier or later than in the controls.
Mean egg production and egg production per ferna^ in each chamber are
shown in Table 3. Egg production per female in treatments ranged from 72
percent of that in controls to zero (Table 3). Egg production per female
was significantly reduced relative to controls i-n HCN treatments of 19.6
yg/liter and higher. Mean percentage hatch of eggs spawned and incubated
in the tests (Table 3) ranged from 83.9 percent in controls to 21.6 per-
cent at 63.6 yg/liter and 0 percent at 80.7 yg/liter (Figure 1). When
3000
FATHEAD MINNOW, FIRST GENERATION
oc 1000
CO
10 20 30 40 50 60
HCN CONCENTRATION, jug/I
Figure 1. Number of fry produced per 100 grams of female fathead minnow
in HCN. Egg production times percentage hatch.
42
-------
TABLE 2. SURVIVAL AND WEIGHT OF FIRST-GENERATION FATHEAD MINNOW
AFTER 28 DAYS, 56 DAYS, AND 84 DAYS OF EXPOSURE TO HYDROGEN CYANIDE*.
Treat-
ment
1
2
3
4
5
6
8
9
10
11
12
Mean HCN**
Concentration
( g/D
Control
Control
Control
5.9
11.4
17.9
24.7
32.8
40.5
57.5
66.8
75.3
88.9
98.1
Percentage Survival***
0-28
Days
64
71
58
80
59
60
51
46
59
49
49
29
19
11
28-56
Days
94
100
100
100
100
98
100
100
100
97
87
91
80
100
56-84
Days
100
100
100
98
100
95
100
97
100
97
100
100
100
100
Mean Weight
(g)
56
Days
0.292
0.236
0.354
0.205
0.274
0.270
0.296
0.382
0.272
0.190
0.220
0.264
0.199
0.190
84
Days
0.581
0.588
0.668
0.580
0.676
0.631
0.688
0.785
0.555
0.458
0.528
0.629
0.560
0.627
*Chambers originally contained 80 larvae. Numbers were reduced to a
maximum of 40 per chamber after 56-day measurements.
**84-day period.
***Survival of fish present at beginning of period.
43
-------
TABLE 3. EGG PRODUCTION, EGG SURVIVAL AND TERMINAL WEIGHTS OF
FIRST-GENERATION ADULT FATHEAD MINNOWS EXPOSED TO VARIOUS
CONCENTRATIONS OF CYANIDE.
Treat-
ment
HCN
(yg/D
Control
5.8
12.9
19.6
27.3
35.8
44.2
63.6
72.8
80.7
100.7
No.
Tests
5
2
2
2
2
2
2
2
2
2
Mean
Eggs
Per
Female
3,476
2,512
1,845
1,468
1,367
1,010
1,119
72
319
242
0
Mean
Percentage
Hatch
83.9
61.6
81.3
56.4
39.3
50.6
12.8
21.6
19.6
0.0
0.0
Mean
Fry/ 100
Grams
Female
2,916
1,547
1,50,0
828
537
511
143
16
62
0
0
Mean Weight
Surviving Adults
(g)
Male
4.42
4.50
4.60
4.31
4.89
4.03
5.00
4.42
4.21
4.49
0.0
Female
2.45
2.60
2.11
2.33
2.06
2.50
2.26
2.63
2.19
2.23
2.21
44
-------
the parent experiment was terminated, weights of survivors of either sex
in the treatment chambers did not differ significantly from weights of
control fish (Table 3).
Fathead FI Generation
Survival of fry in the Fj generation after 28 days was generally
higher than in the parent experiment. Survival averaged 84 percent in
the three controls and ranged from 88 percent at 26.3 yg/liter HCN to 36
percent at 81.0 yg/liter HCN (Table 4). Mortality rate and HCN concentra-
tion were not significant correlated at the 0.05 level. Over the period
of 28 to 56 days, all chambers had 81 percent survival or higher. Mean
length of fish in treatments after 28 days ranged from 116 to 64 percent
of that in controls. The fish at 26.3 yg/liter were significantly longer
than control fish, and those exposed to treatment of 34.8 yg/liter and
greater were significantly shorter than control fish.
Mean total lengths of fish in treatments after 56 days ranged from
105 to 81 percent of mean length in controls, and mean weights ranged
from 122 to 52 percent of that in controls (Table 4). Weights and lengths
of fish in treatments from 5.7 to 52.2 yg/liter were not significantly
different than controls. Mean lengths and weights of fish at 61.6, 70.5,
95.9 and 105.8 yg/liter were significantly different than controls.
Brook Trout - Adults
Survival and Growth—
Adult brook trout were subjected to various levels of HCN for 196
days and showed no significant mortality or growth differences (p>0.05)
associated with cyanide concentration. In treatments of 53.9, 64.9 and
75.3 yg/liter HCN, one fish died in each after temperature was reduced to
9 °C. In the two highest treatments fish showed increased irritability
when temperature was reduced.
Spawning and Egg Production-
Spawning started in controls and at 5.7 yg/liter HCN 145-147 days
after treatment started, but at higher treatments, spawning did not start
until 156-159 days after treatment began (Table 5). The number of eggs
deposited per 100 grams of female varied from 357 in one control to 106
at 75.3 yg/liter HCN (Figure 2). The number of fertilized eggs per 100
grams of female varied from 293 in one control to none at 64.9 yg/liter
HCN. The percentage of live eggs 12 days after hatching varied from 93.6
percent of fertilized eggs in the control to 64.1 percent at 53.9 yg/liter
HCN and 0 percent at 64.9 yg/liter. Sperm mobility was tested at 11 HCN
concentrations but no significant relationship (p>0.05) with HCN was
noted.
Egg Survival and Hatch-
Eggs were incubated and alevins held for 90 days at HCN concentrations
of 5.6 to 77.2 yg/liter at 9 °C and 64-90 percent saturation of oxygen.
No significant differences from controls in percentage hatch was observed
(Table 6). Survival of alevins for 90 days after hatching were not signi-
45
-------
TABLE 4. SURVIVAL, LENGTH, AND WEIGHT OF FATHEAD MINNOW AFTER 28 DAYS
AND 56 DAYS OF EXPOSURE TO HYDROGEN CYANIDE!.
Treatment
Control A3
Control B3
Control C
1
2
3
44
5
63
73
83
9
105
II3
123
HCN
Cone.
(yg/l)
5.7
12.2
20.5
26.3
34.8
43.0
52.2
61.6
70.5
81.0
95.9
105.8
Percentage
Survival^
0-28
Days
85
83
84
80
39
60
88
48
64
58
64
56
36
64
40
28-56
Days
100
98
95
100
94
100
95
100
100
98
100
95
95
100
81
Mean Total Length
(mm)
28 Days
15.0
14.2
13.6
14.5
14.7
13.9
16.6*
12.5*
12.0*
10.8*
11.7*
10.5*
10.0*
10.4*
9.2*
56 Days
29.2
28.7
26.7
28.7
28.0
28.6
29.6
27.6
27.1
26.8
26.7*
25.9*
24.8
24.8*
22.9*
Mean Weight
After
56 Days
(g)
0.238
0.238
0.205
0.234
0.258
0.221
0.277
0.184
0.212
0.199
0.172*
0.160*
0.163
0.150*
0.188*
^Except where noted, chambers originally contained 80 larvae. Numbers were
reduced to a maximum of 40 per chamber after 28-day measurements.
n
Survival of fish present at beginning of period.
3Larvae spawned and hatched in control chambers of parent experiment.
Began with 42 larvae.
Began with 59 larvae.
6Values are significantly different from control values according to
Dunnett's procedure (two-tailed; a = 0.05).
46
-------
TABLE 5. EGG PRODUCTION OF ADULT BROOK TROUT EXPOSED TO HCN
144 DAYS BEFORE START OF SPAWNING.
Control
Control
5.7
11.2
32.3
43.6
53.9
64.9
75.3
Start of
Spawning
147
145
145
159
157
156
157
N
N
Sex
Ratio
M/F
2/4
2/4
2/4
2/4
2/4
2/3
2/3
2/3
2/2
Total Eggs/100 g
of Female
Deposited
237
357
299
162
208
189
187
122
106
Fertilized
128
293
246
137
119
171
124
0
0
PercentageJ-
Viable
(12 Days)
93.6
93.4
89.9
78.1
72.9
86.6
64.1
0
0
1
Formation of neural keel.
47
-------
LLJ
LU
LL
DC
280
240
200
8 160
C/D
8 120
LLJ
Q
S 80
* 40
LU
0
BROOK TROUT
0 10 20 30 40 50 60 70
HCN CONCENTRATION, jig/I
Figure 2. Number of fertilized eggs produced per female brook trout
in HCN.
48
-------
TABLE 6. PERCENTAGE SURVIVAL OF BROOK TROUT EGGS AND ALEVINS
EXPOSED TO VARIOUS LEVELS OF HCN.
HCN
Cone.
(yg/D
Control
5.6
11.3
21.8
33.3
43.5
55.3
67.2
77.2
Mean
No.
Tests
5
4
3
2
3
3
3
2
3
Hatch
Percent-
age
72.5
46.5
70.8
72.0
70.0
72.9
81.3
86.9
76.7
Survival
(90
No.
Tests
3
2
2
2
2
2
2
2
2
of Alevins
Days
Percent-
age
98.6
100.0
100.0
94.0
100.0
100.0
84.0
73.5
30.0
49
-------
ficantly affected at HCN levels of 43.5 yg/liter and lower. At concentra-
tions of 55.3 to 77.2 yg/liter survival was significantly less than con-
trols (p<0.05). At 77.2 yg/liter survival from hatch was 30.0 percent
compared to 98.6-100 percent in controls.
Growth of Brook Trout Alevins—
Length and weight of alevins at hatch was not significantly different
at the various HCN concentrations, but growth over a 90-day period was
markedly affected by increased HCN concentrations (Table 7). The effect
on growth was noted after 30 days and was great at 90 days. Fish at all
treatment levels from 33.3 to 77.2 yg/liter were significantly shorter
and lighter than controls (Figure 3). At 90 days length varied from
41.9 mm in controls to 24.3 mm at 77.2 yg/liter. Weight varied from
1.03 g in controls to 0.16 g at 77.2 yg/liter. At the highest concentra-
tion weight was 15.7 percent of controls. At 11.3 yg/liter growth after
90 days was significantly greater than the control, but at 5.6 and
21.8 yg/liter no significant difference from controls was noted. All
fish at treatment levels from 33.3 to 77.2 yg/liter at 90 days were
significantly slower in growth than controls.
10
20 30 40 50 60
HCN CONCENTRATION, M9/I
70
Figure 3. Growth of brook trout in various concentrations of HCN.
Expressed as percentage of controls.
50
-------
TABLE 7. GROWTH OF BROOK TROUT ALEVINS FROM HATCH TO 90 DAYS EXPOSED
TO VARIOUS LEVELS OF HCN.
Mean Length (mm)
HCN
(yg/D
Control
5.6
11.3
21.8
33.3
43.5
55.3
67.2
77.2
No.
Tests
3
2
2
2
2
2
2
2
2
At
Hatch
13.3
13.6
13.4
12.9
12.2
13.1
12.3
11.8
11.9
90
Days
41.1
41.9
43.3*
42.0
37.3*
36.0*
32.6*
30.2*
24.3*
Mean Weight
At
Hatch
40.0
40.8
40.3
40.0
38.0
40.0
39.7
41.0
39.7
At
90 Days
1.030
1.087
1.135*
1.020
0.722*
0.660*
0.469*
0.335*
0.162*
(g)
% of
Control
100.0
105.0
110.0
99.0
70.1
64.1
45.5
32.5
15.7
*Significantly different than controls.
51
-------
DISCUSSION
Most mortality among fathead minnows occurred in the first 28-day ex-
posure to HCN in both parent and F-j generations (Table 4). The number of
eggs produced and fry which survived were reduced at 196 yg/liter and at
higher concentrations. It is estimated that the highest "no-effect"
level of HCN is between 12.y and 19.6 yg/liter based on egg production.
The lethal threshold for juvenile fathead minnows (defined as the HCN
concentration at which no fish die for 48 hours after continuous exposure
for 96 hours or longer) as determined by unpublished data from our labora-
tory is 119 yg/liter HCN at 25 °C, pH 8.0 and 6.0 rug/liter DO (Table 8).
Comparison of this acute toxic level to the "no-effect" level indicates
that the safe level for fish is between 11 and 16 percent of the acute
toxicity concentration of HCN.
When adult brook trout, prior to spawning, were exposed to HCN, at
all concentrations greater than 5.7 yg/liter, a reduction in the produc-
tion of fertilized eggs occurred. When spawning was successful, egg
viability was not affected adversely at 43.6 yg/liter and lower. Growth
rate of juvenile brook trout during the first 90 days after hatching was
reduced at concentrations of 33.3 yg/liter HCN and higher (Figure 3). At
77.2 yg/liter it was 15.7 percent of controls. On the basis of acute
threshold toxic levels of 88 yg/liter at 10 °C (unpublished laboratory
data) for juvenile fish and 5./ yg/liter HCN as a safe concentration for
successful spawning, the "no-effect" level is approximately 7 percent of
the acute toxic level.
From these chronic exposure tests of two fish species, it is evident
that safe levels of HCN in the environment are much lower than the concen-
trations which will kill fish on short exposure. Where there is con-
tinuous exposure to low levels of cyanide from steel mills or other
sources of cyanide, some fish populations can be adversely affected by
concentrations higher than 7-12 yg/liter.
ACKNOWLEDGEMENTS
The author wishes to acknowledge the contributions of David Lind to
the experiment of fathead minnows, and to Walter Koenst for the experiment
on the brook trout.
REFERENCES
American Public Health Association, American Water Works Association and
Water Pollution Control Federation. 1971. Standard methods for the
examination of water and wastewater. 13th ed.
Broderius, S.J. 1970. Determination of molecular hydrocyanic acid in
water and studies of the chemistry and toxicity to fish of the
nickelocyanide complex. M.S. thesis, Oregon State University,
Corvallis.
52
-------
TABLE 8. NINETY-SIX HOUR LC5Q AND THRESHOLD CONCENTRATIONS OF HCN TO
FATHEAD MINNOWS AND BROOK TROUT JUVENILES (yg/liter).
No.
Tests
6
6
2
4
2
6
02
lmg/1)
6.0
7.0
6.0
6.0
8.8
8.0
°C
Fathead Minnow
15
20
25
Brook Trout
7
10
10
96 h
108
107
119
66
89
82
LCt^n
Threshold
108
107
119
65
88
75
IpH 8.0.
53
-------
Izatt, R.M., J.J. Christenson, R.T. Pack, and R. Bench. 1962. Thermo-
dynamics of metal-cyanide coordination. I. pK, AH°, and ASC values
as a function of temperature for hydrocyanic acid dissociation in
aqueous solution. Inorganic Chemistry, 1, 828 pp.
Leduc, G. 1966. Some physiological and biochemical responses of fish to
chronic poisoning by cyanide. Ph.D. thesis, Oregon State University,
Corvallis.
Mount, D.I., and R.E. Warner. 1%5. A serial-dilution apparatus for con-
tinuous delivery of various concentrations of materials in water.
U.S. Public Health Serv. Publ. 999-WP-23.
Neil, J.H. 1957. Some effects of potassium cyanide on Speckled Trout
(Salv&lsiwA fiowUnatti). Paper presented at Fourth Ontario
Industrial Waste Conference. Water and Pollution Advisory Committee,
Ontario Water Resources Commission, Toronto, Ontario.
54
-------
SECTION 6
MICROBIOLOGICAL INDICES OF THE QUALITY OF WATER AND METHODS
OF THEIR DETERMINATION
V.I. Romanenko
In the majority of cases, microbiological indices may be the best way
to characterize the quality of water used for both drinking and industrial
purposes. Microorganisms are excellent indicators which often exceed the
sensitivity of chemical and physical methods. The cells of microorganisms
react to minute changes in external medium. Under favorable conditions
they start to multiply more rapidly and their metabolism accelerates.
Some species of bacteria can exist only in the presence of a definite
class of chemical substances. Exhaustive knowledge is not available on
the life of millions of microorganisms, however, in the future, the
expanded use of microbiological indicators will undoubtedly develop.
It should be recognized that some of the major questions can be cor-
rectly answered only by microbiological specialists who are well versed
in the details of microbiological technique. In some cases it may be
necessary for the knowledgeable scientist to intentionally depart from
established methodology, while a similar departure in the hands of a
neophyte may lead to false results. When working with pathogenic micro-
organisms is considered, this may lead to serious consequences.
The present communication deals primarily with bacteria. Algae and
small invertebrates, including abundance, or activity, intensity of photo-
synthesis, while important as excellent indices of the condition of a
water-body, are not included in this consideration.
Microbiological indices may be divided into two categories as: (1)
the presence of bacteria, and (2) the intensity of one or another bacte-
rial process (Figure 1).
DETERMINATION OF THE QUALITY OF WATER BY THE NUMBER OF BACTERIA OR PRE-
SENCE OF PARTICULAR SPECIES
Quantity of Microorganisms
The quantity of microorganisms may be judged by their total content,
or by the content of separate physiological groups. Total numbers of
bacteria may characterize the condition of the water-body in general,
55
-------
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ORGANIC SUBSTANCES
CHARACTERIZATION MEDIUM FOR THE PRESENCE OF
SPECIFIC POLLUTANTS (FECES, HYDROCARBONS,
CELLULOSE, SULFIDES, POLYTHIONATES, PERCHLORATES,
OXIDES OF CHROMIUM, SULFATES)
INDICATIVE OF CONDITION OF WATER AND
INTENSITY OF SELF-PURIFICATION OR
PRESENCE OF TOXICANTS WHICH MAY BE
DETERMINED BY CHEMICAL MEANS
INDICATIVE OF CONDITION
OF WATER AND DIRECTION
OF PROCESSES
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i.e., the type of water-body. Through the efforts of Soviet Scientists,
the total number of bacteria has been investigated in detail in waters of
differing type. These investigations may, in general outline, be repre-
sented as in Table 1.
TABLE 1. CONTENTS OF BACTERIA IN WATERS OF VARYING TROPHIC DEGREE.
Type ofQuantity of Bacteria
Water-Body mln/ml Water-Body
Oligotrophic 0.1-0.5 Lakes Onega, Baikal
Mesotrophic 0.5-1.5 Rybinskoe Reservoir
Eutrophic 1.5-10 Tsymlyanskoe, Kakhovskoe
Reservoirs
Distrophic 1.5-3 Lake Melnezers in Latvia
In water-bodies of the oligotrophic type with clean water, the number
of bacteria varies from 0.1 to 0.5 mln/ml. With an increase in the
trophic degree, the number of bacteria also increases. In mesotrophic
water bodies the number reaches 0.5-1.5 mln/ml, and in eutrophic waters,
1.5 to 10 mln/ml. Distrophic waters are distinguished by high water
color values. The content of bacteria in them are the similar to the
mesotrophic condition, but the activity of these forms is considerably re-
duced.
The methodology associated with determination of the total number of
bacteria in water appeared as a result of the development of the ideas of
Vinogradski (1952) on the content of bacteria in soils. There are several
varieties of estimated strengths of bacteria in water (Kuznetsov and
Karzinkin, 1930). The most suitable of these methodologies, used at the
present time by most research workers, is the one suggested by Razumov
(1932).
For calculation of bacterial numbers, 1 to 50 ml of water, depending
on the trophic degree of the water body, is filtered through a membrane
filter with a pore size 0.2-0.3 mm. The filters are dried, stained with
laboratory conditioned erythrosine and the cell production is counted
under the immersion microscope. Calculations are made with due considera-
tion of the volume of filtered water (Rodina, 1965).
Determination of Living Bacteria by the Method of Titer in Sterile Water
Using C-Hydrolysate of Protein
A distinguishing feature of many microorganisms is the fact that they
do not develop on classic nutrient media (meat-peptone agar, meat-peptone
gelatine, etc.). As has been demonstrated (Romanenko, 1973), they grow
57
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well on media with minimum quantity of organic matter, equivalent by com-
position to that in natural water.
Determination of the number of living bacteria should involve the use
of water from the investigated water-body. The water should be collected
in bottles, then decanted into 10 ml test tubes and sterilized in an auto-
clave.
Since sterilization partially destroys the carbonates making the
water more alkaline, following autoclaving, the test tubes should be
placed into an atmosphere of carbon dioxide rendering the water neutral.
The test-tubes are subsequently placed in a stand in the order shown in
Figure 2. To the first test tube is added 1 ml of water by sterile
pippette. After thorough mixing, 1 ml of its contents is transferred
into the second test-tube. The process is repeated until the third test
tube is reached where subsequent transfers are performed in three repli-
cates for a greater statistical confidence. For majority of water bodies,
6-7 dilutions should be made. The seventh or the eighth test-tube serves
as control. Then the test tubes are placed for 7 days in a thermostati-
1 2345678
SEQUENTIAL DILUTIONS
Figure 2. Scheme of the order of test tubes in determination of living
bacteria by titer method. Numbers represent the sequence of dilutions.
58
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cally controlled incubator at a preselected temperature, usually 26 °C.
One drop of the solution of 14C-labeled protein hydrolysate having a
measured activity under a Geiger counter of the order of 0.2 x 10° imp/min
is added to each of the test-tubes by means of a Pasteur pippette, and
the test-tubes are incubated for 2 hours. Their contents are then fixed
with 0.25 ml of formalin and filtered through membrane filters with a
pore size 0.2-0.3 mm. After fixation and filtering, 5 ml of physiological
solution is filtered to remove the excess portions of the labeled pre-
paration. The filters are dried and the radioactivity of bacteria is mea-
sured under the Geiger counter. The final dilution in which the radio-
activity differs markedly from the control serves as an indicator of the
limit of dilution for bacterial reproduction. Using this methodology and
various classes of radioactive substances (e.g., phenol) the presence of
microorganisms and their number in the water samples may be determined.
Number of Saprophytic Bacteria
The number of saprophytic bacteria is the most reliable and sensitive
indicator of water pollution by organic substances of household origin.
This is a classic method used for about 100 years by sanitary organiza-
tions. It was proposed by Koch and has been used for counting of bacteria
in water by a number of different workers. The method of determination
of the number of saprophitic bacteria is quite simple. In the USSR,
standard dry nutrient medium (FPA) is made from fish flesh to which pep-
tone, sodium chloride and 15 percent agar-agar are added.
The medium to be inoculated is prepared in 50-100, or 200 ml flasks
depending on the quantity required. The water is taken into sterile
glassware by special samplers. The most simple model is the sampler of
Meier-Frantsev (Romanenko and Kuznetsov, 1974). Inoculation must be made
within one half of an hour after sampling. Samples may be stored in re-
frigerator or vacuum flask at low temperature for no longer than a day.
Inoculation may be of either the surface or depth type. In the former
case, the FPA should be melted. In laboratory this is best done in
boiling water- The flask with FPA is placed into boiling water until all
the medium is molten (not even smallest lumps of solid medium must be
left, otherwise the analysis will be spoiled). The medium may also be
melted on an open flame of a burner or on electric plate, but the proce-
dure must be carefully watched. The medium is then cooled to 40 °C. In
practice, microbiologists apply the flask to the cheek, if the medium
does not burn, it may be poured into Petri dishes.
For deep inoculation, the water to be tested is added into sterile
Petri dishes by sterile pippettes, then the FPA is poured and all is
thoroughly mixed. In this case, bacteria grow throughout the medium.
The colonies grown in the depths are physically smaller in size. For sur-
face inoculations, the medium is poured into the dishes, and after it
solidifies 0.5 to 1 ml of inoculum is added upon the surface and spread
over with a help of sterile glass spatula. The dishes are incubated at a
room temperature for 10 days. Then the number of colonies are counted,
and estimations are made with due consideration to the dilution.
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A good indicator of the cleanliness of water is the ratio of the num-
ber of saprophytic bacteria to their total number expressed in percent
(Kuznetsov, 1952; Romanenko, 1971). A summary is presented in Table 2.
TABLE 2. RATIO OF THE NUMBER OF SAPROPHYTIC BACTERIA TO THEIR TOTAL
NUMBER AS AN INDEX OF THE CLEANLINESS OF WATER.
Ratio, % Water Water-Body
0.003 or less Very clean Lakes Onega, Ladoga, Baikal
0.03 Clean Reservoirs: Rybinskoe, Sheksninskoe
Ruibyshevskoe
0.3
3 or greater
Dirty
Very dirty
Some parts of Volga River
Collectors of waste waters
Water in which the ratio is 0.003 percent may be considered exceedingly
clean; a ratio of 0.03 suggests clean conditions; 0.3, dirty; and 1-3 and
higher, very dirty.
Content of Bacterial Spores in Waters
Bacterial populations in water are dominated by non-spore forming bac-
teria. The ratio of bacillary forms to other bacterial groups is frequently
equal to 1:10.
In view of some workers, the spore forming bacteria are more often found
in the presence of hard to degrade organic substances, e.g., humic compounds,
etc. (Kholodnyi, 1957). In fact, spore forming bacteria are more abundant
in waters colored by humic substances, and in drainage from peaty grounds.
The number of spores is increased in proportion to the vegetative cells.
Determination of the total number of spore forming bacteria is laborious
and a time consuming procedure. The water or sediment to be tested is inocu-
lated onto agar plates according to the methods of Koch. It is then neces-
sary to wait for some time until the colonies age and begin to produce
spores, a microscopic examination of the colonies is performed using pre-
ferential staining of the spores (Omelyanski, 1932).
It is easier to determine the presence of spores in water than it is to
enumerate the spore-forming bacteria. For this type of determination, two
methods may be used. Both are based on the destruction of the vegetative
cells, and subsequent creation of conditions favorable for germination of
the spores.
Method of Heating—
A sample of water or sediment, either directly or after dilution, is
heated in a water-bath for 10-min. at a temperature of 80 C. Test-tubes
with tested water are then cooled and inoculation is made on MPA, or a mix-
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ture of MPA and yeast-agar.
at i on.
The number of colonies is counted after incub-
Treatment of Samples with 96 Percent Ethyl Alcohol
Romanenko and Daukshta (1975) have shown that the vegetative cells of
microorganisms die almost instantly under the action of strong ethyl alco-
hol, but the spores are preserved for a considerable period of time. This
is the basis for the second method of determining the quantity of spores in
a sample of water or mud.
The spores can be separated from the vegetative cell by several methods.
The tested water may be mixed in a test tube with alcohol and then inocu-
lated. To 5 ml of test water, 10 ml of ethyl alcohol is added (ratio 1:3 in
any volume), and 0.5 ml of the mixture is inoculated into Petri dishes with
MPA by the depth technique (with or without dilution using sterile water).
Another method is to pass a known volume of the test water through a mem-
brane filter, then 3-5 ml of ethyl alcohol is passed through the filter fol-
lowed by 3 ml of sterile distilled water to wash the alcohol from the filter.
The filters are then placed upon a layer of agar in Petri dishes. The number
of spores is estimated after incubation by the number of colonies present.
This method is particularly good in the case of clean water, where the number
of spores is small, because of the concentrating effect of filtering.
The number of spores in mud deposits or sediments, can be determined by
yet another method. A row of test tubes with sterile water for dilution is
used. One test tube, for example N 3 or 4, is filled with 9 ml alcohol (see
Figure 3). Thus, the water is passed through the alcohol in the process of
1
Figure 3. Position of the test-tube (N 2) with ethyl alcohol in the row when
determining the spores of bacteria in mud deposits.
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dilution allowing only living spores to pass the subsequent test-tubes
from which water is inoculated onto MPA.
The Number Specific Bacteria
Consideration of the field of medical microbiology is beyond the scope
of this paper. It should only be indicated that highly dangerous microor-
ganisms causing infectious diseases may be isolated using special media
and methods of enrichment. Isolation of such bacteria indicates a great
danger of such water for the health of people. Such water is quarantined,
and specific measures are to taken to protect the health of the public.
Fortunately such microorganisms are rarely present in water. Usually they
enter the water as a result of illness or from carriers. Some species of
bacteria may be indicators of pollution of the water by a certain type of
organic substances. For example, .species of the filamentous bacteria of
the genus CtadotlvU.*. are indicative of the presence of nitrogen-free or-
ganic compounds in water, and the presence of species of the genus
Spk&eAotituA indicates to pollution of the water by complex organic com-
pounds of nitrogen (Rasumov, 1961).
Coliform liter
The presence of coliform bacilli in water is indicative of a fresh
pollution with feces. In the water, high concentrations of these bacilli,
which are normally inhabitants of the human intestine, indicate pollution,
and the possible presence of pathohenic bacteria. Though certain strains
of coliform bacteria may persist for long periods of time and even multiply
in water, fresh fecal pollution is detected by those strains which only re-
cently entered the water. These are discerned by a characteristic colora-
tion of the colonies.
The number of these bacteria in water is determined by inoculation
with the highly specific Endo medium. The colonies can be grown either on
this medium directly in Petri dishes, or on membrane filters. In the
latter case, the water is passed through the filter, and it in turn, is
placed on Endo agar. Gold-colored colonies with a characteristic metallic
lustre are counted. The coliform titer is taken as the quantity of water
per one coliform-bacillus.
Separate Physiological Groups or Separate Characteristic Bacteria
The presence of a certain class of mineral or organic compounds in
water bodies may be established by the presence or increased content of
specific groups of bacteria. In many cases, it is easier to detect the
presence of such bacteria than it is to establish the presence of certain
chemical substances. It is generally accepted that it is possible to iso-
late separate microorganisms responsible for specific processes. An indi-
cation of the presence of that process is demonstrated by detection of an
increased content of a certain species or genera of bacteria.
Presence of Liquid Hydrocarbons (Oil Pollution) In Water--
Pollution of water or bottom sediments can be estimated by an increased
content of hydrocarbon oxidizing bacteria. For this process the samples
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of water must be inoculated onto the nutrient medium of Tauson, into which
a sterile solution (sterilization is carried out in sealed ampules by re-
peated boiling) of diesel oil, kerosine, or drop of mineral oil is added
as the main source of carbon. After adjusting of pH to the neutral value,
the medium is poured into test tubes. Into each, hydrocarbon and dilutions
of the tested water are added. The test tubes are placed into an incubator
at 26 °C for 5-10 days. At that time, the dilution may be determined in
which the film of bacteria was formed or in which the medium became cloudy.
In clean water the bacteria will either fail to grow, or will develop in
the 1st or 2nd dilution, while in polluted waters they will grow in the
3rd - 5th, sometimes in the 6-7th dilution.
In some cases instead of titer, a time consuming process, the intensity
of bacterial development may be determined. For this, the medium is poured
into 3-5 ml serum bottles, and to each a drop of hydrocarbon and 0.1 ml of
the suspect water are added. The bottles are then incubated. In 5-10
days they are examined, and the intensity of bacterial development (forma-
tion of the film, turbidity) is noted. The water is considered to be
clean if negative, or poor development is observed. When the pollution of
the water by petroleum products is significant, a thick skin-like film
with a white or pink tint rapidly develops, and occasionally the medium
becomes turbid.
Content of Cellulose Degrading Aerobic Bacteria (Pollution of Water With
Cellulose—
At the present time, many wood processing plants discharge wastes into
natural waters. Often wood fibers, lignin, and the like are found in them.
As a rule, cellulose degrading bacteria develop in enormous quantities in
the places of accumulation of wood fibers. They may isolated by inocu-
lating specific nutrient media, e.g., medium of Hatchinson (Romanenko,
1971). This medium contains the principal mineral salts required with
cellulose (filter paper) as the sole carbon source, with this medium, the
inoculum must be used from the 1st to the 6th dilution. Development of
bacteria in the l-2nd dilution indicates the presence of a relatively
small number, while development in subsequent dilutions shows the presence
of pollution with cellulose.
Presence of Sulphides and Thiosu1phates--The presence in water of re-
duced compounds of sulphur may be estimated by the presence of thionic
bacteria. These may be grown on the liquid or solid phase medium of
Beiering. The colonies often have a milky coloration owing to liberated
sulphur. Colonies can be quantified with a help of autoradiography
(Romanenko, e£ o£., 1975). This method utilizes ^c-carbonate added to
the agar medium. Bacteria are grown on membrane filters. After incuba-
tion, the filters are removed, treated with a weak solution of hydrochloric
acid, dried and glued to strip of compact paper equal in size to photo-
graphic film. In a light-free environment, the paper is applied to the
film, and both are rolled. In 2-5 days the film is developed in a contrast
developer. Colonies of thionic bacteria are counted as dark spots on the
film.
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Presence of Perchlorates and Chlorates—In some cases plants and fac-
tories discharge chlorates and perchlorates into receiving waters, e.g.,
NH4C103 or NH4C104- Recently a group of bacteria have been discovered
which quickly reduce this compound under natural conditions (Romanenko,
&£ o£., 1976). It also has been established that in the bottom sediments
of the majority of waters, the processes of reduction of perchlorates are
either slow or absent, but in the waters containing wastes of a given in-
dustry, the number of the specific groups of bacteria greatly increases.
The activity of these bacteria can be assayed using chlorine labeled
perchlorate (NH4 360104). Nutrient medium is prepared as described else-
where (Romanenko, a£ aJt., 1976). The media consists of salts, microele-
ments, vitamine B-]2» acetate, meat-peptone broth and perchlorate, 100 mg/1.
The medium is poured into bottles with ground stoppers 60-70 ml in
volume. To this is added 5 ml of test water, or 100 mg of sediment, and 1
ml of a sterile solution of labeled perchlorate with activity of 0.1 x 10^
imp/min. After 3-5 days the contents of the bottle are filtered through a
filter paper made slight acidic with nitric acid. A one percent solution
of silver nitrate is introduced which precipitates the chlorides. If the
natural content of chlorides in the medium is not great, 2-3 mg of sodium
chloride should be added to precipitate the labeled 36ci ions. The pre-
cipitated chlorides should be filtered through a membrane filter washed
with distilled water (10 ml), dried; and the radioactivity of the pre-
cipitate Ag 36ci measured. Accounting For the initial amount of perchlo-
rate added to the medium, one may estimate the amount which was converted
to chlorides. If the values are close to zero, the given bacteria may be
considered to be absent, and the process of reduction lacking. If reduc-
tion occurs, perchlorates may be reduced to the extent of 50-100 percent
of the added substance. Control experiments with the samples fixed by
formalin should always be conducted.
Presence of Chromates and Bichromates in Water—Chromates and
bichromates enter the waters from residues of electroplating shops, auto-
mobile factories, or chemical plants. Chromium is heavy metal toxic to
many organisms. In 1973 bacteria were isolated which decompose chromates
and bichromates under anaerobic conditions to chromium hydroxide using
these compounds as oxygen donors (Romanenko and Korenkov, 1975). These
bacteria may be used for purification of industrial wastes from chromates
and perchlorates (see above), as well as indicators of chromium oxides in
wastes.
In the places of permanent discharge of chromates, the water in the
near-bottom layers and the surface layer of sediments are rich in chromium
reducing bacteria which can be detected by a special inoculating (Romanenko
and Korenkov, 1975). The medium is prepared in flasks, sterilized and ad-
justed to circum neutral pH. It is then decanted into stoppered test-
tubes. After inoculations, with the test water, the cultures are incubated
for 7-10 days. The presence of the chromium reducing bacteria is indicated
by the medium turning from yellow (hexavalent chromium) to colorless (tri-
valent chromium). Changes in color may be accurately measured on a spec-
trophotometer.
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The presence of chromium reducing bacteria is indicative of pollution
of water by chromates. This is also true in the case where hexavalent
chromium is absent (indicative of decomposition). In this instance it is
impossible to analyse for the presence of chromium with a spectrophoto-
meter.
EVALUATION OF THE QUALITY OF WATER BY THE INTENSITY OF BACTERIAL PROCESSES
Daily Oxygen Consumption for Respiration of Bacteria as an Index of the
Hater Quality
All the organisms inhabiting water constantly consume oxygen for their
respiration. Only in rare instances will oxygen consumption by algae or
zooplankton exceed that of bacteria. In most cases, the role of microor-
ganisms in oxygen consumption and, therefore, in destruction of organic
matter, is greater than that of all other aquatic organisms taken together.
This is demonstrated by the indirect estimations of the number and rate of
reproduction of separate groups of organisms. In all cases, bacteria re-
produce most rapidly, as a further illustration, in the water passed
through a membrane filter where only bacteria remain, the rate of oxygen
consumption essentially does not change (Romanenko and Dobrynin, 1973).
It is well established that respiration is biochemical process in
which oxygen is combined with organic substances. In the end, the process
may be expressed by the following equation:
CH20 + 02 = C02 + H20
It follows from this that 32 weight parts of molecular oxygen are used
per 12 weight parts of carbon of organic matter. Thus, to determine the
amount of decomposed organic matter, the number of milligrams of consumed
oxygen by is multiplied by 0.375 (i.e., ratio 12:32). This application
can be made if the coefficient of respiration is 1, i.e., when carbohy-
drates are destroyed in respiration. Occasionally 0.8 or 0.9 is used
instead of 1. In this case, the value of the accepted coefficient should
be multiplied by 0.375. It should be noted that the experimental data on
respiration coefficients are very few, and often depart from theoretical
consideration.
To determine destruction of organic matter (Vinberg, 1934), 100-150 ml
bottles are filled with water straight from the Ruttner sampler. The pro-
cedure is performed in such a way so that the water will have as little
contact with atmosphere as possible. A rubber tube from the sampler is
placed in the bottom of the bottle and approximately 2-3 volumes are
passed through it. The bottles are closed with ground glass stoppers. In
two bottles, oxygen is measured immediately by the Winkler method (Alekin,
1954). The remaining two bottles are placed in a light-proof bag and
incubated for a standard time: a day in oligotrophic and mesotrophic
waters, 12 hours in more rich waters, 6-12 hours in eutrophic waters in
summer. In winter and summer at low temperatures, the time of incubation
may be decreased to several days. The time utilized should be the minimum
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for which a confident difference is established between the initial and
final oxygen content in bottles. The sensitivity of the oxygen methods
conducted by a skillful specialist may reach 0.05 mg 02/1.
Destruction of organic matter expressed in terms of oxygen is equal to:
°H ~ °k = °n rag/1/day
where:
OH - initial oxygen content (mg/1),
0|< - final oxygen content,
On - oxygen consumption.
Approximate summer values for respiration of the aquatic ecosystems in
water-bodies of various trophic degree are indicated in Table 3.
TABLE 3. OXYGEN CONSUMPTION FOR RESPIRATION OF ORGANISMS IN
WATER-BODIES OF VARIOUS TYPES (SUMMER VALUES).
Type of Water-Approximate Values of OxygenQuality of
Body Consumption, mg 02/1/day Water
Oligotrophic 0.05-0.1 Clean
Mesotrophic 0.1-0.3 Good
Eutrophic 0.3-3 Bad
With organic pollution 3-10 Very Bad
Total BOD
In sanitary microbiology, instead of daily oxygen consumption, the
value of total BOD, i.e., quantity of oxygen which may be used by the
microflora for oxidation of all the fractions of easily degradable or-
ganic matter, is often used (Lapshin, 1952). The experiments are per-
formed at a constant temperature of 20 °C.
Water from the bottom sampler is poured into a flask, its tempera-
ture is adjusted to 20 °C, and filled with a syphon into 6 100-150 ml
bottles. In a case of low oxygen content, the water is saturated with
oxygen by bubbling air through it. In two bottles, the initial content of
oxygen is measured by Winkler method, the remaining two are analyzed for
BOD. Bottles to be incubated are placed in water and ground covers with
the tested water (water lock) are fixed on them.
The difference in oxygen content between the initial water and at the
amount present after 3 days is designated as 6003, and after 6 days as
BOD6- The total BOD is determined by the formula:
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•l
BOD - '
tot. 2a, - a?
a-j = BOD for 3 days,
&2 - BOD for 6 days.
Heterotrophic Assimilation of C02
Heterotrophic assimilation of C02 is assimilition of carbon dioxide by
heterotrophic organisms. As early as 1921, it was shown by A.F. Lebedev
and later by Wood and Werkmann (1936) that heterotrophic organisms assimi-
late a small amount of C02 in their metabolism.
Until the present time, this phenomenon was described only in the
experiments performed at the Institute of Biology of Inland Waters, Academy
of Sciences, USSR (Romanenko, 1964; Sorokin, 1964). By this work it has
been shown that a certain amount of carbon dioxide is assimilated per
given increase in biomass of bacteria. It has also been found that there
is a direct proportionality between assimilation of C02 and respiration.
As a result, it has been established that the ratios between oxygen con-
sumption, increase of bacterial biomass, and assimilation of C02 is equal
to 1000:100:7 (mg 03, mg C, mg C). It should be realized, of course, that
this correlation is not as strict as may be found in chemistry and physics.
However, it may be used for the determination of productivity of bacterial
biomass (Kutnetsov it at., 1966). With the exception of meromictic lakes
in the zones above the layer of hydrogen sulphide, heterotrophic assimila-
tion of C02 prevails over chemosynthesis in water bodies. Heterotrophic
assimilation of CO2 in waters of varying types in shown in Table 4.
TABLE 4. MEAN MID-SUMMER VALUES OF HETEROTROPHIC ASSIMILATION OF C02 IN
WATER BODIES OF VARIOUS TYPES
Type of Water-Assimilation of CO2,Character of
Body g C/1/day Water
Oligotrophic
Mesotrophic
Eutrophic
-
0.1-1.0
1-5
5-70
10-200
Very Clean
Clean
Dirty
Very Dirty
The high sensitivity of the radioactive carbon method enables the
determination of the smallest values of C02 assimilation. The method is
as follows. 50-70 ml of water is placed in a bottle to which 1 ml of 14C-
labeled carbonate, NaH 14C03, is added having a specific activity of
1-5 x 106 imp/min under the Geiger counter. The bottles are placed in
lightproof bags, and the samples are incubated at the temperature of the
water-body for 24 hrs in oligotrophic and mesotrophic waters, 6-12 hrs may
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be used in eutrophic waters. The samples are then fixed with formalin
(0.5 ml per 100 ml of water) and passed through filters impermeable to bac-
teria. In the laboratory, the filters are placed for 10 min. upon a filter
paper moistened with 1 percent solution of hydrochloric acid, dried and
counted.
The content of hydrocarbonates should always be measured. This analy-
sis can be performed by direct titration when the water is clean and trans-
parent, and with distillation when it is highly sedimented or polluted. In
the former case, 100 ml of water is poured into a conical flask and 3 drops
of phenolphtalein solution is added. If the water does not turn pink, 1-2
drops of a solution of an alkali are added. The color is then neutralized
by addition of 0.1 N solution of HC1. After the color has disappeared (pH
8.3), 7 drops of methyl red or methyl orange are added, and titration is
repeated with HC1 until a stable pink color appears. The amount of the
acid used for titration from the time of addition of the second indicator
is multiplied by 12, thus obtaining the quantity in mg C of carbonate in 1
1 of the water.
If the test water is dirty, the carbonates are determined by distilla-
tion from acidified solution into an alkali (Kuznetsov and Romanenko,
1963).
The quantity of carbon dioxide assimilated by microorganisms is cal-
culated by the formula:
where:
ra - heterotrophic assimilation of C02 (pg C I/day),
r - radioactivity of microorganisms in the whole sample of
tested water (imp/min),
C|< - content of carbonates in water (yg C/l),
R - radioactivity of the carbonate solution, added into
the sample, imp/min.
Heterotrophic Assimilation as an Index of Bacterial Development
Heterotrophic assimilation of carbon dioxide as an index of bacterial
development was first used by the author (Romanenko, 1964). It is based on
the proportionality between increase of bacterial biomass and C02 assimila-
tion. For this purpose l4C-labeled organic substances may also be used,
but here difficulties arise owing to the fact that organic substances
quickly decompose, and it is not always possible to establish the location
of the experiment on the non-linear assimilation curve. Further rationale
favoring the use of carbonates is the fact that organic substances are not
introduced with carbonates. The carbonates are essentially neutral sub-
stances and in nutrient media there is almost always an excess of carbonate
as a buffer. It is not desirable to create an excess of organic substances
in nutrient media.
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The influence of different substances upon microorganisms may be deter-
mined using both pure and mixed cultures. One may also study the effect on
bacteria of various substances, for example, heavy metals, antibiotics,
antiseptics, pesticides, or toxic solutions of waste waters. Additionally,
the effect of temperature, pH, redox conditions and the like may be consid-
ered. In all the cases, it is necessary to take into account the presence
in the test water of other carbonate material. While it does not influence
the absolute value of C02 assimilation, it can effect the values of radio-
active uptake.
When studying the influence of toxicants on microorganisms, a nutrient
medium is prepared for bacteria. Weak solutions of meat-peptone broth may
be used. After sterilization and adjusting the pH value circum neutral
conditions, the medium is inoculated with a young culture distributed into a
series of 5-10 ml test tubes. The test is added to each at a required con-
centration, along with 0.1 or 1 ml of the labeled carbonate with an activity
of 0.5-1.0 x 106 imp/min. The test tubes are then plugged with rubber stop-
pers and placed in an incubator for one day after which the samples are
fixed with formalin and filtered through a membrane filter which retains the
bacteria. After treatment with weak hydrochloric acid (1 percent solution)
the filters are dried and the radioactivity is measured. The effect of
various concentrations of the tested substance on bacteria is indicated by
the radioactivity. The samples without the test substance serve as con-
trols. Their radioactivity is assumed to be 100 percent. The results of
one of such experiment on the action of silver ions are shown in Figure 4.
Silver concentrations of 10~8 - 10~7 M have no effect on bacteria. The in-
fluence of these ions was detected beginning with the concentration 10~6 M,
with complete inhibition of metabolism occurring at a concentration of
10-5 M.
This scheme may be used for determination of the toxicity of waste
waters. When the presence thermostable substances is expected, the samples
may be heated to 80 C for 10 min. to kill microflora, and then be added
into test tubes with indicator organisms. Alternatively, one may eliminate
preliminary heating and consider the control.
Determination of the Reserve and Rate of Consumption of Organic Substances
According to the Method of Right and Hobble
The living activity of microorganims may be estimated by the intensity
of consumption of labeled organic compounds (Right and Hobbie, 1965). The
method is based on the regularities of enzyme reactions. It allows the
determination of the speed of circulation of separate organic substances
and their reserves. Labeled glucose or acetate are most frequently used
for this purpose.
The water to be tested is poured into 9 bottles of 30-50 ml each which
are arranged in two parallel rows, four in each, with one remaining for
control of the purity of the labeled substance. To the first two bottles
are added 0.05 ml of the solution, to the second two bottles, twice the
greater amount and so on. To the 9th bottle is added a fixative (Lugol's
solution or formalin) and 0.1 ml of the solution of labeled substance.
69
-------
100
o
o
CD
I-
LU
O
QC
LU
Q_
66
33
0
fluorescens
Bacillus S
megatherium
i
i
10-9 io~8 icr7
SILVER ION CONCENTRATION, moles/l
Figure 4. Effect of Ag ions on bacteria. Analysis using the
heterotrophic assimilation of C02.
70
-------
Incubation is accomplished for one hour at high temperature, or 3-5 hrs.
at low temperatures. The samples are then fixed with Lugol's solution or
formalin, filtered through membrane filters and 5-10 ml of physiological
solution is passed through the filter. After drying, the radioactivity of
microorganisms is determined and the data plotted. On the ordinate the
values R ' t are plotted, where R is the radioactivity of the added organic
substance, t - time of incubation, r - radioactivity of organisms in the
samples. The abscissa displays the concentration of the added organic sub-
stance. The points are connected with a line. The segment on the abscissa
to the left zero on the ordinate axis (Figure 5) corresponds to the reserves
of organic matter in the tested water.
The rate of consumption of a given substance may be calculated by the
f omul a:
v -
v
S)
2000
1600
1200
800
400
0 35 70 105 140
PHENOL CONCENTRATION, jug C/l
Figure 5. Determination of the reserves of phenol in the water of the
Kanskoe Reservoir by the method of Right and Hobbie.
71
-------
where:
V - rate of consumption of a substance (yg 1/hr),
r - radioactivity of microorganisms in the whole sample,
(A+S) - quantity of organic matter in the sample:
A - added into the sample,
S - found on the graph (yg/1),
R - radioactivity of the added substance,
t - time of incubation.
This method may be used not only for the determination of the activity
of microflora, but also for the reserves and the rate of consumption of
separate toxic substances. For example, this technique was applied for
determination of the content and the rate of consumption of phenol in water.
Figure 5 shows the results of one experiment on the reserves of phenol in
the water of the Kamskoe reservoir. The points fit a linear progression
which cults off a segment of the abscissa equal to 13 mg/1 C of phenol.
This method may also be used for the analysis of other organic toxi-
cants. It is necessary only to have a corresponding ^-labeled subbstance.
REFERENCtS
Alekin, O.A. 19b4. Chemical analysis of inland waters L.
Kholodnyi, N.6. 1957. On the methods of quantitative studies of bacterial
plankton. Izbrannye trudy, Kiev, 3.
Kuznetsov, S.I. 1952. Role of microorganisms in circulation of matter in
lakes. M.
Kuznetsov, S.I. and G.S. Karzinkin. 1930. Method of estimation of bacteria
in water. Russ. gidrobiol. zh., 9.
Kuznetsov, S.I., and V.I. Romanenko. 1963. Microbiological study of
inland waters. M.-L.
Kuznetsov, S.I., V.I. Romanenko and N.S. Karpova. 1966. Number of bacteria
and production of organic matter in the water of the Rybinskoe Reser-
voir in 1963 and 1964. In Production and Circulation of Organic Matter
in Organic Matter in Inland Waters, M.-L.
Lapshin, M.I. 1952. Working out the method of purification of waste
water. M.
Lebedev, A.F. I9b2. About assimilation of carbon by saprophytes. Izvest.
Donskogo gos. un-ta, 3. 1921, (Cited in Shaposhnikov V. M. Carbonic
Acid in Metabolism of Heterotrophic Bacteria. Mikrobiol., 21:6. 1952).
Omelyanski, L.V. 1940. Manual on microbiology. M.-L.
72
-------
Razumov, A.S. 1932. Direct method of estimation of bacteria in water.
Its Composition With the Method of Koch. Mikrobiol. 1:2.
Razumov, A.S. 1961. Microbial indices of saprobity of waters polluted by
industrial wastes. III. About Taxonomy of Filamentous Bacteria.
Mikrobiol. 30:6.
Rodina, A.G. 1965. Method of aquatic microbiology. M.-L.
Romanenko, V.I. 1964. Heterotrophic assimilation of C02 by bacterial
flora of water. Mikrobiol. 33:4.
Romanenko, V.I. 1966. Heterotrophic assimilation of C02 as an indicator
of development of bacteria. DAN SSSR Ser. Biol., 168:1.
Romanenko, V.I. 1971. Total number of bacteria in the Rybinskoe Reservoir.
Mikrobiol., 40:4.
Romanenko, V.I. 1973. Multiplication of bacteria on natural water inform.
bull. In-ta Biol. Vnutr. Vod AN SSSR, I/.
Romanenko, V.I. and E.G. Dobrynin. 19/3. Oxygen consumption, dark assimi-
lation of C02 and rate of photosynthesis in natural and filtered
samples of water. Mikrobiol., 42:4.
Romanenko, V.I. and S.I. Kuznetsov. 1974. Ecology of microorganisms of
fresh waters. L.
Romanenko, V.I. and A.S. Daukshta. 19/5. Determination of the number of
bacterial spores with treatment of samples of water and mud with ethyl
alcohol. Inform. Bull. In-ta Biol. Vnutr. Vod AN SSSR, 26.
Romanenko, V.I. and V.N. Korenkov. 1975. Bacterial reduction of ions 004.
Inform. Bull. In-ta Biol. Vnutr. Vod AN SSSR, 25.
Romanenko, V.I., E.M. Peres, V.M. Kudryavtsev and M.A. Pubienes. 1975.
Radioautographic method of determination of thionic bacteria in water-
bodies. Inform. Bull. In-ta Biol. Vnutr. Vod AN SSSR, 26.
Romanenko, V.I. V.N. Korenkov and S.I. Kuznetsov. 19/6. A new species of
bacteria decomposing NH4C104 under anaerobic conditions. Inform. Bull.
In-ta Biol. Vnutr. Vod AN SSSR, 29.
Sorokin, Yu.I. 1964. Importance of dark bacterial assimilation of C02 in
trophies of water-bodies. Mikrobiol., 33.
Vinberg, G.G. 1934. Study of photosynthesis and respiration in water of a
lake. On the Problem of the Balance of Organic Matter. Communication
I. Tr. Limnol. st.v Kosino, 18.
Vinogradski, S.N. 1952. Microbiology of soils (problems and methods), M.
73
-------
Wood, H.G. and C.H. Werkmarm. The utilization of C02 by the propionic acid
bacteria in the dissimilation of glycerol. -J. Bacteriology, 30.
Wright, R.T. and I.E. Hobbie. 1946. The uptake of organic solutes in lake
water. Limnology and Oceanography, 10:1.
ZoBell. 1946. Marine microbiology. Mass, U.S.A.
74
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SECTION 7
AMMONIA AND NITRITE TOXICITY TO FISHES
Rosemarie C. Russo and Robert V. Thomann
AMMONIA
Introduction
Ammonia is a serious pollutant to aquatic life. It enters natural
water systems from several sources, including agricultural and industrial
wastes, and inadequately oxidized sewage effluents. Ammonia is also a
natural biological degradation product of nitrogenous organic matter.
The toxicity to fishes of aqueous solutions of ammonia or ammonium
salts is attributed to the un-ionized (undissociated) chemical species
(NH3) (Chipman, 1934; Wuhrmann, at at. 1947, Wuhrmann and Woker, 1948;
Hemens, 1966), with the ionized species (NH4+) considered nontoxic, or
significantly less toxic (Tabata, 1962). The concentration of un-ionized
ammonia is dependent on the chemical and physical characteristics of the
water, and therefore the toxicity of ammonia to fishes is dependent in
part upon the effect of these variables on the aqueous ammonia equilibrium.
The most important factors affecting this equilibrium are pH, temperature,
and ionic strength. The concentration of un-ionized ammonia increases
with increasing pH and temperature, and decreases with increasing ionic
strength.
The toxicity of ammonia to fishes is also influenced by dissolved
oxygen and free carbon dioxide. A decrease in the dissolved oxygen concen-
tration increases the toxicity of ammonia (Downing and Merkens, 1955;
Merkens and Downing, 1957), possibly because of increased ventilation by
the fish and a corresponding increase in the rate of flow of ammonia
across the gill tissues. Lloyd and Herbert (1960) reported that in waters
of low C02 concentration the toxicity of ammonia may decrease, and attri-
buted this to a reduction of pH at the gill membrane surface, brought
about by the expiration of C02- Other factors which exert an effect on
ammonia toxicity include previous acclimation of fish to low ammonia con-
centrations (Vamos, 1963; Malacea, 1968; Lloyd and Orr, 1969), physical
stress (Herbert and Shurben, 1963), and fish size (Penaz, 1965). Several
researchers have investigated the toxic effect of ammonia in combination
with other poisons (Herbert, 1962; Herbert and Shurben, 1964; Herbert and
Van Dyke, 1964; Vamos and Tasnadi, 1967; Brown, 1968; Brown, utat., 1969).
It is clear that the effects of the toxicants studies are generally
additive; sometimes proportionally, but not always.
75
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Aqueous Ammonia Equilibrium System
In aqueous ammonia solutions, un-ionized ammonia exists in equilibrium
with the ammonium ion and the hydroxide ion. The equation expressing this
equilibrium can be written as:
NH3(g) + nH2°U) - NH3-nH2°(aq) ^ NH4+ + OH~ + ^^U)
As indicated in this equation, the dissolved ammonia molecule exists in
hydrated form; it is hydrogen-bonded to at least three water molecules
(Butler, 1964). The dissolved un-ionized ammonia is represented for con-
venience as NH3; the ionized form is represented as NH4+; and total
ammonia is the sum of these (NH3 + NH4+).
The effect of pH and temperature on the aqueous ammonia equilibrium is
significant. For example, a pH increase from 7.0 to 8.0 within the range
0-30 °C results in a nearly tenfold increase in the concentration of NHs;
a temperature increase of 5 degrees between 0-30 °C at pH 7.0 results in
an NHs concentration increase of 40-50 percent.
There is no convenient method for measuring the concentration of NHs
and NH4+ separately. However, if total ammonia concentration, pH, and
temperature are known, the concentration of NHs may be calculated. Table
1 gives values of percent NH3 in aqueous ammonia solutions of zero
salinity.
TABLE 1. PERCENT UN-IONIZED AMMONIA IN AQUEOUS AMMONIA SOLUTIONS*
Temp.
0
5
10
15
20
25
30
pH Value
6.0
.0083
.0125
.0186
.0274
.0397
.0569
.0805
6.5
.0261
.0395
.0589
.0865
.125
.180
.254
7.0
.0826
.125
.186
.273
.396
.566
.799
7.5
.261
.394
.586
.859
1.24
1.77
2.48
8.0
.820
1.23
1.83
2.67
3.82
5.38
7.46
8.5
2.55
3.80
5.56
7.97
11.2
15.3
20.3
9.0
7.64
11.1
15.7
21.5
28.4
36.3
44.6
9.5
20.7
28.3
37.1
46.4
55.7
64.3
71.8
10.0
45.3
55.6
65.1
73.3
79.9
85.1
89.0
*[condensed from Thurston, at at. (1974)]
76
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In natural waters with low to moderate amounts of dissolved solids
(200-1000 mg/liter), this effect will slightly lower the concentration of
NH3. The magnitude of this effect will vary with the composition of the
water in question. For a water of high pH (8-9) and total dissolved solids
(IDS) of 500 mg/liter, which are predominantly calcium salts, the effect
on the fraction of NH3 present is approximately the same as if the tempera-
ture were lowered one degree. For waters of lower pH (5.5-6), but still
high in calcium, somewhat higher values of IDS (600-700 mg/liter) would be
required to produce a similar effect. For waters in which sodium chloride
is the dominant ionic species, approximately twice these amounts of IDS
would be necessary to produce a change comparable to a one-degree drop in
temperature.
Toxicity of Ammonia to Fishes
Concentration values for ammonia toxicity tests on fishes have been
variously reported as NHs, NHs-N, NH^H, total ammonia, total ammonia
nitrogen, and formula weight for ammonium salts. Calculation of the per-
centage of total ammonia as un-ionized ammonia has also been made in a
variety of ways, sometimes incorrectly. Recalculation of reported values
is not always possible because of failure to report essential water
chemistry parameters. Nonetheless, certain trends have developed which
give some approximation of lethal levels of ammonia for salmonids and some
species of warm water fishes.
In the case of short-term tests on rainbow trout (Salmo gcuAdneAA.) fry
and fingerlings, median lethal concentration (LCso) values as low as 0.2
mg/liter NH3 have been reported (Liebmann, 1960; Danecker, 1964). Other
researchers have reported LCso values ranging between 0.3-0.6 mg/liter
NH3 for tests of one day or less on rainbow trout (Lloyd and Herbert, 1960;
Herbert and Shurben, 1963, 1965; Ball, 1967; Lloyd and Orr, 1969; Smart,
1976), on brown trout (Salmo &iu£ta) fry (Penaz, 1965), and on Atlantic
salmon (Sa&no baton.} smolt (Herbert and Shurben, 1965).
In the case of short-term tests on fishes other than salmonids, Hazel,
at aJL. (1971) reported 4-day LCso values of 1.4 mg/liter NHs for striped
bass (AkMuwe. &a.x.a^UUU>} and 1.0 for sticklebacks (GoiteAo^tetM aco£ea^s).
Colt (1974) reported 4-day LC$Q values ranging from 2.4-3.8 mg/liter NH3
for channel catfish (IctaJtiuJwA punctatuA), and LC§o values ranging from
1.9-3.4 have been observed for fathead minnows (pAjme.ph.alu piome£o6)
(Thurston, unpublished data). A 17-hour LCso of 1.3 mg/liter NH3 was re-
ported for Gambusia (Gambu-6-ca a.^i.nu>) (Hemens, 1966), and a 24-hour
of 2.9 was reported for channel catfish (Robinette, 1976). Lower LC5Q
values between 0.35-0.50 mg/liter NH3 have been reported for 5- to 7-day
tests on bream (Afc-towu^ biama), roach (Rotc£u4 fwutituA), perch (PeAca
j$£uvxta£ctcA), and rudd (Sc.aA.divu.nuA eAytknophthaJbrnuA) (Ball, 1967). In a
longer test on rudd (Water Pollution Research, 1971) the LCso value for 7
days was 0.5 mg/liter NH3 and for 95 days was 0.24.
There are Title published data available on longer-term mortality
tests for fishes of any species. A three-month test on 200 rainbow trout
(Water Pollution Research, 1967) showed that 15 percent died at 0.22
77
-------
mg/liter NHs, and 5 percent died at both 0.11 and 0.06. In four separate
tests of 3-5 weeks' duration, the LC$Q values for rainbow trout fry were
between 0.5 and 0.6 mg/liter NH3 (Thurston, unpublished data).
Deleterious effects of ammonia at sublethal concentrations have been
observed by a number of researchers. Reichenbach-Klinke (1967), in a
series of one-week tests on 240 fishes of 9 species at concentrations of
0.1-0.4 mg/liter NHs, observed swelling of and diminishing of the number
of blood cells, inflammations, and hyperplasia, irreversible blood damage
occurred in trout fry at 0.27 mg/liter NH3- He also noted that these low
NH3 doses inhibited the growth of young trout and lessened their resistance
to diseases. Smart (1976) observed a high incidence of disease, as well
as gill damage, in rainbow trout exposed to 0.30-0.36 mg/liter NH3 for up
to 36 days. Flis (1968) reported that a 35-day exposure of carp (Cyp-^cno6
c.aApi.o) to a concentration of approximately 0.1 mg/liter NH3 resulted in
extensive necrobiotic and necrotic changes and tissue disintegration in
various organs.
Reduction in growth rates for rudd has been observed after 95 days at
concentrations greater than 0.1 mg/liter NH3 (Water Pollution Research,
1971) and for channel catfish at 0.14 mg/liter NH3 after 27 days
(Robinette, 19/6). Smith and Piper (1975) reported a reduction in growth
rates after 6 months and severe pathological changes in gills and livers
of rainbow trout after 12 months' exposure at 0.2 mg/liter NH3. For the
21-day period between egg hatching and swim-up stage, a reduction in
development of rainbow trout (length, weight, and sac absorption) was ob-
served at concentrations of 0.07 mg/liter NH3 and higher (Thurston, un-
published data). Concentrations as low as 0.002 mg/liter NH3 have been
reported to cause gill hyperplasia in fingerling chinook salmon
(OncofikynckuA ti>hcwyti>cha) in 6 weeks (Burrows, 1964.
Rainbow trout have successfully spawned in the laboratory at 0.06
mg/liter NH3 and have produced significant numbers of viable fry (Thurston,
unpublished data).
NITRITE
Introduction
Nitrite is present in only trace amounts in most natural freshwater
systems. In the process of nitrification, ^.e., the biological oxidation
of ammonia to nitrate, nitrite is produced as an intermediate product.
Primary treatment sewage plants discharge large quantities of ammonia and
partially converted ammonia into receiving waters, and as the nitrifica-
tion process proceeds downstream from the discharge point, nitrite levels
above normal may be detected. Of the total nitrogen being discharged by a
secondary treatment sewage plant, a lesser percent will be ammonia and a
higher percent will be nitrate, but also the percentage of nitrite will in-
crease. This percentage is related, in part, to how complete the nitrifi-
cation process has been within the plant before discharge. In some cases,
the amount of nitrite being discharge may raise the concentration of
78
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nitrite in a receiving water so that it may be significant to the stream
biota. Water reuse systems in some fish hatcheries also employ the nitri
fication process to reduce ammonia concentrations. Where these systems
are used, hatchery fish may also be subjected to increased nitrite
levels. It has been shown (Smith and Russo, 19/5) that nitrite induces
methemoglobinemia in rainbow trout. This results in a reduction in the
oxygen-carrying capacity of the blood, and fish may die from anoxia.
Toxicity of Nitrite to Fishes
The amount of published information on nitrite toxicity to fishes is
small, and most available data are from static tests of short duration;
many for 48 hours or less.
Of 13 fish species tested by McCoy (19/2), logperch (PeAcx.no.
was the most sensitive, dying in less than 3 hours at 5 mg/liter N02-N;
the common white sucker (CatoAtomuA comm&iAoyii) was the least sensitive,
surviving 48 hours at 100 mg/liter. A 96-hour LC50 of 1.5 mg/liter N02-N
was reported for mosquitofish (GambaA^a. OL^YIO^) (Wallen, n£aJL., 1957),
anl 10 mg/liter N02-N was reported to be fatal to minnows (PhoxsinuA lae.v,
&QOA) in 14 days (Klingler, 1957). Channel catfish (IctalusiuA punctcutuA)
were studied by Konikoff (1975) and Colt (1974), who reported 96-hour 1C 50
values of 7.5 and 12 mg/liter N02-N, respectively.
The LCso values for 96 hours for rainbow trout (SaJLmo QaJJidn&u.) ranged
from 0.2 to 0.4 mg/liter N02-N, with an asymptotic IC5Q of 0.14-0.15 (Russo
zJt al., 19/4). The susceptibility of cutthroat trout (So£mo cJLaMJui) to
nitrite appears to be comparable to that of rainbow trout. Observed
96-hour LC^Q values for cutthroat trout were 0.5-0.7 mg/liter N02-N, with
asymptotic LCso va1us of approximately 0.4 (Russo and Thurston, 1975).
Chinook salmon (Onc.oihync.huA tbhawytAcha) exhibited 40 percent mortality
in ?A hours at 0.50 mg/liter N02-N (Smith and Williams, 1974), and 96-hour
LC5n values for Chinook salmon were reported to be 0.88 mg/liter N02-N
(West, in, 1974).
There are differences in the reported susceptibilities of fishes to
nitrite. There does appear to be some genuine variability among fish
species, and there may be differences depending on fish size. It should
be pointed out that in some cases differences may be due to variations in
test water conditions. Recent work (Russo and Thurston, unpublished data)
has shown a wide variation in lethal concentrations of nitrite in waters
of different pH and salinity.
REFERENCES
Ball, I.R. 1967. The relative susceptibilities of some species of fresh-
water fish to poisons - 1. Ammonia. Water Res. ]_: 767-775.
Brown, V.M. 1968. The calculation of the acute toxicity of mixtures of
poisons to Rainbow Trout. Water Res. 2: /23-/33.
79
-------
Brown, V.M., D.H.M. Jordan and B.A. Tiller. 1969. The acute toxicity to
Rainbow Trout of flutuating concentrations and mixtures of ammonia,
phenol and zinc. J. Fish. Biol. 1_: 1-9.
Burrows, R.E. 1964. Effects of accumulated excretory products of
hatchery-reared salmonids. Bureau of Sport Fisheries and Wildlife
Research Report 66. G.P.O., Washington, D.C., 12 p.
Butler, J.N. 1964. Ionic equilibrium. Addison-Wesley Publishing Co.,
Inc., Reading, Mass., p. 129.
Chipman, W.A., Or. 1934. The role of pH in determining the toxicity of
ammonia compounds. Ph.D. Thesis, University of Missouri, Columbia,
Missouri, 153 p.
Colt, J.E. 19/4. Evaluation of the short-term toxicity of nitrogenous
compounds to Channel Catfish. Ph.D. Thesis, Univ. of California,
Davis: 94 p.
Danecker, E. 1964. Die Jauchevergiftung von Fischen--eine Ammoniakver-
giftung. Osterreichs Fischerei. 3/4: 55-68.
Downing, K.M. and J.C. Merkens. 1955. The influence of dissolved-oxygen
concentrations on the toxicity of un-ionized ammonia to Rainbow Trout
(SaJtmo gtuAdneAXxt Richardson). Ann. appl . Biol. 4_3: 243-246.
Flis, J. 1968. Anatomicohistopathological changes induced in Carp
(Ci/p-txno6 casip-io L.) by ammonia water. Part 1. Effects of toxic
concentrations. Acta Hydrobiol. 10: 205-224. Part II. Effects of
subtoxic concentrations. Ibid. 10: 225-238.
Hazel, C.R., W. Thomsen, and S.J. Meith. 1971. Sensitivity of Striped
Bass and Stickleback to ammonia in relation to temperature and
salinity. Calif. Fish and Game 5_7: 138-lb3.
Hemens, J. 1966. The toxicity of ammonia solutions to the Mosquito Fish
Baird & Girard). J. Proc. Inst. Sew. Purif. 265-271
Herbert, D.W.M. 1962. The toxicity to Rainbow Trout of spent still
liquors from the distillation of coal. Ann. appl. Biol. 50: 755-777.
Herbert, D.W.M. and D.S. Shurben. 1963. A preliminary study of the
effect of physical activity on the resistance of Rainbow Trout (Salmo
Richardson) to two poisons. Ann. appl. Biol. 52: 321-326.
Herbert, D.W.M. and D.S. Shurben. 1964. The toxicity to fish of mixtures
of poisons. 1. Salts of ammonia and zinc. Ann. appl. Biol. 53: 33-41,
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. Poll. 9_: 89-91.
80
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Herbert, D.W.M. and J.M. VanDyke. 1964. The toxicity to fish of mixtures
of poisons. II. Copper-Ammonia and Zinc-Phenol Mixtures. Ann. appl .
Biol. 53; 415-421.
Klingler, K. 1957. Natriumni trit, ein Langsamwirkendes Fischgift.
Schweiz. Z. Hydro!. 1_9: 5b5-b/8.
Konikoff, M. 1975. Toxicity of nitrite to Channel Catfish. Prog.
Fish-Cult. 37(2): 95-98.
Liebmann, H. 1960. Handbuch der Frishchwasserund Abwasserbiologie - II.
Munchen.
Lloyd, R. and D.W.M. Herbert. 1960. The influence of carbon dioxide on
the toxicity of un-ionized ammonia to Rainbow Trout (Salmo gcUdn&vU
Richardson). Ann. appl. Biol. 48: 399-404.
Lloyd, R. and L.D. Orr. 1969. The diuretic response by Rainbow Trout to
sub-lethal concentrations of ammonia. Water Res. 3; 335-344.
Malacea, I. 1968. Untersuchungen uber die Gewohnung der Fische an hone
Konzentrationen Toxischer Substanzen. Arch. Hydrobiol. 65: 74-95
McCoy, E.F- 1972. Role of bacteria in the nitrogen cycle in lakes. EPA
Water Pollution Control Research Series, 16010 EHR 03/72. 23 p.
Merkens, J.C. and K.M. Downing. 1957. The effect of tension of dissolved
oxygen on the toxicity of un-ionized ammonia to several species of
fish. Ann appl. Biol. 45_: 521-52/.
Penaz, M. 1965. Vliv Amoniaku na Jikry a Pludek Pstruha Obecneho, Salmo
m. fa/u.o. Zoo! . listy 14: 47-54.
Reichenbach-Klinke, H.-H. 1967. Untersuchungen uber die Einwirkung des
Ammoniakgehalts auf den Fischorganismus. Arch. Fischereiwiss. 17: 122-
132.
Robinette, H.R. 1976. Effect of selected sublethal levels of ammonia on
the growth of Channel Catfish ClctaluAuA pu.nc.taMu). Prog. Fish-Cult.
38(1): 26-29.
Russo, R.C., C.E. Smith, and R.V. Thomann. 1974. Acute toxicity of
nitrite to Rainbow Trout (Salmo gcuAdn&ii) . J. Fish. Res. Board Can.
31_: 1653-1655.
Russo, R.C. and R.V. Thurston. 19/5. Acute toxicity of nitrite to
Cutthroat Trout (Salmo cbviki) . Fisheries Bioassay Laboratory Tech.
Rept. No. 75-3, Montana State University, Bozeman: 13 p.
Smart, G. 1976. The effect of ammonia exposure on gill structure of the
Rainbow Trout (Salmo QcuMn&ii] . 0. Fish Biol. 8; 471-475.
81
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Smith, C.E. and R.G. Piper. 1975. Lesions associated with chronic expo-
sure to ammonia. In The Pathology of Fishes. W.E. Ribelin and G.
Migaki (Eds.), University of Wisconsin Press, Madison, pp. 497-514.
Smith, C.E. and R.C. Russo. 1975. Nitrite-induced methemoglobinemia in
Rainbow Trout. Prog. Fish-Cult. 37j3): 150-152.
Smith, C.E. and W.G. Williams. 1974. Experimental nitrite toxicity in
Rainbow Trout and Chinook Salmon. Trans. Amer. Fish. Soc. 103:
389-390.
Tabata, K. 1962. Toxicity of ammonia to aquatic animals with reference
to the effect of pH and carbon dioxide. Bull. Tokai Reg. Fish. Res.
Lab. 34: 67-74.
Thurston, R.V., R.C. Russo, and K. Emerson. 1974. Aqueous ammonia equili-
brium calculations. Fisheries Bioassay Laboratory Tech. Rept. No.
74-1, Montana State University, Bozeman: 18 p.
Vamos, R. 1963. Ammonia poisoning in Carp. Acta Biol. £: 291-297.
Vamos, R. and R. Tasnadi. 1967. Ammonia poisoning in Carp. 3. The
oxygen content as a factor influencing the toxic limit of ammonia.
Acta Biol. Szeged ]_3: 99-105.
Wall en, I.E., W.C. Greer, and R. Lasater. 1957. Toxicity to GcmbuA
-------
SECTION 8
A RESEARCH SYSTEM FOR DEVELOPING FISHERIES STANDARDS FOR WATER QUALITY,
CONSIDERING THE PECULIARITIES OF TRANSFERRING EXPERIMENTAL DATA TO
NATURAL WATER BODIES
L.A. Lesnikov
At present in the USSR, two groups of standards of water quality are
being developed with the purpose of protection of the waters from pollu-
tion: 1) sanitary hygienic, and 2) fisheries MPC (maximum permissible
concentrations). Both groups of the MPC are to be approved by the
Ministries of Health of the USSR, and Fisheries of the USSR respectively,
and will be official interdepartmental standards.
When setting fisheries standards, complex investigations are performed
on water-bodies, and in laboratories. The latter are most used, since
only experimental work enables the investigator to establish clearly the
relation between concentrations of pollutants, and the degree of the dis-
turbance in organisms.
The results of field investigations may be taken into account, but
only for comparison, since alterations in organism response are always
the result of the actions of not only the pollutants, but also of other
natural and anthropogenic factors. In addition, under experimental condi-
tions, it is possible to evaluate the effect of substances on the living
functions of organisms, while in natural waters the influence of pollution
is not direct, but functions through environmental and ecological condi-
tions.
In the USSR there are more than 100 fisheries standards. In 60
percent of cases the fisheries standards are close to the corresponding
sanitary-hygienic ones, but in 40 percent they are more stringent,
occasionally 10-100 times more restrictive (Lesnikov, 1974).
In the USA, only recently investigations were started on the chronic
action of low concentrations of pollutants on aquatic organisms (Brungs,
1972). In the USSR, such investigations have been performed since 1938-
1939 (Eltsina, 1939; Stroganov, 1940; Stroganov, 1941).
From the very beginning, elaboration of the fisheries MPC in the USSR
was based on the following principles formulated by N.S. Stroganov (1941):
83
-------
1. To investigate, as far as possible, the influence of tested
substances on the whole life-cycle of aquatic organisms, or
on its most susceptible stages.
2. To conduct observations for the entire period of a complete
biological cycle (for crustaceans not less than 15 days),
or of its separate stages.
3. To direct attention to the influence of toxic substances
not only on survival, but also on the main physiological
functions of an organism (heart beat, reproduction, breath-
ing).
4. To use in the experiment various organisms, since different
organisms respond to the action of substances in different
ways.
5. To fix permissible levels in waters according to the weakest
biological link.
Later, all the indices of the action of substances on organisms were
subdivided into two categories, principal and supplemental. The princi-
pal category describes those characteristics of the existance of popula-
tions of organisms under natural conditions which are well understood,
i.e., rate of growth, reproduction, death. Supplementary indices are
used to clarify the character of action of a given substance. The supple-
mentary indices are very important in order to establish the causes of
death of organisms in natural waters.
For deeper understanding of the character of action of a substance,
it is necessary to perform experiments on more than one species of aquatic
organism. When choosing test-organisms their role in the circulation of
substances and their relative sensitivity to pollutants was considered.
The organisms are divided into 3-4 groups according to their relative
sensitivity to toxic substances (Lesnikov, 1968, 1969, 1975; Stroganov,
1971). Organisms primarily from the first two groups (the most sensitive
organisms) are taken for the experiments.
In the USSR, a set of tests is used to study the influence of hydro-
chemical indices of water (gaseous and ionic composition of water, pH,
content of organic substances, dynamics of nitrogen compounds, time of de-
composition of pollutant) on:
1. Producers: Scene, Gammasuu pmt&x..
Fish - Salmo xtAx^eoi, Coiegonoi pe£ed and
others, (parallel experiments of
influence on eggs, larvae, less
than 1 year age class and in some
instances, on yearlings).
84
-------
Reducing bacteria - number of bacteria and saprophytes
growing on MPA (Lesnikov, 1973, 1975).
N.S. Stronganov (1975) added a series of experiments on 'aquatic plants
(elodea and others) to these tests.
Investigations
the chronic lethal
lysis of the dependence of
of toxic substances (Jones
estimation of the duration
experiments:
for each level are conducted to reveal the acute lethal,
and the sublethal (inhibitory) effect of toxicants. Ana-
survival time of organisms on the concentration
1957, 1964; Lesnikov, 1973, 1976) enables an
of acute (short-term) and chronic (long-term)
Type of Experiments
Acute
Intermediate
Chronic
Invertebrates
10 days
1 month
3 months
Fish
15 days
3 months
6 months or longer
The boundary between acute and chronic lethal concentrations was taken
to be a characteristic bend of the curve. Between chronic lethal and sub-
lethal concentrations, the position of the asymptote of Stroganov was used
as the determining factor. The bend for different species and substances
was typically between the 4-14th day of the experiment. The duration of
chronic experiments was determined by the detected cases of the remote
negative effect.
For some organisms (Vapknia, and certain algae) the methods of experi-
mentation on populations are established. For Vapkrua, the whole set of
population reactions such as logrythmic growth of the biomass, a regular
transition from parthenogenetic to bisexual reproduction, etc. are repro-
duced. It has been established that when a population reaches its satura-
tion biomass, its sensitivity to pollutants becomes 3-3.5 times greater
than before (Table 1).
TABLE 1. CHANGES IN BOUNDARY CONCENTRATIONS OF 3 POLLUTANTS
INFLUENCING BIOMASS IN PAPHNIA MAGMA.
(Boundary concentrations are estimated on the basis of regression equation)
Pollutants
Sewage from oil re-
finery, %
Sewage from chemical
plant, %
Cobalt chloride,
mg Co++/1
6-9 9-12
5.0 1.7
2.9 1.6
0.1
Days of
12-15
1.4
1.1
0.1
Experiment
15-18
0.7
1.9
0.1
18-21 21-30
1.6
-
0.09 0.02
85
-------
At present, the influence of more than 30 different substances on
jd populations have been investigated, in only two cases were excep-
tions observed. These exceptions were noted: 1) when studying the in-
fluence of sodium chloride on populations of V. magna (adaptation limits
of the Va.phyiLa.to salination exceeded the increase in sensitivity), and 2)
in the case of the influence of sulphate sewage from paper mills on V.
magna. (adaptation to organic substances). In the both instances, the ex-
periments with V. toYiQ4Api.no, yielded typical results. This latter species
adapts to a lesser degree to increased salinity, and is more sensitive to
saprobic pollution. In the case of an. increase of sensitivity of popula-
tion to pollutants, it is necessary to deal with the increase of sensitiv-
ity of individuals in the moment when the saturation biomass in reached
(Lesnikov, 1970). Four main types of effects of substances on the pro-
ductive properties of population can be discerned:
Type 1. The substance increases mortality of individuals with-
out disturbing the functions of growth and reproduc-
tion in individuals. This type is analoguous to the
effect of predation and fishing on the population. To
some extent, the death of some of the individuals is
compensated by intensification of growth and repro-
duction in others. Thus, even the death of a part
of the community may not lead to a notable decrease
in the rate of growth of the biomass.
Type 2. The substance influences the rate of metabolism and,
thus the rate of growth in weight of individuals, but
the reproductive function remains undisturbed. Usu-
ally, the biomass of the test population is close to
that of the control, but is attained at a later time.
Type 3. Reproductive function of the population is disturbed.
The maximum biomass in the test population usually
does not reach that of the control. In Vaphnija., no
cases of formation of ephippia during the period of
high biomass are observed, although these processes
take place in the control. This is the most dangerous
type of the effect.
Type 4. Under the influence of substances, the biomass is
higher than in the control. For the Vaphyiia popula-
tions, this is usually the influence of sewage which
passed through biological treatment (discharge of a
part of activated sludge increases the nutrient base
for VaphruM.). However, if the sewage is derived from
industrial enterprises, the stimulation may be accom-
panied by evident intoxication of a part of the indivi-
duals. In addition, with this type of effect, it is
necessary to account for the possibility of a change in
the water quality which may render it unsuitable for
valuable fish, e.g., the whitefishes and salmonids, which
can be displaced by less valuable fish.
86
-------
Evaluation of the influence of pollutants upon whole populations of
aquatic organisms makes it easier to extrapolate the experimental data to
natural waters. One may expect that similar disturbances will take place
in nature.
It is necessary to take into account the fact that in nature, a pollu-
tant acts to influence a number of other factors. The first attempt at
classification of these factors was made by Wuhrman and Woker (1955). An
improvement of this classification is offered in Table 2.
In addition, usually not one, but a number of pollutants are present in
natural waters. It is generally accepted in the USSR that it is possible
to estimate the sum of the effects of substances as an additive function.
The cases of synergism and antagonism are to be accounted for only in acute
lethal concentrations, since they are less important under the conditions
of chronic action.
Pollutants exert simultaneous action on a number of species of or-
ganisms. The net result depends on the relative sensitivity of these
species to the pollutant. Therefore, creation of a scheme of classifica-
tion of organisms according to their sensitivity to toxicants is directly
related to the question under consideration. The relation of organims to
saprobic pollution is well considered in various saprobity systems. The
relation to toxicants requires special elaboration. Here a scheme of
division is proposed (Table 3) accounting for analogous elaborations in the
USSR (Stroganov, 1971) and in the USA (Muirhead-Thomson, 1971).
A more detailed division of organisms according to their sensitivity is
desirable, but difficult, for two reasons: 1) the indicated classification
is a generalization, and the specific position of many organisms requires
further clarification, since within each of the groups there are further
gradations of sensitivity; and 2) the sensitivity of single species to
various types of toxic substances is different. It is possible that future
considerations will require the creation of not one, but a number of such
systems, while more useful for classification, such an addition will make
the system more cumbersome.
Thus, in analyzing the actual differences between the conditions in
experiments and those of natural waters, the experimental data can be used
with greater assurance if the analytical situation employs natural condi-
tions.
87
-------
TABLE 2. THE INFLUENCE OF FACTORS ON THE CHARACTER OF ACTION OF THE
POLLUTANTS ON AQUATIC ORGANISMS
Type of Influence
Factors
Character of Influence
On properties
of pollutants.
On the time and
condition of the
contact or orga-
nisms with pollu-
tants.
On sensitivity of
organisms to the
action of toxi-
cants.
COcontent in water.
Content of mineral
substances in water.
Content of organic
matter.
pH of the water.
C02 content,
Dynamics of water
masses (velocity of
currents, convection,
stagnation, etc.).
Distribution and mi-
grations of organisms
in a water-body.
Size, age, sex.
Contents of other sub-
stances in water.
Discrepancy between re-
quirements of organism
and environmental con-
ditions.
Weakening of organisms
due to starvation, par-
asites, infections, etc.
Stress conditions of
organism.
Rate of oxidation of pollu-
tants.
Coagulation of some substances,
neutralisation of acids and
alkalis.
Formation of complex compounds
with pollutants absorption.
Change in the degree of disso-
ciation of pollutants, and in
their toxicity.
Change in the direction of
chemical reactions and buffer
properties of water.
Dilution of pollutants, possi-
bility of their concentrating
in certain parts in water or in
bottom sediments.
Possibility of organisms mi-
grating into polluted water.
Differences in sensitivity of
stages. Usually males are more
sensitive than females.
Antagonism and synergism of
toxic substances.
Increasing sensitivity to in-
juring factors, including the
action of pollutants.
The same as above.
Changes in time of reaction and
in intensity of injuring.
-------
TABLE 3. DISTRIBUTION OF ORGANISMS ACCORDING TO THEIR RELATIVE
SENSITIVITY TO TOXIC SUBSTANCES (TOXOBITY)
Group of Organisms
Ecologi-
cal
Fishes
Zoo-
plankton
Zoo-
benthos
Taxonomic 01 i go-
Group toxobity
Salmon All species
Whitefish All species
Perch Zander
Sturgeon All species
Carp
Sheatf ish
Pike
Eel
Cod
Stickleback
Watermites
Cladocerans
Ostracods
Copepods
Rotifers
Ciliates
Zooflagel-
lats
Crustaceans Gammarids
mysids, coro-
phiids, cray-
fish
Harpacti-
cides
Mollusks
Aquatic Ephemerop-
Insects terans
Aquatic
Worms
Toxobity
Betameso-
toxobity
Caspian zan-
der, perch,
ruffe
Bream, white
bream, roach
bleek
Sheatfish
Pike
Burbot
Stickelback
Daphnias,
Si das, pre-
datory clado-
cerans
All species
Calanoids
All species
All species
Bivalvia
Mayflies,
Draggonflies,
Caddis Flies
Alphameso-
toxobity
Carp, cru-
cian carp,
tench big-
head, amur
loach
Pike
All species
Chidorids
Bosminids
All species
Cyclopoids
Bdel lids
Mobile forms
All species
All species
Gastropoda
Chironomids,
Beetles, Bugs
Culicines
Oligochaets
Leeches,
Planaria
Poly-
toxobity
Mobile forms
All species
Tubificids
Lumbriculids
89
-------
REFERENCES
Brungs, W.A. 1972. Effects of pesticides and industrial wastes on surface
water use. River Ecology and Man. New York and London.
Eltsina, N.V. 1939. The influence of the sea salt on development of fresh-
water and adaptation of them to the conditions of increased
salinity. Vopr. ecol. i biotsenol, issul. 4.
Jones, J.R.E. 1957. Fish and river pollution. Aspects of River Pollu-
tion, London, Butterworths.
Jones, J.R.E. 1964. Fish and river pollution, London, Butterworths.
Lesnikov, L.A. 1968. Peculiarities of the fisheries evaluation of the in-
fluence of pollution by hydrobiological data. In col.: Sanitarn.
gidrobiol. i vodnay taksikologiya. Riga, N. 2.
Lesnikov, L.A. 1969. On the development of saprobity systems for the
evaluation of various types of pollution. Tez. dokl. simp, po vodn.
toksikol. L.
Lesnikov, L.A. 1970. Peculiarities of action of pollutants on populations
of aquatic organisms. In col.: Voprosy vodnoi toksikologii. M.
Lesnikov, L.A. 1973. Theoretical and methodical aspects of elaboration of
the fisheries MPC. Vodnye resurcy, N 4.
Lesnikov, L.A. 1974. Protection of waters from pollution from the point of
view of fisheries. Izv. GosNIORH, V. 98.
Lesnikov, L.A. 1975. Methodic directions for establishing MPC of harmful
substances in the waters of fisheries. L.
Lesnikov, L.A. 1975. Expansion of the saprobity system and extrapolation
of experimental data into fisheries waters. In col.: Formation and
Control of the Quality of Surface Waters, Issue 1.
Lesnikov, L.A. 1976. Comparison of different methods of conducting of
experiments in aquatic toxicology. Izv. GosNIORH, V. 109.
Muirhead-Thomson, R.C. 1971. Pesticides and freshwater fauna. London and
New York.
Stroganov, N.S. 1940. Toxicology of aquatic animals in relation to the in-
fluence of sewages on a water-body. Zoo!., zh., V. 19.
Stroganov, N.S. 1941. Theoretical foundations of the solutions of problem.
Lichen, zapiski MGU, Issue 60.
90
-------
Stroganov, N.S. 1971. Methods of determination of toxicity of water
medium. In coll.: Methods of Biological Investigations in Aquatic
Toxicology. M.
Stroganov, N.S. 1975. Tin-organic compounds and living processes in
aquatic organisms. M.
Wuhrman, K. and H. Woker. Influence of temperature and oxygen tension on
the toxicity of poisons to fish. Verh. internat. Verein. Limnol.,
V. 12.
91
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SECTION 9
COLLAGEN AND HYDROXYPROLINE IN TOXICOLOGICAL
STUDIES WITH FISHES
Foster L. Mayer and Paul M. Mehrle
Chronic toxicity studies with fish are expensive, high-risk endeavors,
requiring from 10 months to one year to conduct. Such studies include
growth, reproduction, and survival of adults, and growth and survival of
the offspring. As a consequence, there has been much interest in the de-
velopment of alternative methodologies that provide similar information
with less expenditure of time and effort. Grant and Schoettger (1972)
stated that the monitoring in fish of biochemical factors that can be cor-
related with toxicant exposures and residues, provides a useful means of
anticipating the subtle, adverse impacts of organic contaminants on the
fish. To date, however, investigators have used biochemical measurements
alone in many studies to arrive at rather broad conclusions, without deter-
mining the ultimate effects on growth, reproduction, and survival (i.e.,
whether the chemical changes observed were within the adaptive capacity of
the fish). Therefore, attempts were made to assess the possibility of
using biochemical factors as indicators or predictors of growth and devel-
opment in fish, thus decreasing the time required for chronic toxicity
determinations. Growth of fish is usually evaluated by measuring weight
and length; however, biochemical changes due to intoxication should occur
before reductions in growth. As potential indicators of growth and devel-
opment, backbone collagen and the hydroxyproline concentration in collagen
were selected, and these measurements were incorporated into basic studies
with toxaphene to provide information for the establishment of water qual-
ity criteria (Mayer e£o£., 1975, 1977; Mehrle and Mayer, 1975a, 1975b,
1976).
RATIONALE FOR MONITORING HYDROXYPROLINE AND BACKBONE COLLAGEN
Collagen is the major fibrous protein of all vertebrates and most of
the invertebrate phyla (Piez and Likens, 1958). Its most important func-
tion in vertebrates is to serve as the major component of the organic ma-
trix of connective tissues and bones. The collagen molecule is unique in
its amino acid content (Harrington and Hippel, 1963); the amino acids
hydroxyproline and proline combined make up about one-tenth and glycine
another third of the total amino acid composition in collagen. In animal
tissues, hydroxyproline is found only in the protein's collagen and elas-
tin. Since the total amount of elastin is very small in comparison with
92
-------
that of collagen, and since the hydroxyproline content of elastin is only
about one-tenth that of collagen, its contribution to the total hydroxy-
prolin content is negligible (Green it al., 1968). The synthesis of colla-
gen, like that of other proteins, occurs on the ribosomes in fibroblasts,
osteoblasts, and chondroblasts. However, hydroxyproline and hydroxylysine
in the collagen molecule are not derived from incorporation of the free
amino acid into the polypeptide, but instead are derived from the hydroxy-
lation of their respective precursors, proline and lysine, after incorpor-
ation of proline and lysine into the polypeptide protocollagen. The en-
zyme collagen hydroxylase, or peptidyl proline hydroxylase, which begins
its activity during gastrulation, catalyzes hydroxylation; vitamin C,
a-ketoglutarate, and ferrous ion serve as cofactors for the enzyme
(Mussini it al., 1967).
The importance of collagen in animals in shown by its wide distribution
and many functions during growth and development. One of its major func-
tions is to serve as the structural support for bones. Dried bone consists
of one-third organic matrix and two-thirds minerals. About 90 percent of
the organic matrix is collagen, and the rest consists of mucopolysaccha-
rides, mucoproteins, and lipids (Nusagens &taJt.t 1972). Calcification
and mineralization take place around and within the collagen fibrils in
bone and, as development proceeds, the deposition of calcium and phosphate
produces mature bone.
The use of collagen as a representative "differentiated" protein in
the study of embryonic development has been reported in amphibian embryo-
logical investigations (Green it at,, 1968; Rollins and Flickner 1972).
Collagen synthesis, though repressed during the first cleavage stages of
the embryo of a frog (Xenopo6 laiv-a>), begins during gastrulation and in-
creases 500-fold through neutrulation, hatching, and posthatching stages.
Also, decreased hydroxyproline excretion in urine has shown promise as a
detector of nutritional deficiency and reduced growth rates in humans
(Whitehead and Coward, 1969). Fish continue to grow throughout life, and
their vertebrae continue to elongate and enlarge with growth. It has
therefore hypothesized that backbone development should increase in pro-
portion to increase in growth, and that increases in collagen and hydro-
xyproline would be indicators of this growth.
RELATION OF COLLAGEN AND HYDROXYPROLINE TO GROWTH
Brook trout (So£ve£oto6 fiontinati*), fathead minnows (P-umphaLn>
ptiomitaA), and channel catfish (Ic£o£uAoi punctaJuA) were continuously ex-
posed to, toxaphene in water; the proportional diluter systems used were
modeled after Mount and Brungs (1967) and modified as recommended by
McAllister it oJL. (1972). The diluter systems delivered five concentra-
tions of toxaphene, with a dilution factor of 0.5 between the concentra-
tions, and a control (Table 1). We used flow-splitting chambers as
designed by Benoit and Puglisi (1973) to thoroughly mix and divide each
toxaphene concentration for delivery to duplicate exposure tanks. Artifi-
cial daylight was provided by the method of Drummond and Dawson (1970),
and water temperatures were maintained within +_ 0.2 °C.
93
-------
TABLE 1. SUMMARY OF EXPERIMENTAL CONDITIONS AND SAMPLING PERIODS USED DURING CONTINUOUS EXPOSURE OF
BROOK TROUT, FATHEAD MINNOWS, AND CHANNEL CATFISH TO TOXAPHENE9.
Toxaphene
Species and Concentration
Life Stage (ng/1 )
Brook trout
Eggs 39-502
Fry
Fathead minnows
Experiment 1 94-727
Experiment 2 13-1/3
Eggs
Fry
Channel catfish
Adults 49-630
Eggs
Fry
Age At
Water Initiation
Temperature of Exposure
(°C) (Days)
9
25
25
12-26
26
26
eyed eggs
10
40
Ob
0
2.5 years
0^
0
Duration of Growth Biochemical
Exposure Determinations Determinations
(Days) (Days) (Days)
22 before hatch
90 30, 60, 90 7, 15, 30, 60, 90
150 60, 90, 150 150
295 30, 98 98
5
30 30 30
<
100 50, 100 100
7 1
90 5, 30, 60, 90 15, 90
Experimental details are given by Mayer vt al. (1975, 1977) and Mehrle and Mayer (1975a, and 1975b).
Eggs and fry were produced and remained in the exposure units.
-------
Hydroxyproline and backbone collagen were found to be sensitive bio-
chemical indicators of growth in fathead minnows exposed to toxaphen (Table
2). The correlation between hydroxyproline in backbone collagen and fish
weight was high, with a coefficient of determination (r2) of 0.982, and the
relation between collagen concentration and weight was even higher (r2,
0.990). Coefficients of determination were also relatively high for brook
trout and channel catfish. The measurement of hydroxyproline has some ad-
vantages over the measurement of collagen in that hydroxyproline can be
directly determined in eggs and whole fry, whereas collagen is determined
indirectly, except in fish large enough to permit analysis of the backbone
itself. The impact of toxicants on collagen and hydroxyproline metabolism
is probably greatest during early life stages of fish because the young fish
are generally more sensitive than older fish to toxicants, and have more
rapid developmental rates. However, in a preliminary study with the dime-
thylamine salt of 2,4-D and fathead minnows (Mayer and Mehrle, 1974), it was
found that hydroxyproline changed little in backbone collagen, but that col-
lagen itself decreased significantly (P<0.05). These results indicate that,
when possible, both hydroxyproline and collagen in backbone should be mea-
sured to facilitate toxicological interpretation.
The use of collagen and hydroxyproline as predictors of growth effects
shows some promise, but has not been fully delineated (Table 3). The reduc-
tion of growth caused by toxaphene in brook trout occurred 23 to 30 days
after effects were observed in hydroxyproline content. In the first study
(Mehrle and Mayer, 1975a), the growth, and the collagen and hydroxyproline
concentrations in adult fathead minnows were significantly reduced (P<0.05)
at all toxaphene concentrations. The hydroxyproline concentration in back-
bone collagen of adults was significantly reduced (PO.05) in toxaphene con-
centrations as low as 54 ng/1, whereas growth was significantly reduced only
in the 97 and 173 ng/1 exposures of a second study (Mayer &t <*£., 1977). In
the resulting fry, however, growth was more sensitive than hydroxyproline as
an indicator of the effects of toxaphene. Growth of channel catfish fry was
not reduced by toxaphene until 30 days after the eggs hatched, but the hy-
droxyproline content of eggs from exposed adults was significantly reduced
(P<0.05). The effects of toxaphene on hydroxyproline first appeared at ex-
posure concentrations of 72 ng/1, whereas effects on growth first appeared
at 299 ng/1. However, the reduction in hydroxyproline was related to bone
development, and numerous fish in the 72 through the 630 ng/1 exposures had
broken backs (Mayer e£ o£., 1977; Mehrle and Mayer, 1976). Also, survival
was significantly reduced (P<0.05) in concentrations of 792 to 630 ng/1 and
TABLE 2. RELATION BETWEEN BACKBONE DEVELOPMENT AND WEIGHT
IN FISH EXPOSED TO TOXAPHENE
P Value of
Coefficient of Determination (r2) Correlation
Species Hydroxyproline
Brook trout
Fathead minnow
Channel catfish
0.962
0.982
0.974
Collagen
0.964
0.990
0.982
Coefficient (r)
0.005
0.0005
0.0005
95
-------
TABLE 3. STATISTICAL SIGNIFICANCE3 OF THE EFFECTS OF TOXAPHENE ON GROWTH
AND HYDROXYPROLINE CONCENTRATION IN FISH.
Species, Life Stages
and Exposure Period
(Days)
Brook Trout
Fry
0
7
15
30
60
90
Fathead Minnows
Adults
98
Fry
30
Channel Catfish
Adults
100
Eggs
7
Fry
5
15
30
90
Exposure Concentration (ng/1) of
Toxaphene, and Statistical Significance
(1) or Nonsignificance (0) of Effects
on Growth (Left Column) and Hydroxy-
proline Concentration (Right)
39
,
0 -
- 0
- 0
0 1
1 1
1 1
13
0 0
0 0
49
0 0
- 0
0 -
0 0
0 -
0 0
68
0 -
- 1
- 1 -
0 1
1 1
1 1
25
0 0
0 0
72
0 0
- 1
0 -
0 1
0 -
0 1
139
0 -
- 1
- 1
1 1
1 1
1 1
54
0 1
1 0
129
0 0
- 1
0 -
0 1
0 -
0 1
288
0 -
- 1
- 1
1 1
c
c
97
1 1
1 1
299
0 0
- 1
0 -
0 1
1 -
1 1
502
0 -
- 1
- 1
c
c
c
173
1 1
1 1
630
0 0
- 1
0 -
0 1
1 -
1 1
Statistical significance, P=0.05.
Not determined.
CA11 fish had died.
96
-------
growth rates of the surviving catfish fry may not have been fully representa-
tive of the original populations.
EFFECT OF OR6ANOCHLORINE CONTAMINANTS ON COLLAGEN FORMATION
Various reports have stated that vitamin C is involved in the hydroxyla-
tion of drugs and chemicals in the liver of mammals (Axelrod qJL at., 1954;
Levin
-------
to
-5
n>
CL
o
X
Liver <^ Cl
oo
CU OJ
r+ c*
_i.
-t <
n> n>
_o
c n>
-i. 3
-S N
n>^
< tti
CU
3
O
Hydroxylating enzymes
vitamin C
Storage, further
degradation,
elimination
(Prolineand HVdroxylating enzymes Hydroxyproline
B°ne) lysine — anH
vitamin C
and
hydroxylysine
Collagen
n>
-5
DJ
3
Q.
O*
O
3
fD
-------
C in bone, however, was significantly reduced (P<0.05) in fish exposed to
toxaphene for 90 days, and was low in all fish, including the controls, at
150 days. This response in the controls was probably due to the chronic
effects of the diet itself. After 90 days exposure, backbone collagen was
significantly reduced (P<0.05) only in the highest concentration; at 150
days, however, it was significantly reduced in all toxaphene treated fish.
Thus, when fish are exposed to an organochlorine contaminant such as toxa-
phene, the increased use of vitamin C by the liver in hydroxylative detoxi-
cation mechanisms may reduce the amount in the bones by as much as 50 per-
cent. This reduction of vitamin C in bone is believed to inhibit the
formation of hydroxyproline from proline and reduces collagen formation.
CONCLUSIONS
Biochemical characteristics such as hydroxyproline and collagen concen-
trations in bone can be used as indicators, and within limits, predictors
of growth in fish thereby shortening chronic toxicity tests. Although
growth can be directly related to collagen and hydroxyproline metabolism
in fish, the mechanism by which growth is reduced is not known. Other
biochemical processes requiring vitamin C could also be affected when
large amounts of the vitamin are used by the liver in detoxication of
organic contaminants through microsomal hydroxylative enzymes.
ACKNOWLEDGEMENT
This research was sponsored in part by the U.S. Environmental Pro-
tection Agency through Contract No. EPA-IA6-0153(D) and EPA-IAG-141(D).
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101
-------
SECTION 10
EXPERIMENTAL TESTING OF TOXICITY OF WATER MEDIA AND INCREASING
OF THE SENSITIVITY OF BIOLOGICAL TESTS
L.P. Braginski, V.D. Bersa, T.I. Birger, I.L. Burtnaya,
F.Ya. Komarovski, A.Ya. Malyarevskaya, E.P- Shcherban
General increase of anthropogenic pollution of the hydrosphere raises
the problem of quantitative and qualitative characterization of pollu-
tants and evaluation of their biological danger. Additionally, the ques-
tion of establishing analytical and control methodologies of wide applica-
bility for assaying the toxicity of the water medium based on evaluation
of biological effects of toxicants is of paramount importance. One of
the principal ways of solving the problem is the application of biologic
tests. These tests have enjoyed wide spread acceptance in Europe
(Bringmann and Kuhn, 1959; Liebmann, 1960; Stanislawski, 1969), in the
USA (Katz, 1971; Environmental Protection Agency, 1972; Federal Water
Pollution Control Administration, 1969) and in the USSR (Braginski, 1971;
Lesnikov, 1971; Anon., 1959; 1971, 1966; Stroganov, 1971).
The main advantage of the biological tests are simplicity and avail-
ability of methodology, high sensitivity of the test organisms to the
minimum concentrations of toxic agents, speed, and the fact that expen-
sive reagents and equipment are not required. The main principle of bio-
logical testing is extremely simple. It is used to establish confident
differences between the experiment (medium containing toxicant) and the
control (clean water) in any indicative biological parameter test or-
ganism. Both alternative (life-death) and graded (percent of the experi-
ment as contracted with the control) experiments are used to indicate com-
plete or partial inhibition of essential functions of test organisms
under the influence of the test water or toxicants in certain concentra-
tions.
Discrimination between two types of test organisms is made: 1) indi-
cative, and 2) representative ones. The first category implies the use
of organisms with the greatest degree of sensitivity to toxicants, the
second implies the use of organisms that most fully represent a given
ecosystem (the crustacean Ep-uk
-------
(water, mud, vegetation thickets, etc.), and degree of sensitivity to
toxicants. The latter is specific for each species and varies within a
very great range: from nannograms to milligrams and grams per litre
(Alekseev, 1970). Investigations of both Soviet (Alekseev and Antipin,
1976; Filimonova, 1974; Anon., 1975) and the filter feeders (especially
the cladocerans) are the most sensitive test organisms, and hence, the
most widely used. Different species of Cladocera, however, have their
own specific sensitivity, and between the laboratory test cultures (pure
lines) and natural populations there are also substantially great dif-
ferences in resistence to toxicants.
In the world literature there are a number of biological tests sug-
gested. These have been described in appropriate reviews (Katz, 1971;
Stanislawska, 1969). Many of these tests have not received international
recognition, and are being used primarily within national laboratories, or
have been encorporated within the limits of regional agreements (Anon.,
1966).
The present communication deals mainly with the authors' efforts,
representing a contribution of one laboratory to the given problem. Since
algae are used only rarely for toxicological testing, the topic will be
discussed in some detail.
Test cultures of algae are grown either as pure cultures, or as
samples from natural waters at the time of mass development of some
species (e.g., 5'te.pka.no'dcicui ha.ntzc.kLL in spring, or Mic.noc.ytsit> aqjui-
Q4.no&a in summer). There may be essential differences between the re-
sults obtained on these species, since laboratory cultures are more deli-
cate, and have been developed under artificial conditions, frequently on
complex media. However, the benefits of uniform results and synchrony
cannot be overlooked.
The simpliest test is the one involving the death of the unicellular
algae in the presence of toxicants. A quantitative ratio of living to
dead cells in the test culture is established by means of microscopy.
Illustrative of this technique is the testing of various ions. For
example, a solution containing copper and ammonium is so toxic that even
at a concentration of 0.05 mg/1 active from the living cells is practi-
cally absent, and at 0.5 mg/1 no cells survived (Table 1).
TABLE 1. RESULTS OF THE TOXICITY OF A COPPER-AMMONIUM SOLUTION ON
TEST CULTURE OF MICROC/STIS AERUGIWOSA (LABORATORY STRAIN,
5-DAY EXPERIMENT)
Concentration
mg/1
Control
0.05
0.1
0.5
1.0
Living
68
0
0
0
0
Percentage
Dead
32
10.7
96.8
100
100
Dying
0.0
89.3
3.2
0
0
103
-------
The method of low level luminiscence developed by a group of Moscow bio-
physicists enables an evaluation of the toxicity of the minimum concentra-
tions of toxic agents.
In addition to the luminiscence method, differentiation of living and
dead algal cells may be performed with a help of dyes (0.1 percent
neutral red), reagents (TTX) and fluorochromes.
Occasionally, dead and living algae may be discerned even when using
vital microscopy, or the dark field technique. These techniques reveal
tological disturbances (plasmolysis, disintegration of the chromatophore
and cell walls) under the action of toxicants in large filamentous algae
Cladophoia., RkizocJLoYu.u.m and others (Braginski, 1972; Anon., 1959). The
Institute of Biology of the Ukranian Academy of Sciences has proposed a
number of cytochemical tests which enable observation of disturbances of
the living activity of blue-green algae in a toxic medium. These methods
involve determination of permeability of cell membranes when staining
with nitrosine, determination of ascorbic acid content, sulphur hydryl
group determination, measurement of enzymatic activity of cells, and ob-
servation of the redox potential of test cultures (Osterov, 1968).
Advances in understanding of the physiology of algae allow the use of
a whole complex of experimental methods including the celloscopic counting
of cells in test cultures, the determination of the chlorophylls, caro-
tenoids and other pigments (Anon., 1975). All these tests are important
since they take into account the possible harmful effect of toxicants,
not only upon animals, but also upon the components of primary production
of waters. To date, insufficient attention has been given to the primary
producers when ascertaining the ecological threat of chemical pollutants.
In this regard, of substantial interest are the investigations for direct
determination of inhibition of photosynthesis (Knepp, 1969).
The authors proposed a number of modifications of the Knepp test,
including the use of Sce,ned&smu4 acuuninatuA, S. bjugatuA, and the diatom
Ste-pkanociibcuA kantzckli (Bereza, 1972, 1973) instead of 5. quadfUcaadu^
as test species, and a modification of the oxygen method for determina-
tion of primary production and destruction of phytoplankton in the pre-
sence of toxicants. These modifications differ from the traditional
method in that a toxic component of required concentration is added to
the bottles as a control. The experiment then determines primary pro-
duction and destruction by the generally accepted method (Winberg,
I960). Investigations have shown that this test is not universal and
may have an indicative value when dealing primarily with ions of heavy
metals.
Under the action of heavy metals, the correlation of primary produc-
tion and destruction is vastly altered. Copper, for example, influences
both processes, while zinc increases destruction (Table 2).
In investigations of pollution of water by stable pesticides capable
of inhibiting phytoplankton photosynthesis, the test in which the inten-
104
-------
TABLE 2. PRIMARY PRODUCTION AND DESTRUCTION IN PHYTOPLANKTON
SAMPLES (BLUE-GREEN + DIATOM) UNDER THE ACTION OF HEAVY
METALS. VALUES INDICATED AS PERCENT OF THE CONTROL
Concentration
of Ions, mg/1
MO'2
5.10'1
1.10"1
1.10'3
MO'2
i.io-1
1.0
Photosynthesis
10.6
17.3
4.2
58.7
116.0
102.0
31.4
Destruction
38.8
20.0
11.1
120.0
180.0
550
450.0
105
-------
sity of gross photosynthesis is determined, may be a very distinct indica-
tor of preservation of toxic substances in the water (Table 3).
In water containing herbicides, photosynthesis is inhibited in all
algae, including the filamentous forms which may be used as indicators
for visual detection of substances inhibiting photosynthesis. The test
is performed in vessels with small lumps of filamentous algae (5-10 g wet
weight), which, in a toxic medium, settles to the bottom, and does not
become covered with bubbles of oxygen, but in the control they remain sus-
pended in water and actively liberate oxygen. This test may be also
performed on a glass slide with solitary filaments of Cladopkoia, or in a
concave slide with a suspension of any planktonic algae. In both cases,
the presence of toxic factor is indicated by the absence of formation of
the bubbles of oxygen.
The most useful indicator organisms are aquatic invertebrates since
they are more sensitive to toxicants than algae. Cladocerans, rotifers,
larvae of mayflies and chironomids, garmarids, isopods, copepods, ostra-
cods, bivalvia and gastropod molluscs are routinely used. Each of these
organisms have their own specific features of behavior, biology, and
reactions to toxic materials which must be taken into account when per-
forming experiments.
TABLE 3. GROSS PHOTOSYNTHESIS IN A WATER-BODY TREATED WITH DIURONE
AT THE TIME OF A BLUE-GREEN ALGAL BLOOM.
Days of
Experiment
Background
Temperature
°C
21
Experiment
(Diurone 0.2 mg/1 )
02, Percent of
Saturation
111.8
Photosynthesis
mg/1 hr-1
12.0
Control
02, Percent of
Saturation
224.2
1 hour after
addition 28
1
3
5
7
10
12
15
23
25
23.5
23.0
23.0
19.5
17.0
57.2
0
0
0
0
0
0
0
0
0
0
0
0
0
0
124.5
152.8
165.2
71.8
66.0
109.8
154.0
106
-------
In common tests on survival, the reaction to overwhelming intoxication
is identical in all species, i.e., death. The percentage of mortality,
for a given time may, however, vary significantly depending on the sensi-
tivity of organisms, their anerobiotic ability, their lipid content, the
degree of oxygen saturation of the medium, and many other factors. In
addition to studies of mortality, visual behavioral reactions are also
quite indicative. At present, experience in analysis of behavioral reac-
tions in organisms is limited to two classes of toxicants, pesticides and
heavy metals. It is possible that the reactions described are not univer-
sal, and the action of other toxicants is manifested in a different way.
For example, the well known peculiarities of behavior of aquatic organisms
in the presence of phenol (Alekseev, 1970) differs essentially from those
described below for a number of species.
Cladocerans in a non-toxic medium move by leaps, rarely settling onto
the bottom. The movements of the antennae are even. The heart rhythm in
different species varies from 200 to 300 (occasionally up to 500) beats
per minute, the eyes are brightly pigmented, and the body is multicolored.
In a toxic medium the movements are predominantly rotatory and revolving
around the body axis or along a spiral. The latter is especially
characteristic reaction to toxic pesticides. As intoxication progresses,
the crustaceans lie immobile, the body contracts convulsively, the
antennae jerks, the heart beat diminishes to a single uneven contraction
per minute, and the eyes become depigmented. The body may acquire a red
coloration. The females abort immediately after transfer into the toxic
medium, shedding both eggs and embryos. After prolonged exposures to low
concentrations of toxicant, embryonic abnormalities may arise. These
anomolies may include twisted antennae, underdeveloped eyes, and the like.
The copepods do not usually manifest symptoms of disturbed behavior
in toxic media. The rotifers, however, pass into a state of anabiosis,
change in body length, and cease to feed (observable microscopically by
decoloration of the intestine). Counts of living, anabiotic, and dead
rotifers in a Fuks-Rozental Counting Chamber or other hemocytometer esta-
blish the toxic effect quantitatively in comparison with the control.
In toxic media, the usually quick moving gammarids (Gommo^uxS pu£ex,
6. IcLCiLAtAsU) become sluggish and nearly immobile. Under the influence
of heavy metals, their bodies may acquire a red tint. Aquatic sowbugs
react similiarly.
Chironomid larvae in water move in a spiral fashion. On a suitable
substrate, they begin case construction. Under the influence of toxic
substances, their body is convulsively stretched and straightened, they
lie immovable on the bottom and fail to construct cases. In hemoglobin
containing species (Ch&ionomuA p£cuno4oi, C. Aw) the red color
may assume a greenish blue, or disappear altogether.
The image forms of chironomids appear to be rather sensitive to the
presence of DDT residues in storage organs and tissues of fishes (Bereza,
1972). In contact with tissues of predatory fishes (fat, brain) con-
taining considerable quantities of accumulated DDT, chironomids are para-
107
-------
lized instantly, or they manifest clear symptoms of insectides poisoning:
tremor of limbs and wings, disturbances of coordination, convulsions, and
death in 20-30 min. after exposure. To check corresponding symptoms of
intoxication and the rate of their development relative to the DDT content
in tissues, the material was analyzed by the gas chromatographic method
and scale of conventional values of the level of accumulation of DDT in
tissues of fish is proposed (Table 4).
This test has a dignostic value when analyzing the causes of mass
mortality of fish. Accumulation of DDT i.e. the vital organs of fish, es-
pecially in the brain, may be the cause of sudden catastrophic death in
stress situations, e.g., after a sharp increase in water temperature,
during spawning, and as a consequence of mobilization of fat reserves and
resultant appearance of DDT dissolved in lipids in the blood streams.
Detection of insecticides in the brain tissue-of long preserved, or even
petrified fish, enables identification of the role of DDT as one of the
causes of death, while the chemical analysis of such tissue is very com-
plicated, and requires equipment for gas chromatography.
The oligochaet Tubx^ex tab^nx. in non-toxic media normally maintains
a vertical body attitude, swaying evenly like a wheat field in the wind.
On the bottom, they form tangles. In toxic media the bodies of the worms
stretch convulsively, the movements become disordered, and under deep in-
toxication, the worms lie immobile, unentangled on the bottom, and the
reddish coloration of the body disappears (evidently as a result of hemo-
globin degradation as in chironomid larvae).
TABLE 4. TESTS ON DDT CONTENT IN STORAGE TISSUES OF FISH
Species of
Fish, Tested
Tissue
Weight
g
DDT Content
(DDT+DDD+DDE)
mg/kg
Toxic Effect
On Image of
Chironomids
Mark
Zander
Intestinal 4.0
fat
Liver 4.0
25.7
0.70
Instant death of ++++
all the insects.
Clear symptoms of +++
intoxication. Agony.
Death in 20-40 min.
Carp
Liver
4.0
0.145
30-40 percent are
dead in 1 hour.
Zander
Muscle Tissue 4.0
Roach
Muscle Tissue 2.0
0.03
0.00
Death in 1 hour.
All are living for
6-8 hours.
108
-------
In clean water, the molluscs hnodonta, Un-io, and Spka&uum periodi-
cally open the valves and protrude both foot and syphons. In toxic media,
the valves are either permanently closed or widely opened and the syphons
are protruded. In the gastropods, the reactions to toxicants are quite
diverse and peculiar. U-iv^paAu^ vlv^paAuA in toxic median retracts the
body deeply into its shell, tightly shuts the lid and becomes enveloped in
a thick layer of musilage. The large pond snail, Lcmnaea &tagnaLu>, ac-
tively grazes on plants and filters suspended particles in clean water,
its body periodically protruding from its shell. The defensive reflex
consisting of retracting the body into the shell as a result of external
stimuli is clearly expressed in this mollusc. It normally produces a
tape-like excrement, however, the water in the vessel remains clean and
transparent, evidently owing to bactericidal effect of the excreted musi-
lage and possibly to antibiotic substances. In toxic media, this mollusc
does not feed, or the intensity of ingestion of food is greatly dimini-
shed, and the body is hidden in the shell. Other essential disturbances
in behavior, notably in sexual behavior (Bereza, 1973), are evident.
Under acute intoxication, this mollusc falls out of its shell.
The sensitivity of tests on survival of invertebrates may be consider-
ably increased when experiments are conducted at elevated temperatures
(Table 5). The sensitivity of biological tests is further increased with
sharp changes in temperature.
When undertaking such testing, it is necessary to consider a number
of factors, including: 1) hydrochemical peculiarities of water, its
oxygen content, pH; 2) the degree of adaptation of the test organisms to
the experimental conditions, their lipid content, age, sex, developmental
stage; 3) temperature; and 4) the degree of pollution of the habitat.
However, in cases without significant statistical differences between
the experiment and the control, and the test is of short duration, all of
these conditions are of secondary importance, especially when drawing con-
TABLE 5. INCREASE IN SENSITIVITY OF BIOLOGICAL TESTS ON SURVIVAL
OF AQUATIC ORGANISMS AT 30 °C
Toxicant
ZnS04-7H20
CdS04
CuS04'5H20
ZnS04-7H20
ZnS04-7H20
CdS04
Lethal Concentration, mg/1
Test Organisms
Vaphvua. magna
(females with eggs)
Vaphnia magna
(young)
Aie££a4 aqucuttcuA
Aie££o6 aqao^ccoa
at 18 °C
1.0
0.2
0.02
1.09
39.0
32.0
at 30 °C
0.005
0.002
0.009
0.014
1.52
0.20
109
-------
elusions about the toxicity of the water. Further evaluation requires
special chemical analyses, toxicologic investigations, and a detailed
ecological study of the water body.
The quantitative aspect of application of biological testing is more
complex and requires an approach to the characteristics of measurable
functions. Any function changing under chemical action may function as
an indicator, although the more simply measured functions are advanta-
geous to use. Various functional disturbances in highly organized aquatic
organisms may be evaluated on the basis of application of various physio-
logical and biochemical methods (Komarovski, 1971, 1972; Malyarevskaya
and Birger, 1973).
At present, elementary statistical principles are being used in
biological testing. However, it is impossible to avoid the influence of
permanently acting factors (time, temperature) and yet, it is necessary
to reduce expenditures of time and labor to a minimum. The most reason-
able approach is the use of the principle of the All Factor Experiment
(AFE). The scheme AFE 22 enables the acquisition of reliable data from
four experiments. This seems applicable to the situation of stabilized
temperature, with due consideration of the factors of concentrations, and
time of action of toxicants. When three variable factors are considered,
the scheme AFE 23 gives confident information using eight experiments.
Such a material can be easily interpreted both analytically and graphi-
cally, using a system of three coordinates. With a mass accumulation of
data an electronic computer can be utilized.
In conclusion, it should be noted that the development of the method-
ology leads to three variations of application: 1) qualitative tests for
toxicity of medium, 2) quantitative characteristics of toxic effect, and
3) quantitative determination of toxicants. The latter task is the most
complicated, and practically insoluble. In some cases, biological
testing lacks sufficient sensitivity when compared with chemical analysis.
For example, the test of Knepp compared with analytical chemical methods
has shown that this test detects the presence of some toxicants only at
concentrations equal to 5-10 MPC, which is certainly insufficient for a
quantitative conclusion.
More indicative are the specific chemical tests, which reflect the
effect of a given substance, and provide a general indication of the
polluting substance. Concentrations are established by consecutive di-
lutions of the suspected toxicant. Parallel testing using a sensitive
species is employed. The effective concentration corresponds to the
dilution at which similar toxic effects are displayed. In view of the
difficulties of identifying numerous substances in natural water, and the
necessity for complex and expensive apparatus, often unsuitable for work
in the field, the quantitative tests may not only have analytical value,
but may also be more economic than direct determination of toxicants by
chemical methods.
110
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REFERENCES
Alekseev, V.A. 1970. Study of acute phenol intoxication in some species
of aquatic insects and arachnids. Gidrobiol. zh. V. 6, N. 5.
Alekseev, V.A., B.N. Antipin. 1976. lexicological characteristic and
symptom complex of acute phenol intoxication in some freshwater crust-
aceans and molluscs. Gidrobiol. zh. V. XII, N. 2.
Anon. 1965. Bioluminiscence. M. V. 21.
Anon. 1959. Life in the USSR freshwaters. V. 4.
Anon. 1971. Methods of biological investigations in aquatic toxicology.
M.
Anon. 1975. Methods of physiological and biochemical investigations of
algae in hydrobiology. Kiev.
Anon. 1975. Tin containing organic compounds and living processes in
hydrobionts. M.
Anon. 1966. Unified methods of investigations of quality of water. In
coll.: Metody biol. i mikrobiol. analiza vod. M., Part 4.
Bereza, V.D. 1972. Using of culture of Sce.ne.d&>mtu b^jagcutu^ (Turp)
Kutz. for evaluation of toxic effect of pollutants on phytoplankton
with a help of A-Z-test by Knepp. In Coll.: Experimental Aquatic
Toxicology. Riga, N. 3.
Bereza, V.D. 1973. Ste.pha.no cLL& coi kantzckU. Grun. as test-object for
indication of pollution of waters by sewage containing toxic sub-
stances, In Coll.: Experimental Aquatic toxicology. Riga, N. 4.
Braginski, L.P. 1971. On some principles of choosing test-objects for
studies in aquatic toxicology. In Coll.: Criterion of Toxicity and
Principles of the Methods in Aquatic Toxicology. M.
Braginski, L.P. 1972. Pesticides and life in water bodies. Kiev.
Bringmann, G., R. Kuhn. 1959. Vergleichende wasser-toxicologische
Untersuchungen an Bacterien, Algae und Kleinkrebsten, gesundkeit
ing., V. 80.
Crosby, D.G., R.J. Tucker. 1971. Accumulation of DDT by Vaphnia magna.
Environ. Sci . and Technology, V. 5, N. 8.
Federal Water Pollution Control Administration. 1969. Water quality
criteria. Rep. of the Nat. Techn. Advisory Committee to the Secre-
tary of the Interior, April 1, 1968. Wash., D.C.
Ill
-------
Filimonova, V.I. 1974. The influence of baitex on some representatives
of aquatic fauna in the Karelia. Gidrobiol. zh., V. 10, N. 3.
Frear, D.E., J.E. Boyd. 1967. Use of Vaphn^a magna for the microbio-
assay of pesticides. I. Development of standardized techniques for
rearing Pap/ixtui and preparation of dosage-mortality curves for pesti-
cides. J. Entomol., V. 60, N. 5.
Katz, M. 1971. Toxicity bioassay technique using aquatic organisms.
Wat. and Wat. Pollution Handb. N.Y., V. 2.
Knepp, 6. 1969. Investigations of self-purifying capacity of rivers and
its disturbances by industrial sewage. In Coll.: Limnological In-
vestigations of the Danube. Kiev.
Komarovski, F.Ya. 1972. Experimental investigations of the toxicity of
the complex substances "Monurox" for fishes under the conditions of
long-term experiment. In Coll.: Experimental Aquatic Toxicology.
Riga, N. 3.
Komarovski, F.Ya., N.A. Popovitch. 1971. Investigations of the toxic
effect of urea derivatives on fishes under the conditions of long-
term experiment. In Coll.: Experimental Aquatic Toxicology. Riga,
N. 2.
Liebmann, H. 1960. Handbuch der Frishwasser und Abwasser Biologie.
Munchen-R. Oldenburg.
Lesnikov, L.A. 1971. Methods of evaluation of the influence of water
from natural water-bodies on Vaphnia. magna.. In Coll.: Metodiki
biol. issled po vodnoi toksikol., M.
Malyarevskaya, A.Ya., T.I. Birger, at at. 1973. The influence of the
blue-green algae on metabolism in fishes. Kiev.
Osetrov, V.I. 1968. On the application of cytochemical methods in
studies of the living activities of the blue-green algae. In Coll.:
Bloom of Water. Kiev, Issue 1.
Stanislawska, J. 1969. Zastosowanie biotestow do wydrywania substancji
chemichnych w wodzie. Ecol. Polska, V. 15, N. 3.
Stroganov, N.S. 1971. Methods of determination of toxicity of water
medium. In Coll.: Methods of Biological Investigations in Aquatic
Toxicology. M.
United States Environmental Protection Agency. 1972. Pesticides in the
Aquatic Environment. Wash., D.C.
Winberg, G.G. 1960. Primary production of waters. Minsk.
112
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SECTION 11
CHRONIC EFFECTS OF LOW LEVELS OF HYDROGEN SULFIDE
ON FRESHWATER FISH
Lloyd L. Smith, Or.
INTRODUCTION
The potential effects of hydrogen sulfide on fishery ecosystems have
not been fully realized because most survey work has measured neither con-
centrations below 0.5 mg/1, nor accumulations near the sediment/water
interface in areas of continuous H2$ production. Areas within a few
centimeters of the bottom where fish eggs and young fry occur are rarely
sampled. Since it has been assumed that levels of hydrogen sulfide po-
tentially dangerous to fish life occur only under conditions of low oxygen
concentration, it has been believed that the adverse effects of low oxygen
will control fish populations before the toxic effects of sulfides will
be manifested. Undissociated hydrogen sulfide is the toxic form. At pH
9.0, approximately 1 percent is undissociated, at 6.7 about 50 percent,
and at 5.0 about 99 percent.
The significance of the work described in this study is primarily a
demonstration of the toxic effect of very low concentrations of hydro-
gen sulfide which frequently are found over natural organic bottoms, in
the vicinity of sludge beds, and in areas where hydrogen sulfide is formed
from waste effluents or comes directly from industrial operations. Colby
and Smith (1967) found levels of hydrogen sulfide within 20 mm of the
bottom which varied from 0.02 to 0.2.0 mg/1 in a river with a major wall-
eye tetj.zoAte.dion vWiuim vi&L&um Mitchell) fishery. Dissolved oxygen at
these locations was adequate to maintain fish life. Here maintainance of
the adult population depended on inward migration of fish rather than
natural reproduction. In natural spawning areas for northern pike ( E4ox
£ucxu6 Linnaeus), Adelman (1969) reported hydrogen sulfide concentrations
near the bottom commonly in the range of 0.03-0.08 mg/1, and occasionally
as high as 0.22 mg/1 during the spawning period. Scidmore (1956) working
in a Minnesota lake during the winter found 0.3 and 0.4 mg/1 H S with 6.0
and 3.6 mg/1 0 , respectively. Adelman and Smith (1970) showed that eggs
and fry of northern pike were affected by low levels of hydrogen sulfide.
The experiments summarized in the present report used four species of
fish, brook trout (So£ve£oio6 &owUnati& Mitchill); bluegill (Lepom-a
mac/iocA-i/tai Rafinesque); fathead minnow (Vixnuphatoj* p/iome£aA Rafinesque);
and goldfish (Caficu>A-uu auAatiu Linnaeus). The purpose of this study was
113
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to determine "no-effect" levels of hydrogen sulfide based on long-term
tests, and to relate them to acute toxicity levels.
MATERIALS AND METHODS
The work described here was done at the University of Minnesota labor-
atories using well water (Table 1). Continuous flow-through apparatus was
used in all tests. Small species and early life history stages were
tested in equipment described by Colby and Smith (1967), and Mount and
Brungs (1967). Adult fish of larger species were tested in fiberglass
tanks and all other tests were done in glass or acrylic test chambers.
Molecular hydrogen sulfide concentrations were maintained with sodium sul-
fide solutions, and pH was adjusted to provide the desired test concentra-
tion. Analyses of water from the center of each test chamber were made 3
times each day in acute tests and every 2 days in chronic tests. Using
Method C of Standard Methods of Water Analysis (1971), undissociated
hydrogen sulfide was determined by calculation. Bluegills and some fat-
head minnows were wild stock from local lakes or streams. Other fish
were hatchery stock or laboratory reared. The 1X50 values were cal-
culated by standard methods using probit techniques.
Acute Toxicity
Acute tests were made with 4 to 5 concentrations of hydrogen sulfide
arranged in a logarithmic series and one control. Temperature in various
tests ranged from 6.1 °C to 26 °C depending on species, life history
stages, and specific objectives (Table 2). Duration of tests was from 4
to 12 days to determine the 96-hr LCso and the threshold LC$Q. The
threshold value was considered to be attained when no death occurred in
48 hours. Eggs, sac fry, swim-up fry, juveniles, and adults of fish
species were tested. The LC5Q at 96-hr varied from 0.515 mg/1 H£S at
6.1 °C with fathead minnow sub-adults, to 0.007 mg/1 H2S at 24 °C with
fathead fry. Values ranged between these limits for other species, life
history stages, and temperatures (Table 2).
The effect of temperature on resistance was large in fathead minnows.
The 96-hr LC$Q concentration increased from 0.021 mg/1 at 24 °C to 0.515
mg/1 at 6.1 °C (Table 2). This twenty-five fold change in tolerance
occurred principally between 10 and 6.1 °C. Goldfish tested at tempera-
tures between 14.1 and 26 °C showed 96-hr LC$Q values varying from 0.145
mg/1 H2S at the lowest, to 0.063 mg/1 ^S at the highest temperature
(Adelman and Smith, 1972).
Threshold LC5Q concentrations were lower than those for 96-hr in most
tests, but not markedly so, except in the goldfish at lower temperatures.
Threshold values for the various life history stages had the same general
relationships to each other as the 96-hr LCso values.
Long-Term Tests at Low Concentrations
Since acute toxicity of a material may be a poor index of long-term
effects of sub-acute concentrations of a toxicant on a fish population, a
114
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TABLE 1. ANALYSIS OF WELL WATER9 VALUES EXPRESSED IN
MG/L
Item Concentration
Total hardness as CaC03 220
Calcium as CaC03 140
Iron 0.02
Chloride <1
Sulfate <5
Sulfide 0.0
Fluoride 0.22
Total phosphates 0.03
Sodium 6
Potassium 2
Copper 0.0004
Manganese 0.0287
Zinc 0.0044
Cobalt, nickel <0.0005
Cadmium, mercury O.0001
Ammonia nitrogen 0.20
Organic nitrogen 0.20
aWater was taken from well head before aeration and heating;
pH 7.5.
115
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TABLE 2. 96-HOUR AND THRESHOLD LCso OF HYDROGEN SULFIDE (MG/L) FOR
BROOK TROUT, BLUEGILL, FATHEAD MINNOW AND GOLDFISH
Species
Brook
Trout
Bluegill
Fathead
Goldfish
No. of
Stage Tests
Egg
Sac fry
Swim-up fry
Juvenile
Egg
Sac fry
Swim-up fry
Juvenile
Adult
Egg
Fry
Sub-adult
Sub-adult
Sub-adult
Sub-adult
Sub-adult
Egg
Fry
Sub-adult
Sub-adult
Sub-adult
3
2
1
2
2
1
1
6
8
6
o
4
1
1
7
6
1
1
21
21
21
Temp.
°C
9.0
9.0
12.5
8.0
22.0
22.0
22.0
20.0
20.0
24.0
24.0
6.1
10.0
15.0
20.0
24.0
22.1
21.6
14.1
20.0
26.0
96-hr LCso
Mean (mg/1)
_ «•
0.031
0.022
0.025
0.140
—
0.009
0.028
0.030
0.035
0.007
0.515
0.150
0.057
0.036
0.021
0.022
0.025
0.145
0.083
0.063
Threshold
Mean (mg/1)
0.054
0.030
0.019
0.019
__
0.017
0.008
0.028
0.030
0.035
0.006
—
--
--
—
--
_-
—
0.084
0.071
0.060
LC50
Days
9
5
7
12
__
9
8
4
4
4-8
6
—
--
--
--
--
—
—
11
11
11
116
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series of tests of extended duration at low levels of hydrogen sulfide were
run on the test species (Table 3). An unfavorable response was assumed to
occur when growth, survival, or reproduction were adversely effected. A
"no-effect" concentration was identified when no adverse effect on these
parameters was noted. Physiological responses may occur at these concen-
trations. In some species stimulation occurred at the lowest levels of
treatment and resulted in better long-term performance than that exhibited
by the controls.
Tests were conducted for 45 to 826 days in the various species (Table
3). The temperature at which various tests were run varied from 11.8 to
24 °C with different species. The "no-effect" concentration varied from a
minimum concentration of <0.001 mg/1 ^S at 24 °C in bluegills, to 0.010
mg/1 H2S at 18.6 °C in one goldfish test. Bluegills were the most sensi-
tive species. A comparison of trout with the warmwater species is diffi-
cult because the former were tested at lower temperatures. The life his-
tory stage at which different fish species were first subjected to hydrogen
sulfide had varied influence on the final "no-effect" level in the differ-
ent species. The concentrations designated as "lowest effect level" were
the lowest concentrations of molecular fS which showed a measurable
TABLE 3. CHRONIC EFFECT OF SUBLETHAL CONCENTRATIONS OF HYDROGEN SULFIDE
(MG/L) ON TROUT, BLUEGILL, FATHEAD MINNOW AND GOLDFISH
Species
Brook
Trout
Bluegill
Fathead
Goldfish
Stage
Started
Adult
Fing.
(5 g)
Eggs
Juvenile
Juvenile
Adult
Adult
Adult
Eggs
Eggs
Juvenile
Juvenile
Eggs
Juvenile
Adult
Days
45-751
120
316
826
165
200
288
97
84
404
112
373
430
294
294
Temp.
°C
12.9
13.0
22.4
11.8
24.0
15.0
20.2
23.6
23.0
24.0
20.0
21.3
21.5
18.6
18.6
No-Effect
Level
<0.006
0.007
o
0.002^
<0.001
—
0.003
<0.001
0.005
0.004
0.005
0.007
0.007
0.010
0.005
Lowest
Effect
Level
0.006
0.009
0.002
0.003
0.002
0.002
0.006
0.001
0.007
0.007
0.011
0.019
0.009
0.025
0.010
Factor
of
Effect3
R
G
G+S
G+S
G+S
G
G+S
R
G+S
G+S+R
S
G+S
G+S
G+S
G+S
1
Two tests.
^Reproduction inhibition at 0.002 mg/liter.
R, reproduction; G, growth; S, survival.
117
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adverse effect. The two most useful indicators of adverse effects were
growth and reproductive rates (Tables 4, 5, and 6).
Two experiments on bluegills were conducted, one started with young-of-
the-year fish exposed for 826 days, and one with adults exposed for 97 days
prior to spawning. At low H2S levels, the percentage increment in weight
after short exposure was greater than controls (Table 4). After 826 days
exposure to 0.007 mg/1 ^S, the mean weight of the fish started as young-
of-the-year was approximately 63 percent of controls. In a second test,
after fish started as adults which were exposed for 97 days to levels of
TABLE 4. INCREMENT IN WEIGHT AND SURVIVAL OF BLUEGILL STARTED AS
YOUNG-OF-THE-YEAR AND EXPOSED TO VARIED CONCENTRATIONS OF HYDROGEN
SULFIDE FOR 826 DAYS AT 02 OF 6.0 MG/LITER AND MEAN TEMPERATURE
OF 17.8-18.5 °C VARIED SEASONALLY—WEIGHT EXPRESSED AS MEAN
IN GRAMS AT SUCCEEDING INTERVALS
ExposureH2S Concentration (mg/liter)
(Days) Control 0.002 0.004 0.007
Weight (g)
56 4.02 4.50 5.00 5.12
392 50.81 46.34 34.37 43.35
826 99.91 98.71 90.35 63.05
Percentage Survival
28
362
392
420
826
100
100
100
100
100
100
100
100
100
100
90
90
90
90
90
100
100
100
100
70
TABLE 5. WEIGHT AND REPRODUCTION OF ADULT BLUEGILLS EXPOSED TO VARIED
CONCENTRATIONS OF HYDROGEN SULFIDE FOR 97 DAYS AT 23.5-23.9 °C
Concentration (mg/1)
Control 0.001 0.002 0.003 0.008
Mean Starting Weight
(g) 75.9 84 78.2 76.1 76.6
% Increase in Weight 55 58 70 59 54
No. Eggs/Female 17,562 12,795 6,157 0 0
No. Eggs/g of Female 155.5 100.8 51.1 0 0
__ _
-------
TABLE 6. GROWTH OF FATHEAD MINNOW IN 112-DAY EXPOSURE TO HYDROGEN
SULFIDE AT 23 °C, pH 7.7, Q2 6.4-7.3 MG/LITER - EXPRESSED
AS MEAN WEIGHT IN GRAMS OF THREE REPLICATIONS AT THE END
OF SUCCEEDING 28-DAY PERIODS
28 Days
56 Days
84 Days
112 Days1
Total % Increment
Survival %
No. Eggs/Female2
No. Spawning
Total Females
112 Days
140 Days
168 Days
252 Days
o
Total % Increment
Survival %
No. Eggs/Female
No. Spawnings
No. Females
Control
0.106
0.455
0.685
1.234
1,064
97
1,181
92
11
Control
0.616
0.965
1.249
2.160
250
97
912
116
21
First
0.0004
0.111
0.421
0.688
1.114
904
90
1,332
80
13
Second
0.0007
0.387
0.799
1.167
2.374
513
93
1,130
101
14
Generation
0.0012
0.116
0.382
0.595
1.066
819
93
444
30
9
Generation
0.0013
0.649
0.960
1.222
2.190
237
90
413
88
16
0.0031
0.106
0.399
0.714
1.180
1,057
93
1,314
60
9
0.0037
0.533
0.812
1.079
2.022
280
90
535
39
9
0.0061
0.115
0.359
0.610
1.097
854
50
1,109
24
4
0.0069
0.391
0.655
0.830
2.036
423
64
799
30
5
1
2$tart of spawning in first generation
JVfter 297 days exposure to H2S.
After 404 days exposure to H2S.
-------
hydrogen sulfide up to 0.008 mg/1, there was no appreciable difference be-
tween treatments and controls, except at 0.002 mg/1 where there was a
significant increase in growth. Reproduction (number of eggs deposited
per gram of female) in the second experiment was significantly reduced at
0.001 mg/1 H2S and completely inhibited at 0.003 and 0.008 mg/1 H2S (Table
5). In the experiment started with young-of-the-year no spawning occurred
after 826 days at concentrations of 0.002 mg/1 and higher. The failure of
egg deposition appeared to be caused by the inhibition or absence of normal
spawning behavior, since apparently average numbers of viable eggs for fish
of comparable size were found in ovaries of non-spawning fish.
A two-generation, long-term experiment on fathead minnows (three repli-
cations) was also run to determine the "no-effect" levels of hydrogen sul-
fide. the test was started with eggs which were hatched and carried
through to spawning adults. The second generation was continued with eggs
from females reared in the same hydrogen sulfide test levels as the first
generation. The first generation was continued for 297 days and the second
for 404 days. Growth measurements were all made prior to the start of
spawning in both generations. The day length was reduced and lengthened
during the second generation to induce spawning, which resulted in a
lengthening of the total exposure period. Hydrogen sulfide ranges in the
first generation were 0.0004-0.0061 mg/1, 0.0007-0.0069 mg/1 during the
first 112 days of the second generation, and 0.007-0.0069 mg/1 during the
remainder of the cycle. For both cycles, mean temperature was 23 °C, pH
7.7, and 0 6.4-7.3 mg/1 (Table 6).
Growth in weight after 112 days in the first generation was less at
all test levels. In the second generation, after 252-day exposure, there
was an apparent growth stimulation at 0.0007 mg/ H2S. Growth inhibition
in the second generation occurred in the early periods, but was less in
later periods, suggesting that early effects of exposure are greatest, and
tend to lessen as growth proceeds. These results may be influenced to
some degree by greater mortality of smaller fish in the hydrogen sulfide
treatments. Survival was much lower than in the control at the highest
treatment in both generations. In lower treatments, survival was also
lower than the control, but not markedly so.
The success of spawning, measured by the number of eggs produced per
female, did not appear to be affected in the first generation at any level
of hydrogen sulfide treatment, although at the highest level, female sur-
vival was substantially lower than in the control. In the second genera-
tion, there was an apparent stimulation of egg production at 0.0007 mg/1,
but a reduction at higher levels. The number at the highest level may
have been increased by mortality of smaller females.
DISCUSSION
Both toxic and long-term toxic concentrations of hydrogen sulfide have
been shown to be lower than levels commonly found in natural and polluted
waters. Because sites of sustained hydrogen sulfide concentrations in the
ecosystem are frequently overlooked, or low levels are not identified, the
120
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importance of this toxicant in potential fish producing waters is often not
evaluated. Comparison of LC5Q levels with those which have adverse effects
after long exposure indicates that the 96-hr LC$Q may be 3 to 8 times
higher than the safe levels. The present work, and that of Smith and Oseid
(1974), who examined the effect of hydrogen sulfide on early life history
stages of 8 species of freshwater fish, show that a safe level of hydrogen
sulfide which will insure survival and growth of a fish population, and
adequate survival of all life history stages will generally be between
0.002 and 0.004 mg/1 at 20 °C. In bluegills, the level is significantly
lower, and this pattern may be followed in other species not tested.
REFERENCES
American Public Health Association, American Water Works Association, Water
Pollution Control Federation. 1971. Standard methods for the examina-
tion of water and wastewater. 13th ed. New York, Am. Pub. Health
Assoc., Inc. 874 p.
Adelman, I.R. 1969. Survival and growth of northern pike (Eiox £ucx:a6 L.)
in relation to water quality. Ph.D. Thesis, Univ. of Minnesota, St.
Paul, Minnesota. 195 p.
Adelman, I.R. and L.L. Smith, Jr. 1970. Effect of hydrogen sulfide on
northern pike eggs and sac fry. Trans. Amer. Fish. Soc. 99: 501.
Adelman, I.R. and L.L. Smith, Jr. 1972. Toxicity of hydrogen sulfide to
goldfish (Co^o64xxc6 ausiatuA) as influenced by temperature, oxygen, and
bioassay techniques. J. Fish. Res. Bd. Canada 29: 1309.
Colby, P.J. and L.L. Smith, Jr. 1967. Survival of walleye eggs and fry on
paper fiber sludge deposits in the Rainy River, Minnesota. Trans.
Amer. Fish. Soc. 96: 278.
Mount, D.I. and W.A. Brungs. 1967. A simplified dosing apparatus for
toxicological studies. Water Res. 1: 12.
Scidmore, W.J. 1956. An investigation of carbon dioxide ammonia and
hydrogen sulfide as factors contributing to fish kills in ice covered
lakes. Minn. Dept. Conser. Bur. Res. Plann., Invest. Rep. 177: 9.
Smith, L.L., Jr. and D.M. Oseid. 1974. Effect of hydrogen sulfide on
development and survival of eight freshwater fish species. In: The
Early. Life History of Fish. Blaxter, J.H.S. (ed.). Springer-Verlag,
New York. p. 416-430.
121
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SECTION 12
THE BEHAVIORAL ASPECTS OF AQUATIC TOXICOLOGY
B.A. Flerov
As a result of conspicuous progress in the study of behavior of
aquatic animals (Anon., 1972, 1975; Flerov, 1965), this science is be-
coming widely applied in aquatic toxicology.
As early as the 1950' s, a number of manuscripts appeared demonstra-
ting the peculiarities of action of pharmacological drugs by studies of
the behavioral reaction of fish. These efforts (Abramson and Evans,
1954; Abramson at al., 1958; and Evans
-------
on reproductive function, but a direct consequence of suppression of
sexual behavior in the animals.
The changes in behavior of aquatic animals are not only clear indica-
tions of intoxication, they are also the first symptoms of disturbances
of living activity. Just as an experienced physician can diagnose ill-
ness in his patient using only the patient's behavioral symptoms, a
rather attentive toxicologist can speak of intoxication of an organism by
judging its behavior. Therefore, behavior may be used as a sensitive test
for toxicity of the water medium. In this regard, conditioned reflexive
activity and learning of animals are of special interest.
Investigation of the conditioned reflexive reactions of fish under
the influence of toxic substances have not achieved prominance until quite
recently, in spite of the great experience gained by the Academician I.P.
Pavlov and his followers on the pathophysiology of higher nervous activity
under various generic pathogenic effects (Dolin, 1962; Frolov, 1944;
Ivanov - Smolenski, 1952).
Investigations of the effect of phenol in sublethal concentrations
upon locomotory defensive and locomotory feeding conditioned reflexes in
gold carp (Flerov, 1965) have shown that the disturbances of the higher
nervous activity of fish are of a general, nonspecific character, similar
to pathological alterations in functioning of the cerebral cortex in
mammals. These symptoms are manifested in inhibition of differentiation,
decrease in percentage of demonstration of positive reflexes, prolonga-
tion of the latent periods of positive reflexes, and, finally, in com-
plete suppression of conditioned reflex activity (Figure 1). The charac-
ter and degree of disturbances of the reflexes depend upon the typologi-
cal peculiarities of the higher nervous activity of a fish. In a fish of
a weak type, pathological alterations begin earlier, and are displayed to
the greatest degree (Flerov, 1973). Comparison of the sensitivity of the
conditioned reflex method with other physiological methods has shown that
it is order of magnitude more sensitive.
Alterations in the higher nervous activity may be successfully used
as a quality sensitivity test for determination of the quality of water.
The most varied methods of investigation of the conditioned reflex
activity may be applied for this purpose. In recent time this methodology
is being used to estimate the effects of low concentrations of mineral oil
(Kasymov and Rustamova, 1969), heavy metal ions (Krasnov, 1971; Wier and
Mine, 1970), and pesticides (Anderson and Peterson, 1969; Anderson and
Prins, 1970; Hatfield and Johansen, 1972; McNicholl and MacKay, 1975).
Water toxicologists have long been using disturbances of equilibrium
reflex as an index of intoxication in fish. Little attention, however,
has been paid to careful observations of all the symptoms of intoxication,
and to their objective recording.
Two examples illustrate this observation. The first example considers
the symptoms of acute intoxication in fish as a result of exposure to
123
-------
0 > 100
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- H
n
^*k/ y J«y W
~" PHENOL I i
— 4DD£Df J fm% At
I jk jm^ . ^% j^ ^^ \*^ *•* ^* «^
••^ I *^* * ^ ^* * *^^^W ^^ ft rt/ tF~&.tf\ 1
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— „ ^
y v». -r' **v\ x %v*"r'vVv*v-''s
— * » / v T
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/
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^**f**-f^^-r^*''~ — -^r v '\^ * " '^ "^V^^V^tj rv
1 1 1 1 1 1 1 1 1 1 I 1
v^ UNTREA TED CONTROLS
*""*V. «__«• Df\OI It A Tlf\Hi T"DC A TCn tA/ITLJ & Mf^ //
•~\\- ^\ "POPULATION TREATED WITH 6 MG/L
_MfV\ f
hMc£5£w^tefe
i i i i i i i i i i i i
r/V""\
"D
Lx
1 1
"*^^
^*»*,^*^
1 1
^*VT% •-
v V
I 1
DLJCHt/^ 1
PHENOL
1 1
20
40
60 80
TIME, days
100
120
140
Figure 1. Changes in the conditioned reflex activity in the
common guppy (LtbUtut, fLe£Lc.ula&i!>) under the influence of
sublethal concentrations of phenol.
124
-------
several different classes of toxic compounds (Table 1). In spite of the
fact that there are common features in the manifestation of pathology
(e.g., violent general locomotory activity, disturbance of equilibrium re-
flex on intoxication with phenol and polychlorpinene), characteristic
specific features are also revealed. Thus, under the influence of poly-
chlorpinene, in contrast to phenol, such specific symptoms as moving to
the surface and swallowing of the air are observed in fish during the
phase of the violent locomotory activity.
On intoxication with chlorophos, a prolonged inhibition stage and
darkening of coloration due to the opening of chromatophores are charac-
teristic. An abundant excretion of mucilage is observed as a result of
exposure to detergents.
A second example is to be found in the symptoms of intoxication in one
of the representatives of invertebrates, the medicine leech (HiAudo m&di-
CA.naLLf>: Annelida). Exposure of this organism to solutions of the toxic
substances noted above yields more specific reactions (Figure 2).
Intoxication with polychloropinene is first noted when the organism
rolls the anterior body segments ventrally (Figure 2, 1-4). A few hours
later, convulsions develop followed by immobility and death.
TABLE 1. SYMPTOMS OF INTOXICATION IN CARP EXPOSED TO THE SHORT-TERM
ACTION OF TOXIC SUBSTANCES
Toxicants Symptoms of Intoxication
Phenol Violent locomotory activity (fish perform im-
petuous rushes, frequently breaking their
snout against the walls of aquarium). Dis-
turbance of the equilibrium reflex (swimming
on the side); convulsions; immobilization;
death.
Polychlorpinene Prolonged increase in general exitability to
acoustic and tactile stimuli, violent swimming
activity. Fish swims to the surface and swal-
lows the air. Disturbance of equilibrium re-
flex; immobilization; fish floats up and dies.
Chlorophos Increase in general exitability to acoustic
and tactile stimuli. Sudden general inhibi-
tion (weak reactions to external stimuli, low
general activity); disturbance of coordina-
tion; tremor of body muscles; significant in-
tensification of coloration; sinking of the
fish to the bottom; immobilization; death.
"Lotos-71" Abundant excretion of mucilage; immobiliza-
tion; death.
125
-------
Figure 2. Symptoms of intoxication of medicine leech in solutions
of polychlorpinene (1-4), chlorophos (5-8), and phenol (9-12).
See text for explanations.
126
-------
In solutions of chlorophos, initial symptomology is manifested by
coiling of posterior segments, and twisting them into a spiral. Grad-
ually, twisting of the whole body into a tight spiral occurs, so that the
movement ceases (Figure 2, 5-7). The leech subsequently straightens, the
gullet opens and swallowing of air occurs. Both volume and weight of the
organism increases (Figure 2, 8).
In a solution of phenol, initial disorderly locomotor activity and co-
ordinative disturbances are followed by a looped attachment to the walls
of the vessel (Figure 2, 9). The leech then drops to the bottom, convul-
sions develop and characteristic constrictions appear in the body (Fi-
gure 2, 10-12). Such observations enable an estimation of the character
and specificity of action of a harmful compound, suggest a course of study
of actual functions responsible for the development of the pathological
process, and enable classification of the intoxications.
The most important problem presently facing aquatic toxicology is the
question of adaptation of the organisms to a new environmental factor,
the toxicants. Because of their active relationship with the environ-
ment, aquatic animals can avoid the harmful effects of such factors,
though defensive behavior incorporating the avoidance reaction.
The ability of fish to avoid toxic solutions under laboratory condi-
tions has been widely considered (European Inland Fisheries Commission,
1965; Hansen etal., 1972, 1974; Ishio, 1965, 1969; Jones, 1951, 1957,
1964; Shelford, 1971; Sprague and Drury, 1969). The experiments have
been performed using many species of fish (carp, crucian carp, minnow,
loach, trout, etc.) and various toxic compounds: cyanids, phenols, salts
of zinc and copper, carbonic acid, ammonium, chlorine, hydrogen sulphide,
pesticides, detergents, and industrial wastes. In general, the results
of the experiments have shown that various species of fish avoid the
zones with sublethal concentrations of toxic substances.
Ishio suggests that the avoidance reaction obeys the law of Weber-
Phekhner, i.e., the response is proportional to the logarithm of the irri-
tant intensity (concentration of the toxic substance). However, this
reaction varies widely, depending on the species of fish, and the chemi-
cal properties of the toxicants. Its manifestion may be strong, weak, or
entirely absent. For example, phenol is actively avoided by the carp,
but the contaminant is not avoided, even in lethal concentrations, by
salmon ids. Moreover, some toxic substances are even preferred by fish.
Thus, low alkalinity, ammonium hydroxide, and low concentrations of
copper salts possess attracting properties. The reaction of trout to
dissolved chlorine is most interesting (Sprague and Drury, 1969). A dis-
tinct avoidance reaction is observed in the range of low (0.001-0.01 mg/1)
and high (1 mg/1) concentrations, but in the range of medium concentra-
tions, a clear preference is revealed. It is difficult to explain the
difference in reactions. Studies of associated physiological mechanisms
are still lacking. However, perception and discerning of toxic substances
is largely provided by organs of smell and taste, and to the so-called or-
gan of general chemical sense. For example, fish can discern phenol and
127
-------
parachlorphenol in concentrations as low as 0.0005 mg/1 (Hasler and Wisby,
I ,/Dw ) •
There are still few data on the avoidance reaction to toxic substances
in aquatic invertebrates. Studies have utilized freshwater insects,
spiders, leeches, and marine crustaceans. The forms of behavior pre-
venting toxic exposure to animals are varied. Beetles usually crawl out
of toxic solutions along the walls of the experimental vessels and occa-
sionally even fly out of the solutions. Hymenoptera jump out and fly
away, and water spiders escape by running along bits of grass placed into
the vessel. Animals incapable of rapid movement manifest defensive be-
havior in other ways. The larvae of many flies and some butterflies use
their cases as means of protection from toxicants. Some caddis fly larvae
in toxic media will build thicker case walls. Chironomids bury themselves
deeper into the mud. Molluscs close the valves of their shells tighter
and for a longer period (Alekseev and Flerov, 1972).
The most dangerous toxic substances (pesticides) are, however, poorly
avoided by invertebrates. Again, the medicinal leech will serve to illu-
strate this fact. A comparison of avoidance of substances belonging to
different classes of chemical toxicants (Table 2) shows that pesticides
are either not avoided at all (chlorophos), or avoided only at their
lethal concentrations (polychlorpinene). These substances are evidently
not "unpleasant", causing pain for the leech. Herein lies the insidious-
ness of pesticides. In contrast, phenol and "Lotos-71" were actively
avoided by the animals.
The avoidance reactions are of great importance for adaptation. For
practical purposes, it is essential to know the range of concentrations
in which these reactions are manifested. However, questions arise rela-
tive to the extrapolation of experimental results into the natural system.
There are some observations on migrations of marine fish from the
areas polluted with petroleum wastes. Data also exists on impoverishment
of species composition of communities which may be explained, not only by
the death of some species, but also by escape of others from the polluted
TABLE 2. TOXICITY OF SOME SUBSTANCES FOR MEDICINE LEECH AND THEIR
THRESHOLD CONCENTRATIONS (M6/L) PRODUCING AVOIDANCE REACTION
Exposure 48 hrs.
Maximum Tolerated
Toxicants Concentrations LC5Q
"Lotos-71"
Phenol
Polychlorpinene
Chlorophos
150
275
2.5
0.05
190
290
5
0.3
LCiQO
300
100
10
0.6
Threshold
Concentration
1
50
5
No avoidance
128
-------
areas. The initial comparison of avoidance reactions in the laboratory
and in the field yields barely consoling results. Mature salmon actively
avoiding salts of copper and zinc under experimental conditions displayed
virtually unnoticeable reactions in nature (Sprague, 1971). Studies of
this kind are very important, and their further development is required.
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II: Psychobiological effects on the Siamese Fighting Fish. Science,
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Abramson, H.A., B. Weiss, and M.O. Baron. 1958. Comparison of effect of
lysergic acid diethylamide with potassium cyanide and other respira-
tory inhibitors on the Fighting Fish. Nature, V. 1, N. 4616.
Alekseev, V.A., and B.A. Flerov. 1972. Effect of phenol on photoreac-
tions and resistance of CfuAonomtt6 pliwoi>uu> and Larmoaktviu a.qu.cutic.a_.
Inform. Bull. Inst. Biol. vnutr. vod AN SSSR, N. 13.
Alekseev, V.A., and B.A. Flerov. 1972. Reaction of avoidance of toxic
solutions of phenol by some aquatic insects and arachnids. Inform.
Bull. Inst. Biol. vnutr. vod AN SSSR, N. 14.
Alekseev, V.A., and B.A. Flerov. 1972. Some peculiarities of behavior
of aquatic invertebrates (In4£c-£a, M&c/mo-tda) under conditions of
phenol intoxication. In Coll.: Behavior of Aquatic Invertebrates,
Brook.
Anderson, J.M., and M.R. Peterson. 1969. DDT: Sublethal effects on
brook trout nervous system. Science, V. 164, N. 3878.
Anderson, J.M., and H.B. Prins. 1970. Effect of sublethal DDT on simple
reflex in brook trout. J. Fish. Res. Bd. Canada, V. 27.
Anon. 1972. Behavior of aquatic invertebrates.
Anon. 1975. Behavior of aquatic invertebrates.
Anon. 1975. Invertebrate learning. New York, V. 3.
Bengtsson, B.E. 1974. The effect of zinc on the ability of the minnow,
PdoxxtmM ph.oxA.nuA L., to compensate for torque in a rotating water
current. Bull- Environ. Contan. Toxicol., V. 12, N. 6.
Besch, W.K., I. Juhnke, and A. Kemball. 1972. Standartisierung des
Fischwarntestes. Schr. Reine Ver. Mass. Boden Lufthyg, Stuttgart,
N. 37.
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Braginski, L.P., I.L. Burthaya, and E.P. Shcherban. 1972. Some peculiar-
ities of behavior of aquatic invertebrates in toxic medium and thier
importance for entomology. In Coll.: Behavior of Aquatic Inverte-
brates, Borok.
Davy, F.B., H. Kleerekoper, and P. Gensler. 1972. Effects of exposure to
sublethal DDT on the locomotor behavior of the goldfish
. J. Fish. Res. Bd. Canada, V. 29, N. 9.
Dolin, A.O. 1962. Pathology of the higher nervous activity.
European Inland Fisheries Advisory Commission. 1965. Water quality cri-
teria for European freshwater fish. Report on Finely Divided Solids
and Inland Fisheries, Rome.
Evans, L.T., L.H. Geronimus, C. Kotnetsky, and H.A. Abramson. 1956.
Effect of ergot drugs on B&tta. .*p£enctett6 . Science, V. 123.
Flerov, B.A. 1965. Influence of phenol on conditioned reflex activity in
fishes. Gidrobiol. zhurn., V. 1, N. 3.
Flerov, B.A. 1965. Influence of low concentrations of phenol on locomo-
tory, feeding activity and growth in crucian carps. Voprosy ikhtiol.,
V. 5, N. 1.
Flerov, B.A. 1969. Influence of subtoxic concentrations of phenol on
sexual behavior of Leb-c4te6 A.e££ca£a£u4. In Coll.: Physiology of
Aquatic Organisms and Their Role in Circulation of Organic Matter.
L.
Flerov, B.A. 1973. Experimental investigations of phenol intoxication
in fishes. In Coll.: Influence of Phenol Upon Hydrobionts. L.
Flerov, B.A., and L.N. Lapkina. 1976. Avoidance of solutions of some
toxic substances by the medicine leech. Inform. Bull. Inst. Biol.
vnutr. vod AN SSSR, N. 30.
Foster, N.R., A. Scheier, and J. Cairns. 1966. Effects of ABS on feeding
behavior of flagfish 3onda.n<>JULoi ^o'u.dae.. Trans. Am. Fish. Soc.,
V. 95.
Frolov, Yu.P. 1944. Higher nervous activity in toxicoses. M.
Hansen, D.J., E. Matthews, S.L. Nail, and D.P. Dumas. Avoidance of pesti-
cide by untrained mosquitofish, Gambu^^La. a^/tw^A. Bull. Environ.
Contam. Toxicol., V. 8, N. 1.
Hansen, D.J., S.C. Schimmel, and J.M. Kelther. 1973. Avoidance of pesti-
cide by grass shrimp ( Po£aemone£e^ pug-io) . Bull. Environ. Contam.
Toxicol., V. 9, N. 3.
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Hansen, D.J., S.C. Schimmel, and E. Matthews. 1974. Avoidance of Aroclor®
1254 by shrimp and fish. Bull. Environ. Contam. Toxicol., V. 12,
N. 2.
Hasler, A.D., and W. Wisby. 1950. Use of fish for the olfactory assay
of pollutants (phenols) in water. Trans. Am. Fish. Soc., V. 79.
Hatfield, C.T., and P.H. Johansen. 1972. Effects of four insecticides on
the ability of Atlantic Salmon (S&tmo kaJLoJi) to learn and retain a
simple conditioned response. J. Fish. Bd. Canada, V. 29, N. 3.
Ivanova, V.I. 1961. On the mechanism of action of aminazine on feeding
locomotory conditioned reflexes in fishes, pigeons and rabbits.
Zhurn. vyssh. nervn. deyat., V. 7, N. 11.
Ivanov-Smolenski, A.G. 1952. Sketches on pathophysiology of the higher
nervous activity (by data of Pavlov and his school). M.
Ishio, S. 1965. Behavior of fish exposed to toxic substances. Adv.
Water Pollut. Res. V. 1.
Ishio, S. 1969. Discussion. Adv. Water Pollut. Res., Pergamon Press.
Jones, J. 1951. The reactions of the minnow, PkoxJ.nu.A pfooxxtnoa (L.) to
solutions of phenol, ortho-cresol and para-cresol. J. Expl. Biol.,
V. 28.
Jones, J. 1957. Fish and river pollution. Aspects of river pollution.
London.
Jones, J. 1964. Fish and river pollution. London.
Kasymov, R.Yu., and Sh.A. Rustamova. 1969. Influence of different con-
centration of petroleum upon the dynamics of conditioned reflexes in
the young of the sturgeon. Mater, nauchn. sessii TsNIORH, Astrakhan.
Krasnov, S.K. 1971. Methods of feeding locomotory conditioned reflexes
in fishes. In Coll.: Metodiki biol. issled. po vodn. toksikolog.
M.
McNicholl, P.6. and W.C. Mckay. 1975. Effect of DDT on discriminating
ability of rainbow trout (Salmo gcuAdne.>u.). J. Fish. Res. Bd. Canada,
V. 32, N. 6.
Ogilvie, D.M., and J.M. Anderson. 1965. Effect of DDT on temperature
selection by young Atlantic Salmon, Salmo t>oJt
ptLom&loA (Rafinesque) and its changes induced by copper salt CuSO,.
Pol. Arch. Hydrobiol., V. 18, N. 4.
131
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Peterson, R.H. 1973. Temperature selection of Atlantic Salmon (Salmo
&aJt(Vi) and brook trout (So£ve£mo6 &on£inatl!>) as influenced by
various chlorinated hydrocarbons. J. Fish. Res. Bd. Canada, V. 30,
N. 8.
Shelford, V.E. 1917. An experimental study of effects of gas waste upon
fishes, with special reference to stream pollution. Bull. 111. State
Hab. Nat. Hist., V. 1.
Sprague, J.B. 1971. Measurement of pollutant toxicity to fish. III.
Sublethal effects and "safe" concentrations. Water Res., V. 5.
Sprague, J.B., and D.E. Drury. 1969. Avoidance reactions of salmonid
fish to representative pollutants. Adv. Water Pollut. Res., V. 1.
Swedmark, M., B. Braaten, E. Emannuelsson, and A. Granmo. Biological
effects of surface active agents on marine animals. Mar. Biol.,
V. 9, N. 3.
Weir, P.A., and C.H. Mine. 1970. Effects of various metals on behavior
of conditioned goldfish. Arch. Environ. Healt., V. 20, N. 1
William, T., Waller and 0. Cams, Jr. 1972. The use of fish movement
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132
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SECTION 13
GEOLOGIC POLLUTION PROBLEMS OF LAKE SUPERIOR
Albert B. Dickas, Ph.D.
ASBESTIFORM MINERALS
Background
On 17 February, 1972, the U.S. Environmental Protection Agency first
stated that the presence of the trace mineral cummingtonite in the Wis-
consin portion of Lake Superior constituted interstate pollution. This
statement was made part of preliminary evidence being gathered in the now
famous "Reserve Mining Case" (Reserve Mining Company versus United States
of America, No. 74-1291). In December of that same year the first public
statement was made of concern over the relationship of the presence of as-
bestos and potential public health hazards of residents of the Lake
Superior Basin (Great Lakes Research Advisory Board, International Joint
Commission, 1975). Since then a number of detailed investigations of the
distribution and health effects of asbestos have been undertaken.
Prior to this time, the region was internationally known as an econo-
mic source of silver (historically), copper, as well as high grade (hema-
tite) and low grade (taconite) iron ore, but not asbestos in any commer-
cial quantity. Thus little research has been done on the distribution of
this mineral within the Great Lakes area. Although the effects of inhaled
asbestos are reasonably well known to be the causative agent of the
disease asbestosis (a scaring of the lungs by increased fibrous tissue
growth), the effects of ingested asbestos have only recently been con-
sidered by the scientific and medical community.
Mineralogy, Chemistry and Morphology
The term asbestos is defined by Dana (1954) as constituting the fib-
rous varieties of serpentine and amphibole, the fibers of which are some-
times very long, fine, flexible, easily separable by the finger and look
like flax. The term is derived from the Greek for "incombustible".
Generically, the term is used to describe fibrous hydrated silicates, con-
sisting of 40-60 percent silica in combination with oxides of iron, magne-
sium and other metals. The minerals differ in their chemical and physical
properties, such as fiber diameter, flexibility, tensile strength and sur-
face properties.
133
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A conventional description would include highly perfect cleavage, sub-
conchoidal to uneven fracture, vitreous to pearly luster, black, white, green
brown and pink in color and with an uncolored streak. Pleochroism is strong
with deeply colored varieties with absorption usually Z>Y>X (Dana, 1954).
Mineralogically there are two main groups of asbestos minerals: ser-
pentine and amphibole. Common types, crystal class, chemistry and identi-
fication characteristics are given in Table 1.
Water is considered an essential constituent of all types. Common fiber
lengths for all of the asbestos minerals are within the range of 0.2-10 ym
(ym = 10-6 meter). These values depend somewhat on the laboratory method
employed to disperse the fibers for measurement.
For bulk analysis, the techniques of infrared spectroscopy, differential
thermal analysis and X-ray diffraction are employed. In order to study in-
dividual fibers, transmission electron microscopy, selected area electron
differaction and electron microprobe analysis methodology are employed. While
acceptance precision can be achieved through these expensive and time-con-
suming programs, the superiority of any one method has yet to be demonstrated
(Great Lakes Research Advisory Board, International Joint Commission, 1975).
Geological Sources
Chrysotile asbestos occurs in serpentine form that has been altered from
(a) ultrabasic rocks such as peridotite or dunite or (b) magnesium lime-
stones or dolomites (Bateman, 1950). The former occurrence yields about 90
percent of the world's asbestos supply. Amphibole varieties are found in
slates, schists, banded ironstones and as lenses and pockets in peridotite and
pyroxenite.
TABLE 1. CLASSIFICATION OF ASBESTOS. DATA FROM GREAT LAKES RESEARCH
ADVISORY BOARD, INTERNATIONAL JOINT COMMISSION, 1975
Group
Serpentine
Amphibole
Amphibole
Amphibole
Amphibole
Name
Chrysotile
Anthophyllite
Cummingtonite*
Grunerite*
Tremolite**
*That particular amphibol
Mining Company tailings
**That particular amphibol
Superior.
Chemistry
t(Mg,Fe)2Si04]
[(Mg,Fe)Si(h]
[(Mg.Fe)SIO,]
[(Fe,Mg)SiO?]
[(Ca.MgJSi^i
Crystal
Monoclinic
Orthorhombic
Monoclinic
Monoclinic
Monoclinic
Characteristics
Hollow Curved Fibers:
00^250 A; IDR50 A; of-
ten occur in bundles
and change shape in
lung liquids.
Straight:
do not oc-
cur in bun-
dles and do
not split or
change shape
in lung liq-
uids
OD>2500 A
OD>1500 A
OD>1500 J"
OD> 600 A
e generally released into Lake Superior by Reserve
discharge of 67,000 tons/day.
e found in natural lacustrine sediments of Lake
134
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Generally speaking asbestos might be found wherever basic and ultra-
basic rocks have been serpentinized (the conversion of ferromagnesium
minerals or rocks to aggregates of serpentine minerals) by autometamorphism
(metamorphism of igneous rock by its own volatile fluids) (Bayly, 1968),
thus forming chrysotile; or by load metamorphism (deep burial accompanied
by mineralizing vapors), thus forming amphibole asbestos. This brief des-
cription of asbestos sources and formation fits quite nicely the geological
conditions found within the Lake Superior Basin.
Source
The history of taconite mining in the Lake Superior Basin is the his-
tory of low grade iron ore beneficiation which resulted from the effective
exhaustion of primary high grade hematite ores. Taconite is well cemented,
ferruginous chert and slate. In order to be useful, such rocks must be
beneficiated from material containing 25-30 percent iron to material con-
taining as much as 65 percent iron. This high grading is accomplished by
crushing and grinding the taconite to the point where in excess of 90 per-
cent of the material is finer than flour. The purpose is to isolate the
iron "ore" from the silica "gangue", so that the ore can be separated by a
magnetic process.
A large quantity of water is employed in this process, both for con-
tinued washing and sizing of the material, and as a medium for handling.
In the beneficiation process, the iron "flour" is made into "green" pel-
lets, approximately 1.27 centimeters (0.5 inches) in diameter by rolling a
mixture of bentonite clay and magnetic grains in a large revolving drum.
The "green" pellets are then baked to 1316 °C (2400 °F) in a kiln where
they are converted to taconite pellets, a very desirable blast furnance
feed. In the process, up to 37,850 liters (10,000 gallons) of water are
used for each ton of pellets produced. In addition, the ratio of waste
tailings to concentrated pellet production is approximately 2/1, that is,
two million tons of tailings to 1 million tons of pellets (Great Lakes
Research Advisory Board, International Joint Commission, 1975).
Pollution problems associated with this beneficiation process relate
to the dumping of tailings into Lake Superior and the voluminous use of
water, the state in which it is left, and where the waste is disposed.
All are concerns of the health and environmental community.
Treatment
Treatment of asbestos containing waters falls into two principle
methods:
1. Ordinary filtration by sand or diatomaceous earth, which
has proven to be approximately 90 percent effective.
2. Chemical coagulation with iron salts and polyelectrolytes
followed by filtration, which is more than 99 percent
effective.
135
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Possible Health Effects
Although the effects of inhaled asbestos are reasonably well docu-
mented, the effects of ingested asbestos have only recently come under
study. The Great Lakes Research Advisory Board indicates the following oc-
cupational hazards:
1. Asbestosis; increased fibrous tissue growth of the lung.
Disease will appear after 10-40 years of occupational
exposure.
2. Pleura! calcification; a deposition of insoluble salts
in the lung lining. Occurs after an approximate 20 year
latency period.
3. Mesothelioma of the chest and abdominal cavity lining;
with a latency period of 20-40 years, this disease was
considered, until recently, quite rare.
Much less is known regarding ingested asbestos. It has been shown
that the high rate of stomach cancer in Japanese is linked to their use of
rice dusted with asbestositic talc (Merliss, 1971). Laboratory tests em-
ploying rats show that asbestos will accumulate in the brain and in tissue
surrounding the small intestine. It may also cause malignant tumors in
the kidney, lymph-nodes and brain (Pontefract and Cunningham, 1973). How-
ever, present knowledge of public health aspects of asbestos in drinking
water supplies is inadequate. In consideration of the potential of this
problem, the significance of a possible 20 year delay following even short-
term exposure must be given proper perspective (U.S. Circuit Court of
Appeals, 1974).
Considering direct biochemical effects to Lake Superior, taconite tail-
ings deposition has shown to bring about the following results (Federal
Water Pollution Control Administration, 1970):
1. A reduction in the abundance of fish food production suf-
ficient to create five (5) percent reduction in commer-
cial and sport fishing.
2. Chemical analysis projections, based upon daily discharge
of 67,000 tons of tailings, indicates daily discharge of
copper, nickel, zinc, lead, chromium, phosphorus and man-
ganese ranging from 1,860 kilograms to 285,310 kilograms
(4,100 to 629,000 pounds).
RED CLAY TURBIDITY
Based on the discharge analysis of major tributary systems, open lake
turbidity and shoreline recession rates, it is estimated that the total
gross erosion into the U.S. portion of Lake Superior is in excess of 4.8
million tons/yr. (a literature review of this subject will yield a wide
136
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estimate due to obvious field sampling difficulties and variability of
discharge rates and storms with time of season). Along the north shore of
the lake the estimated average annual yield is 1.1 ton/km2, a relatively
low figure due to geology (basically igneous and Precambrian strata), soil
types, vegetation and land use. Along the south shore, where Pleistocene
lake development sequences left a thick, exposed layer of easily erodible
red clay, estimates range in excess of 6.8 ton/km2 (Great Lakes Basin
Commission, 1975). Thus, the subject of turbidity is basically one of red
clay erosion along the Lake Superior south shore.
Of particular concern from the viewpoint of sediment erosion control
has been the question of ultimate source of such lake turbidity: tribu-
tary inflow, lake shore erosion by high water levels, storm activity, and
resuspension of previously deposited clay. The latter source is of signi-
ficant debate considering the circulation patterns of Lake Superior.
While the inflow and outflow rate of this lake is small in comparison to
the water mass, the lake water is not standing still. It is kept in con-
stant motion principally by the wind, which not only generates the visible
surface waves (in excess of 4.9 meters), but stirs and mixes the water
throughout the lake. Both water movements and rates of mixing are in-
fluenced by the formation of temperature (and associated density) thermo-
clines. In the summer, Lake Superior becomes divided into an upper layer
of warm readily circulating water, the epilimnion, and the lower layer of
cold, relatively undisturbed water, the hypolimnion. The contact between
these two zones where rapid temperature changes takes place is termed the
thermocline. When the lake is stratified, the hypolimnion is essentially
physically and chemically isolated from the remaining water. In Lake
Superior, nearly 95 percent of the lake's volume is in the hypolimnion
(Federal Water Pollution Control Administration, 1969). The summer strati-
fication begins to develop in mid-July, with the epilimnion reaching its
maximum temperature of approximately 21 °C in August. In the winter
months, the lake can be considered, for all practical purposes, to be iso-
thermal .
Because currents in the lake are motivated principally by wind, and
the winds are variable, the horizontal movements of lake waters exhibit
infinite variety, and frequent changes in both direction and speed. The
net circulation is counter clockwise, with the possibility of large cy-
clonic eddies occurring in the western arm (Great Lakes Basin Commission,
1975). Upwelling occurs in the lake when winds cause horizontal surface
movement of water away from the shore and the surface waters are replaced
by colder, deeper water (Upper Lakes Reference Group, International Joint
Commission, 1977).
Considering such currents, storms, and precipitation cycles, it has
recently been estimated that lake turbidity is primarily due to shoreline
erosion of lacustrine clay by storm currents. The most recent available
data (Sydor, 19/5) indicates rates for the three causes to be:
1. Tributary erosion into Lake
Superior: >bUO,OUO metric tons per year
137
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2. Shoreline erosion into Lake
Superior: > 4,000,000 metric tons per year
3. Resuspension of lacustrine >300,000 metric tons per_year for
clay in Lake Superior: depths of <" 21 meters (7 70 feet).
Chemical Effects
The water quality of western Lake Superior is directly affected by ex-
tensive erosion of the glacial-lacustrine red clay deposits (Dikas et at .,
1973). In addition to loss of property value, increased turbidity in
drinking water, and a decrease in aesthetic value, Bahnick (1976) has
attributed the following aquatic chemical changes to red clay turbidity.
Parameter Turbid Conditions Deep Water Conditions
Suspended Solids
Alkalinity (ppm CaCOJ
Hardness (ppm CaCO )d
Calcium ^
Magnesium
Sodium
Iron
Orthophosphate
Chemical Oxygen Demand
>110
> 42
> 50
> 15
> 3
> 1.5
> 19
> 0.04
(Mg 0/1 )> 16
ppm
ppm
ppm
ppm
ppm
ppm
ppm
ppm
ppm
^ 0.5 ppm
^ 40 ppm
'\, 46 ppm
•\- 1 4 ppm
•\> 2.7 ppm
^ 1.3 ppm
•v 4.0 ppm
•v 0.02 ppm
••o 3 ppm
Biotic Effects
In addition to human health effects, the problems of biotic effects of
red clay turbidity centers on the overall effect upon the fisheries
resource of Lake Superior. Recent studies have approached this problem
from the viewpoint of turbidity effects upon species composition, feeding,
predation, distribution, mortality and growth. Swenson (1975) has found
the following preliminary results:
1. Changes in food habits of major species can be expected to
result indirectly from turbidity through its influence on
light penetration, fish distribution and distribution of
plankton suites. Increased predation by pelagic smelt on
larval herring as a result of turbidity may have resulted
in the decline of commercial lake herring population as
smelt can be expected to leave the bottom during turbid
periods and increase predation pressure on the herring
larval.
2. Turbidity may have indirect effects on walleye feeding
success, rates and time.
3. Tank experiments indicate lake trout have a preference for
low turbidity, while walleye showed a demonstrated pre-
ference for the highest turbidity levels.
138
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In addition, red clay deposition would be expected to clog potential
spawning grounds in the shelf areas of the lake. Finally, on a more quali>
tative basis, some Lake Superior commercial fishermen claim that if the
clay material is readily visible in suspension during the time of freeze-
up, then the winter fishing catch is greatly reduced.
This subject is a very complex one and continues to be studied in the
Lake Superior Basin.
REFERENCES
Bahnick, Donald A. Department of Chemistry, University of Wisconsin-
Superior, personal communication.
Bateman, Alan M. 1950. Economic mineral deposits. 2nd Edition, John
Wiley and Sons, Inc. 290-294 p.
Bayly, Brain. 1968. Introduction to petrology. Prentice Hall, Inc.
371 p.
Dana, Edward S. 1954. A textbook of minerology. 4th Edition, 16th print-
ing by William E. Ford, John Wiley and Sons, Inc. 574 p.
Dickas, A.B., J.W. Horton, R.D. Morden, P.M. Ruez and W.A. Swenson. 1973.
Environmental effects of harbor dredging Superior-Duluth Harbor. Cen-
ter for Lake Superior Environmental Studies (University of Wisconsin-
Superior), contract publication No. 6.
Federal Water Pollution Control Administration, Great Lakes Basin. 1969.
An appraisal of water pollution in the Lake Superior Basin.
Federal Water Pollution Control Administration, Great Lakes Region. 1970.
An appraisal of water pollution in the Lake Superior Basin.
Great Lakes Basin Framework Study. 1975. Water quality (Appendix 7).
Great Lakes Basin Commission under sponsorship of the U.S. Environmen-
tal Protection Agency.
Great Lakes Basin Framework Study. 1975. Erosion and sedimentation (Ap-
pendix 18). Great Lakes Basin Commission under sponsorship of the
U.S. Environmental Protection Agency.
Great Lakes Research Advisory Board. 1975. Asbestos in the Great Lakes
Basin with emphasis on Lake Superior. 35, A-6 p.
Merliss, R.R. 1971. Talc-treated rice and Japanese stomach cancer.
Science, V. 173, 1141-42.
Pontefract, R.D. and H.M. Cunningham. 1973. Penetrations of asbestoc
throughout the digestive tract of rats. Nature, vol. 246, 352-53.
139
-------
Swenson, W.A. 1975. Influence of turbidity on fish abundance in western
Lake Superior. Progress Report for U.S. Environmental Protection
Agency, Project no. R-802455-02.
Sydor, M. 1975. Turbidity in extreme western Lake Superior. Unpub-
lished report.
U.S. Court of Appeals, Eighth Circuit, Reserve Mining Company, oJt al.t vs,
United States of America, ^t aJL. No. 74-1291, June 4, 1974.
Upper Lakes Reference Group. 1977. The water of Lake Huron and Lake
Superior, Vol. Ill (Part B), Lake Superior. Report to the Inter-
national Joint Commission.
140
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SECTION 14
EXPERIMENTAL APPLICATION OF VARIOUS SYSTEMS OF BIOLOGICAL
INDICATION OF WATER POLLUTION
G.6. Winberg
In the Soviet Union, there is no generally accepted system of evalua-
tion of water pollution by hydrobiological indices. As in a number of the
European countries, the most widely utilized was the saprobian system of
Kolkwitz and Marson particularly modified by Zelinka and Marvan, Pantle
and Buck, and Sladecek. At present, work has been started on the relative
evaluation of various methods of biological analyses of water pollution.
It should be noted that hydrobiological analysis may be used for two es-
sentially different purposes: 1) to obtain a relative evaluation of the
water quality at as given time, and 2) to obtain data objectively charac-
terizing the condition of aquatic ecosystem, intended to be used subse-
quently for the study of long-term alterations.
Comparative evaluation of various systems of hydrobiological analysis
of polluted waters has been made by the staff of the Laboratory of Fresh-
water and Experimental Hydrobiology of the Zoological Institute of the
USSR Academy of Sciences. For this purpose during 1973-1975, hydrobiolo-
gical samples were collected at 26 stations representing various degrees
of pollution on several rivers of the Leningradskaya region, including
the rivers: Izhora, Luga, Vuoksa. In the Kaliningradskaya region the
tributaries of the system of the Pregol were included as was Moskva River.
On all the rivers, the samples were collected in July and August of 1973-
75 with the exception of the Izhora River, which was considered as a model
and where the samples were taken on eight occasions representing all the
seasons of the year. Detailed investigations were made of the phytoplank-
ton and periphyton (V.N. Nikulina), planktonic ciliates (T.V.
Khlebovitsch), zooplankton (M.B. Ivanova, L.A. Kutikova, A.V. Makrushin)
and zoobenthos (A.F. Alimov, N.P. Finogenova, E.V. Balushkina, S.Ya.
Tsekholikhin). The degree of pollution on each of the stations was
characterized by hydrochemical (N.G. Ozeretskovskaya, V.V. Bulion) and
bacteriological (M.F. Fursenko) data. Total counts of bacteria were made,
the number of heterotrophic bacteria (plate counts on MPA) and the
associated heterotrophic activity were determined by the method of Right
and Hobbie. Of the 6 classes of polluted waters, only the classes II-V
were represented, i.e., very clean (class I) and very dirty (class VI)
water were not found.
141
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Application of various modifications of the method of saprobic indica-
tor organisms (including the methods of Knepp, Pantle and Buck, Zelinka
and Marvan, Rotshain, Sladecek) was made to collections of phytoplankton,
planktonic ciMates, rotifers and crustaceans. The results lead to very
similar evaluations of the quality of the waters.
Detailed algological investigations on the 5 rivers showed that 50
percent of the species observed and nearly 80 percent of the dominating
species are cited in the lists of saprobic indicators by Sladecek. This
enabled a comparison of the methods of treating the data by Knepp,
Zelinka and Marvan, Rotshine and Sladecek. The greatest possibility of
differentiation of the stations with varying degrees of pollution are pro-
vided by the estimation of the mean saprobic valency according to Zelinka
and Marvan. This method, while useful, can not compensate for the ad-
vantages of simplier methods which generally lead to similar results.
This is especially true of the method of Sladecek, or to be more correct,
his modification of the method of Pantle and Buck.
All the 5 investigated rivers must be classified as 3 mesosaprobic
waters in spite of evident differences in their degree of pollution
according to the saprobic index. Within this class, however, the rivers
are arranged in a sequence by ranking the mean saprobic index correspond-
ing to relative pollution.
Of 120 species of planktonic crustaceans and rotifers, 97 are included
in the list of Sladecek. Estimation of the saprobic index lead to the
same general conclusions as were drawn by data available on the phycology.
The methods based on the application of the lists of saprobic indicators
of phytoplankton, planktonic crustaceans and rotifers generally reflected
the correct varying degrees of pollution of the investigated rivers, but
they displayed only poorly the differences between stations on the same
stream, especially when the influence pollution levels were low. This
fact naturally established boundaries for the application of these
methods.
The results of the zoobenthic assays lead to other conclusions rela-
tive to the application of the system of saprobic indicator organisms.
The proposed systems of indicator organisms appeared to be not applicable
in many cases for conditions in the USSR. One of the reasons is the
difference between the fauna of the Middle European countries, and the
rivers of the USSR. For example, 132 species of macrobenthic forms are
presented in the table of indicator organisms. Of 170 species of benthic
animals from our collections, only 17 can be found in the table of indica-
tor organisms. The average number indicator species in a sample usually
did not exceed 28 percent. At some stations on the Moskva and Vuoksa
Rivers, the indicator organisms in the zoobenthos were absent entirely.
The difference between the fauna and the number of the indicator species
will be even greater for the areas of the Far East, Kamchatka, Sakhalin,
Middle Asia, the Caucases and the like.
142
-------
It is difficult to agree with those values for the saprobic valencies
and indicator weights which Zelinka and Marvan give for Tub^ex Tittup ex,
LtmnodsuJLuA ho{^meA^t^fu., and Votamothtvix. motda.vie.nb-i!>. The first two
species are indicated as characteric of a and p saprobic zones, and the
latter only for a saprobic zones. At the same time, it is well known
that all of them are typical also of 8 saprobic zone, and T. £ubx^e.x is
one of the primary species of oligochaets in oligotrophic lakes.
Other varieties of the system of Kolkwitz and Marson (Knepp, Pantle
and Buck) contain arbitrary evaluations of the number of organisms. These
do not seem to be sufficiently correct to be applied to benthic animals.
The use of the terms "many" "few" in these systems for quantifying
organism abundance will have a variable meaning and cannot be applied
with certainty.
Of current wide use for the evaluation of the degree of pollution is
a system based on the application of large numbers of benthic taxa. It
has been evident for some time that groups of aquatic insect larvae occur
in clean waters. Oligochaets, on the other hand, easily resist pollutions
and attain great abundance in the sediments enriched with organic matter.
Therefore, it is not surprising that indices which account for abundance
or biomass of oligochaets, or their separate species (Parele, 1975; Carr
and Hiltunen, 1965; Goodnight and Whitley, 1961; Zahner, 1964, and 1965),
or compare the ratio of the biomass of insects to that of oligochaets
(King and Ball, 1964) are the most widely occurring.
Evaluation of the degree of river pollution with these indices has
shown that some of them are probably correct only for those water-bodies
for which they were proposed (American Great Lakes, Boden Lake, the
Daugava and the Lielupe Rivers). The index of King and Ball does not
account for the seasonal dynamics associated with numbers of insect
larvae. Thus, one time collections may lead to incorrect values. The
index which deserves attention is the one suggested by Zahner which
considers the number of oligochaets, T. iub^ex and species of the genus
LimnodbvituA. In this system, seasonal dynamics of oligochaets are con-
sidered in quite an unusual way.
The method with the greatest perspective for biological analysis of
polluted waters using the composition of the bottom animals seems to be
the one proposed by Woodiwiss (1964) for the Trent River. The undis-
puted advantage of this method is that it unites the principles of indi-
cator values of separate taxa (distinctly fewer than the indicator
species list), and the principle of decreases in diversity of the fauna
under the conditions of pollution. It is important that Woodiwiss1
system of "grouping" is understood to be rather broad. For some animals
this implies separate species (larvae Plecoptera, Ephemeroptera), for
others it suggests large taxa (e.g., the famely Tubificides). At the
same time, this system reflects a simplification of trophic relationships
with respect to pollution, e.g., the decrease in the numbers or disappear-
ance of predatory animals.
143
-------
Having evaluated the data by Woodiwiss' method, stations with a
biotic index 7-9 were placed into the category of clean waters, 5-6 into
moderately polluted, 4 was considered polluted, and 2-3 indicative of
dirty waters. Extreme gradations, especially clean (Index 10) and very
dirty waters (Index 0-1) were not encountered. This evaluation of the
degree of pollution reflects rather objectively the actual situation in
parts of the rivers investigated. The calculated values of the biotic
index showed a good correlation with such chemical indices of pollution
as BODs, and biochromate oxygen consumption (Figures 1 and 2). These fi-
gures show that the value of the index regularly decreases with an in-
crease in BODs and COD.
It is especially important that such an essential factor for distribu-
tion of benthic animals as bottom type did not interfere with the evalua-
tion of the degree of pollution when using Woodiwiss1 technique. Thus,
samples taken on the cleanest station on the Izhora from silt gave the
same high index value of 5 as those taken from stones on the same station.
Samples taken from the same bottom types, subject to different degrees of
pollution, had different values of this index. For example, samples
taken from stones in various parts of the stream had indices of 4 and 9;
from clean sands, 2.5, 7; from silted sands, 2-6; and from silts 2-5.
This provides an assurance that the values depended primarily on the de-
gree of pollution. The great advantage of the method of Woodiwiss is its
simplicity, it does not require identification of species of benthic ani-
mals.
However, when using this method, one should remember that under the
conditions of sparse fauna, especially on pure sands, more samples must
be collected for more correct evaluation. Otherwise, unjustifiably low
values of the biotic index may be obtained.
The method of Woodiwiss has been used on some English and French
rivers. Investigations associated with this study have shown that it may
be used on the water-bodies of the West, North-West and Center of the
European USSR. For wider applicability over all of the USSR, it is
necessary to perform special investigations of the fauna of different
area. In this way the peculiarities of the fauna in various zoogeogra-
phical regions w.ill be considered.
Recently, indices of species diversity are used to evaluate the
degree of water pollution. Among the most frequently used is the Shannon
calculation of the Wilm and Dorris index. This index was applied to the
present study on samples of phytoplankton, zooplankton, zoobenthos as a
group, and separately for the chironomid larvae. It was found that the
values calculated by this index are not solely a function of the degree
of pollution. The diversity index calculated by the composition of
zoobenthic samples exposed to similar levels of contaminants had lower
values on stations with uniform habitats. Further, when species of large
size ranges prevail, the index also becomes slower. Considerable
seasonal variation also occurs at a given station. The hatch of aquatic
insects and departure of the imago stage will have a substantial impact
on the seasonal dynamics of the zoobenthic community. Pollution is only
144
-------
10
in
Q
O
co
• IZHORA RIVER STATIONS
A MOSKVA RIVER STATIONS
O PREGOLYA RIVER STATIONS
i
i i
i i
4 6
BIOTIC INDEX
10
Figure 1. Correlation between Woodiwiss' biotic index and BODC
145
-------
40
O 30
o:
o
20
O
DC
o 10
00
0
MOSKVA RIVER
STA TIONS
PREGOLYA RIVER
STATIONS
I I I I I I I
0
246
BIOTIC INDEX
8
Figure 2. Correlation between Woodiwiss' biotic index and bichromate
oxygen consumption.
146
-------
one of the possible causes of a decrease in the species diversity index.
Thus, this methodology may only be used in conjunction with other tecni-
ques as one of the comparative methods.
Analysis of the data obtained in this study enabled the development
of a new methodology for assessing the level of water pollution.
In various systems advocating the use of indicator organisms, atten-
tion is usually directed to the macrobenthos, ignoring the organisms of
meiobenthos. However, meiobenthic organisms may serve as good indica-
tors of the degree of water pollution. The investigations of the com-
position of meiobenthos performed by S.Ya. Tsckholikhin as a part of this
study on various parts of the Moskva River have shown that the representa-
tives of two subclasses of nematods (Adenophorea and Secernantea) may be
successfully used as indicators of pollution. The subclass Secernantea
tend to occur in places containing large quantities of organic matter.
Adenophorea, however, prefer unpolluted waters. Ratios of the numbers of
the representatives of these subclasses may serve as an index of the pre-
sence and degree of pollution. It is apparently sufficient to identify
these organisms to the order classification. This, of course, presents
no serious difficulties. Further, as a result of their world-wide distri-
bution, no geographical restrictions are imposed for using nematods as
indicators of pollution.
In the lists of indicator organisms proposed by different investiga-
tors, the number of the chironomid larvae does not exceed 10, and most
frequent used are the larval forms identified to genus. Some representa-
tives of the family Chironomidae are considered to be most numerous in
polluted water, e.g., C/uAonomoi, PioclaciwA} P&ectto&m^ptM. Further
definition of the use of species of chironomid larvae as indicators of
water pollution is presently impossible because of the lack of taxonomic
detail for this group and the need for further understanding of the
ecological requirements for separate species of this group.
It does appear, however, that the use of universally occurring
chironomid larvae holds real potential for hydrobiological analysis.
Investigations by E.V. Balushkina as a part of this study have shown that
a regular change in the ratio of chironomid larvae belonging to the sub-
families Chironomidae, Tanypodinae, Orthocladiinae takes place in
polluted waters. Clean waters are dominated by Orthocladiinae larvae,
and polluted waters by Tanypodinae larvae. A pollution index (K) may be
developed based upon the relationship between the representatives of
these three subfamilies:
„ °t + 0.5 ach
i\ =
°
-------
The value a = N+10, where N is relative abundance of individuals of
each of the subfamilies in percent of the total abundance of the chirono-
mid larvae. The value 10 is introduced to set limits for changes in the
value of the index, K. For example, an increase of this number leads to
decrease in the range of possible values of K, and simultaneously to de-
crease in its sensitivity. At 10, an optimum relation of the gradation
of the index and its sensitivity is attained. Since in clean waters the
relative abundance of the Orthocladinae larvae is close to 100 percent,
and in the most polluted waters the abundance of Tanypodinae larvae
approaches 100 percent and the larvae of the subfamily Chironomidae in-
habit both clean and polluted waters, the indicator value of chironomids
for the evaluation of the index K is reduced to one half.
Possible changes in the value of this index in natural waters lie
within the limits of 0.09 and 21. Determination of the value of this
index of subfamily composition of the chironomid fauna in the rivers
studied in this investigation, and in other reports (Gromov, 1950), have
shown regular increases with water pollution. In the cleanest waters, K
values varied from 0.136 to 1.08, and in the most polluted, from 0.9 to
11.5. Identification of chironomid larvae to subfamily is not difficult.
Estimation of the index value K is relatively simple, and it appears to
accurately reflect the degree of pollution of a river.
A critical review of methods using oligochaets for evaluation of
water quality has shown that the most suitable index is that of Goodnight
and Whitley (1961). In the opinion of N.P. Finogenova and A.F. Alimov,
an index characterizing the role of oligochaets in the total biomass, but
not in the total abundance of animals may be developed in additon to
Goodnight and Whitley1 s. The value of this index increases with an in-
crease in pollution.
The littoral zooplankton community of polluted waters is characterized
by a decrease in the total number of crustaceans with an increase in
pollution. Simultaneously, as has been shown by M.B. Ivanova in this
study, a regular decrease in species composition and abundance of clado-
cerans occurs, and copepods dominate over cladocerans. In the most
polluted areas, the crustacean zooplankton is represented only by cyclo-
poids. The least sensitive to pollution appears to be EucyclopA 4e/t-
All the indices suggested by this study have the distinct advantage
of less rigorous taxonomic requirements. These indices also use broader
taxomic categories, which naturally increases their wider applicability.
It is probable, however, that evaluation of the degree of pollution can
not be based solely on these indices. They should be considered as
supplemental, since the validity of each may be different in various
situations.
Further investigations are required to make the methods of hydro-
biological analysis more exact, to determine the most substantial systems
of analysis, and to clarify both the lists and indicator value of separate
species under various conditions, and in various geographical regions.
148
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It is impossible to develop any system of hydrobiological analysis suit-
able for all conditions and all communities of aquatic organisms. When
working with different communities, under different conditions, and with
different purposes, it is necessary to use different methods, choosing
the most suitable for a given case.
REFERENCES
Carr, J. and J. Hiltunen. 1965. Changes in the bottom fauna of western
Lake Erie from 1930-1961. Limnol. and Oceanogr., V. 10.
Goodnight, C.J. and L.S. Whitley. 1961. Oligochaetes and indicators of
pollution. Proc. 15th Ind. Waste. Conf., Purdue Univ. Ext. Ser., V.
106.
Gromov, V.V. 1950. Benthic fauna of the Kama from the mouth of the
Belaya to junction with the Volga. Izvest. Estestv. nauchn. in-ta
pri Molotovsk. univ., V. 13, N. 1.
King, D.L., and R.C. Ball. 1964. A quantitative biological measure of
stream pollution. J. Wat. Pollut. Control Fed., V. 36.
Parele, Z.A. 1975. Oligochaets of the mouth areas of the Daugava and
Lielupe Rivers, their importance for sanitary biological evaluation.
Thesis. Tartu.
Sladecek, V. 1973. System of water quality from biological point of
view. Ergebnisse der Limnologie, Hf. 7. Archiv fur Hydrobiol.
Beiheft, B. 7.
Woodiwiss, F.S. 1964. The biological system of stream classification
used by the Trent River Board. Chemistry and Industry, V. II.
Zahner, R. Beziehungen zwischen dem Auftreten von Tubificidae und
Zufuhr organischer Stoffe im Bodensee. Intern. Revue ges. Hydrobiol
B. 49.
Zahner, R. Organismen als Indicatoren fur den Gewasser zustand. Arch.
Hygiene und Bacteriologie, B. 149.
149
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SECTION 15
STRUCTURAL AND FUNCTIONAL CHARACTERISTICS OF SESTON
AS INDICES OF WATER POLLUTION
A.P. Ostapenya
Plankton has long been a traditional part of ecological investigations
related to water pollution. Accumulated data show convincingly that ses-
ton, including planktonic, detrital and mineral suspensions are an impor-
tant, distinct structural component of aquatic ecosystems, functioning as
a single entity. Seston actively influences the quality of water. This
influence is diverse, and is evident by its action on the production and
destruction stages of the biotic circulation.
Structural and functional characteristics of the seston are suffi-
ciently sensitive for use as an indicator in the evaluation of water pol-
lution.
The concentration of suspended substances in unpolluted waters may
vary within a rather wide range. A survey of literature values shows
that the concentrations of seston in unpolluted lakes, depending on
trophic type, varies from 0.1 to 70 mg dry wt/1. In rivers, even greater
concentrations of the suspensions may be observed. However, in spite of
differences in content of suspended materials, each may be characterized
by definite mean concentrations of seston. Since, in general; the influx
of nontoxic pollutants causes an increase in the content of suspended
substances in water, the zones of pollution within a water-body may be de-
lineated by the increase in concentration of the seston. The River
Svisloch serves as a typical example of a polluted stream. As a result
of year round observations in 1973, it has been demonstrated that at all
stations situated both above and below the source of pollution, no regu-
lar seasonal changes in concentration of seston are observed. This
apparently suggests that autochtonous suspension plays a minimal role,
since its concentration is closely associated with seasonal changes in
the production processes. On the cleaner parts of the river, the concen-
tration of seston varied from 7 to 25 mg/1. Below the outfall, the con-
centration was greatly increased. Further downstream, approximately 50
kilometers, the content of the seston in the water continued to increase,
apparently at the expense of heterotrophic synthesis. During all seasons,
seston concentrations were approximately three times greater than the
unpolluted sections of the river. The distribution of seston at various
stations on the River Svisloch during August of 1973 is shown in Figure 1.
150
-------
The extent of the polluted zone is evident from the concentrations of sus-
pended material in the river. Approximately 160 km below the source, ses-
ton values again achieve levels characteristic of non-polluted portions
of the river.
Depending on the intensity of the production processes and the pre-
sence of nontoxic substances in waters, the correlation between dissolved
and suspended organic matter varies markedly. Usually, dissolved organic
substances in unpolluted waters greatly exceed suspended materials. The
processes of eutrophication and pollution of waters leads to a consider-
able increase in suspended organic matter. Thus, in mesotrophic water
bodies, suspended organic matter constitutes about 10 percent of dissolved
materials, while in eutrophic waters, the relative content of suspended
matter reaches 80 percent or higher. In heavily polluted streams, the
content of suspended substances may greatly exceed the content of dis-
solved organic matter.
In the unpolluted part of the Svisloch River, the seston makes up 39
percent of the dissolved organic matter. In a significant part of the
river below the source of pollution, the relative content of seston rises
to 14 percent of dissolved organic matter (DOM). As a result of self-
purification processes, within 60 m, the relative content of the seston
drops to 17 percent. Thus, the correlation between dissolved and sus-
pended organic matter may serve as a convenient indication of the zones
of eutrophication and pollution.
Relative chlorophyll content in seston is also a rather sensitive in-
dex of pollution, and may be used for evaluation of the influence of pol-
lution on aquatic ecosystems. Generally, the relative chlorophyll con-
tent in the seston of the polluted zone is higher than in a clean part of
the river. Table 1 shows data on the relative chlorophyll content in the
seston of three rivers in Belorussia. All the three streams are moder-
ately polluted by domestic and industrial wastes. The relative chloro-
phyll content in the seston below the source of pollution increased in the
Pripyat, Western Dvina, and Neman by 25, 100, and 70 percent, respec-
tively.
Under the stress of heavy pollution by waste waters, the relative
chlorophyll content in the seston may notably decline as a result of large
quantity of allochtonous suspensions. Data on the relative chlorophyll
TABLE 1. RELATIVE CHLOROPHYLL CONTENT OF SESTON IN CLEAN AND
POLLUTED PARTS OF RIVERS
Chlorophyll in Seston, % ~~
River Clean Part Pol luted Part
Pripyat
Western Dvina
Neman
0.040
0.052
0.224
0.050
0.108
0.390
151
-------
X
Q_
o
DC
O
0.4
0.3
O I-
£2
> H 0.2
H- 2
< O
-I o
LU
" 0.1
O
DC
LU
Q_
0
STATIONS:
1 AND 2 ARE ABOVE THE ZONE OF POLLUTION
3, 4 AND 5 ARE IN THE ZONE OF ACTIVE DECOMPOSITION
6, 7 AND 8 ARE IN THE ZONE OF RECOVERY
2345 6
SVISLOCH RIVER STATION
8
Figure 1. Distribution of seston in the River Svisloch
in August of 1973.
152
-------
content in the Svisloch River are given in Figure 2. Above the source of
pollution (Stations 1 and 2), the fraction of chlorophyll in the seston
made up 0.2 percent. In the polluted zone (Stations 3, 4, and 5), it
dropped to hundredths of a percent. Downstream, the processes of self-
purification lead to a considerable increase in the chlorophyll content
in the seston.
Specific Oxygen Consumption (SOC) is an important functional index,
characterizing the biological activity of seston. It represents the
amount of oxygen consumed by a unit mass of suspension, per unit time.
According to Margrave (1972) SOC values for suspensions of various ori-
gin, composition and degree of dispersion lie within a range of 0.002-
0.240 mg 02/mg organic matter per day. While the relative values of SOC
are apparent, for many scientific and practical purposes, including the
evaluation of water pollution, it is necessary to fully understand how
this index depends on the trophic status and the pollution degree. In
this regard, a determination of SOC by seston has been made on lakes of
various types polluted to varying extent by domestic and industrial
sewage.
0.4
0.3
I
Q_
O
DC
O
X
O I-
-5
> K 0.2
h- Z
< O
-1 O
LU
DC
I-
•z.
LLJ
O
DC
LU
Q_
0.1
0
STATIONS:
1 AND 2 ARE ABOVE THE ZONE OF POLLUTION
3, 4 AND 5 ARE IN THE ZONE OF ACTIVE DECOMPOSITION
6, 7 AND 8 ARE IN THE ZONE OF RECOVERY
234 5 6
SVISLOCH RIVER STATION
8
Figure 2. Relative chlorophyll content in seston of the River Svisloch
in June of 1973.
153
-------
The mean SOC values in three lakes are given in Figure 3. In the meso-
trophic waters of Lakes Naroch, the seston SOC was the highest averaging
0.062 mg 02/mg organic matter per day for the vegetation season. With an
increase in the trophic status, the SOC notably decreases, reaching a
level of 0.036 mg 02/mg/day in eutrophic Lake Myastro, and 0.016 mg 0?/mg/
day in ultra-eutrophic Lake Batorin.
The SOC of seston with a greater fraction of allochtonous organic mat-
ter as a result of sewage pollution is demonstrated by studies of the
Svisloch River. In clean waters of this river, the SOC made up 0.08 mg
02/mg of organic matter per day. In polluted portions, the value
obtained was 0.147 mg 02/mg of organic matter per day. A similar SOC
value (0.13 mg 02/mg/day was observed for a polluted part of the Dnieper
River.
In general, the data suggests that concentrations of SOC in unpolluted
waters range from 0.01 - 0.10 mg 02/mg of organic matter per day. In the
presence of easily degradable allochtonous materials, the SOC values asso-
-
"
0.06
c/)
o <=
o.y
2 g 0.04
LU O)
CD o
^~ 01
CN
Ho 0.02
LU
Q.
CO
0
Lake Naroch
(mesotrophic)
Lake Myastro Lake Batorin
(eutrophic) (highly eutrophic)
Figure 3. Specific Oxygen Consumption (SOC) by seston in three
representative lakes.
154
-------
ciated with seston increase to 0.15 mg Og/mg of oganic matter/day. It
appears that the metabolic activity of river seston is higher than that
of the lakes.
The photosynthetic activity of phytoplankton is an important
component ofthe living fraction of seston. It depends on the quality of
water, and may be used as an index of pollution. Under the influence of
moderatepollution by nontoxic wastes, photosynthesis in the zone of
pollution rises, apparently at the expense of enrichment of the water in
nutrients. The Neman River graphically demonstrates this reaction in
response to an influx of sewage (Figure 4). Below the inflow,
photosynthesis increased by three fold. Depending on the intensity and
character of pollution, photosynthesis either increase or decrease
markedly. In all cases, a notable deviation of photosynthetic activity
from the average level characteristic of clean waters is observed in the
polluted zone, it should also be noted that such generally accepted
indices as oxygen consumption is more highly variable than photosynthetic
activity, and yields less precise results when pollution levels are low.
15
co
3 10
CM
0
en
£
^ 5'
<
o
•^*\
i S~-
~J . -
'• x'*^-^ >A
x y ^x • ' \ _j
/ \ XX \ -
/ ^ / \.
' ^ \ :
i i i i i i i i i i ^r
1.2 |~
1 0 ^
i »\j \.^i
O
_ _ en
0.8 E
n« Q
123456 7 8 9 10 11 12
STATION
Figure 4. Relationship of photosynthesis ($) and distruction (D) in
the Neman River in August of 1975.
155
-------
In the past, investigators have used the ratio of phytoplankton photo-
synthesis to oxygen consumption by seston and by dissolved organic matter
(/D). This ratio is also a rather variable index, and apparently, can
not be used for the evaluation of pollution. Much better results are ob-
tained by the use of the ratio $/R, where R is total consumption of oxygen
by the water and sediments. Figure 5, the ratio /D and $/R for the
Pripyat River are compared. The ratio $/R serves as a better indicator
of pollution than the ratio of $/D (Figure 5, Stations 1 and 7).
Thus, the structural peculiarities and functional indices of seston
reflect the biological activity of suspended substances, and react mea-
surably to pollution. As a result, they may be used for evaluation of
the degree of pollution and estimation of water quality. However, be-
cause of the diversity of pollutant types and variations between water
masses, some variation in measured values can be expected. In some in-
stances, an increase in the index value is observed, and in others a de-
crease is noted. It is important, however, that in all the cases, a not-
able deviation from the average clean water statistical norm is observed.
Q
12 345 6
7 8 9 10 11 12
STATION
Figure 5. The ratio of */D and $/R on the Pripyat River.
156
-------
The use of structural and functional indices of seston for evaluation
of water quality holds promise. Many parameters characterizing suspended
matter and its participation in the biotic circulation can be determined
by present instrumentation and, thus, even finally may be incorporated in
programs of automatic sampling of the control of water quality.
157
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TECHNICAL REPORT DATA
(Please read Instructions on the reverse before completing)
REPORT NO.
EPA-600/3-78-076
!™ANDSUBTITLE Proceedings of the First and Second USA-
USSR Symposia on the Effects of Pollutants Upon Aquati
Ecosystems Volume I-Duluth, Minnesota, USA Symposium-
3. RECIPIENT'S ACCESSION-NO.
4.
5. REPORT DATE
August 1978 issuing date
6. PERFORMING ORGANIZATION CODE
7. AUTHOR(S)
Environmental Protection Agency - USA
Soviet Academy of Sciences - USSR
8. PERFORMING ORGANIZATION REPORT NO.
9. PERFORMING ORGANIZATION NAME AND ADDRESS
US-USSR Joint Agreement on Environmental Protection-
Project .02-.02-1.3, Effects of Pollutant Upon Aquati
Ecosystems and Permissible Levels of Pollution.
US EPA, Grosse He, Michigan 48138
10. PROGRAM ELEMENT NO.
1BA769 1BA608
•11. CONTRACT/GRANT NO.
1? SPONSORING AGENCY NAME AND ADDRESS
Environmental Research Laboratory - Duluth, MN
Office of Research and Development
U.S. Environmental Protection Agency
Duluth, Minnesota 55804
13. TYPE OF REPORT AND PERIOD COVERED
In-House
14. SPONSORING AGENCY CODE
EPA/600/03
15. SUPPLEMENTARY NOTES
Prepared in cooperation with the Institute for the Biology of Inland Waters,
Soviet Academy of Sciences, Borok, Jaroslavl Oblast, USSR.
16. ABSTRACT
This publication represents the proceedings of two symposia conducted
jointly by the US EPA and the Academy of Sciences of the USSR. The first symposia
(Volume I) was held in Duluth, Minnesota, USA on October 21-23, 1975, and the
second (Volume II) was held in Borok, Jaroslavl Oblast, USSR during June 22-26,
1976. The published papers from these symposia contain both broadly based review
papers, designed to familiarize attendees with a wide cross-sectional representa-
tion of ecologically related activities in each country, and narrowly specific
state-of-the-art scientific discussions. The presentations focus upon methodology,
historical aspects, microbial and abiotic degradation processes, trace metal
problems, effects of toxicants, proposed species indices, and studies of fate and
transport of pollutants.
7.
KEY WORDS AND DOCUMENT ANALYSIS
DESCRIPTORS
b.IDENTIFIERS/OPEN ENDED TERMS
c. COSATI Field/Group
Freshwater
Phosphorus
Nitrogen
Pesticides
Fishes
Stream Flow
Bioassay
Communities
Phytoplankton
Nutrients
Waste Treatment
Toxic Substances
Macrobenthos
Microbiota
Water Quality Criteria
Great Lakes
Max imum
Permissible
Concentrations
57H
68D
3. DISTRIBUTION STATEMENT
Release to Public
19. SECURITY CLASS (This Report)
Unclassified
21. NO. OF PAGES
412
20. SECURITY CLASS (This page)
Unclassified
22. PRICE
EPA Form 2220-1 (9-73)
158
U. S. GOVERNMENT PRINTING OFFICE: 1978 — 757-140/1435
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