United States
          Environmental Protection
          Environmental Research
          Duluth MN 55804
          Research and Development
of the First
and Second USA - USSR
Symposia on the Effects
of Pollutants
Upon Aquatic
          Duluth , Minnesota
          USA Symposium
          October 21 - 23 ,1975

          Volume II
          Borok , Jaroslavl Oblast
          USSR Symposium
          June 22 -26 ,1976


 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.

                                             August 1978
               SYMPOSIA ON THE
Volume I:   Duluth, Minnesota, USA Symposium
            October 21-23, 1975

Volume II:  Borok, Jaroslavl  Oblast, USSR Symposium
            June 22-26, 1976
          DULUTH, MINNESOTA  55804

                   VOLUME I

               SYMPOSIUM ON THE
             October 21-23, 1975
              Duluth, Minnesota
                  Edited by

               Donald I. Mount
           DULUTH, MINNESOTA   55804

    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.

                           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.

                           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.


    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,

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

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容xperiment  1	141
 13       Change in component values in time容xperiment  2	142
 13       Change in component values in time容xperiment  3  	  143
 13       Change in component values in time容xperiment  4	144
 13       Change in component values in time容xperiment  5	145
 13       Change in component values in time容xperiment  6	146
 13       Change in component values in time容xperiment  7  	  147
 13       Change in component values in time容xperiment  8	148
 13       Change in component values in time容xperiment  9  	  149
 13       Change in component values in time容xperiment  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

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

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

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

    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

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
         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

11.   Toxic Organic Residues in Fish
       Howard E.  Johnson	    115

12.   Balance of Organic Matter in the Ecosystem of the Rybinskiy
       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,
       Carlos M.  Fetterolf, Jr	    219

                                 SECTION 1


                              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預l I except for drinking purposes.
 1-USSR,  Biology  Faculty  of Moscow State  Univ.


     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

     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

               10    15   20



                                                      IO    15   20
Figure  1.   Processes of mineralization in solutions with triethyl stan-
           nic  chloride.  K幼ontrol;  10.1  mg/1; 21 mg/1; 310 mg/1;
           450 mg/1.  Along the abscissa妖ays  of the experiment.

Key:  a.   Triethyl stannic chloride.

     b.   BPK [biokhimicheskoye potrebleniye  kisloroda, BOD biochemical
          oxygen demand].

     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 10100 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 35 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.


                Table 1.  PERMISSIBLE CONCENTRATIONS (tng/1) OF TOXIC AGENTS FOR BACTERIA  (saprophytic
                          and nitrifying)
Formation of N02
Formation of N03



                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






             6     10     14
                   10    14
18    22
Figure 2.   Processes of mineralization  in solutions of pyror-70, K幼ontrol,  BOD:
           110 mg/1; 21 mg/1;  350 mg/1; 4100 mg/1;  5300 mg/U 11  mg/1;
           210 mg/1.  Along the  axis  of the abscissa妖ays of the test.
Key:  a.  BOD.




/ \
/ \
/ \
- / \
/ \
; \
/ \
/ \
i \
i \




                                           12 16 20 24 28 32 36  40 44 48 50
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:   1100 mg/1; 210 mg/1; 31 mg/1;
N02:   110  mg/1; 2100 mg/1; 31 mg/1;
N03:   110  mg/1; 11 mg/1;  3100 mg/1.

            Along the axis of the abscissa妖ays  of  the test.

Key:   a.  BOD.
                     Tests in

    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.

    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.

                                 SECTION 2


                            Clarence M. Tarzwell

    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.

    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.

 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葉heir 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


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

    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


 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

     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-


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

    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.


     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

     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.

    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


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

    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


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

    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


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

    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

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.


    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


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.


    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.

    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


personnel engaged in extensive field investigations of the effects of
malaria control熔il 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


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.


    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


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-

    2)  requirements would apply to all sections of the country,
        and thus there would be no incentive to move to escape

    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


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

    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


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

    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.


    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


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.

    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.

    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.

                                 SECTION 3


                      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.


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 -


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


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

    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.

                                 SECTION 4

                           ENDPOINTS  IN BIOASSAY

                            Lloyd L.  Smith, Jr.
    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.


    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


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,




                                           TIME - DAYS
              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.

    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


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

    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-






    ^ 200

    8  100

    CJ   80

i  <ー
                   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

                           TEMPERATURE   (C)
Figure 3.  Tolerance of bluegills (Legomis macrochirus) to molecular cyanide
          (HCN) at various temperatures.  (Dotted portion of curve is

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.

    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.

        96-hr LC50       MATCd      Application
        (yg/liter)    (yg/liter)      factor
Fathead minnow
  (Mount and Stephan, 1967)
(Organophosphorus  pesticide)

         9,000         200-580

            89         3.6-7.1
                      (Organophosphorus pesticide)
Fathead minnow
  (Adelman, Smith, and
  Siesennop, 1976)
Fathead minnow
  (Smith, unpublished)



Fathead minnow                 7,200
  (Pickering and Gast, 1972)
  (Eaton, 1974)
Fathead minnow
  Soft water
    (Mount and Stephan, 1969)
  Hard water
    (Mount, 1968)

Brook trout
  Soft water
    (McKim and Benoit, 1971)




0.33-0.55    0.00017-0.00027
5.3-17.3     0.049-0.162
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.


Brook trout
Rainbow trout
Fathead minnow
             (EXPRESSED AS yg/LITER)a
Black bass (largemouth)
Rainbow trout
Coho salmon
  Yellow perch                 9         3,060             13
aMacek and McAllister, 1970.


    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.

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:

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.

                                 SECTION 5


                              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.

    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-


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羊apid motor
activity擁n anesthetized crucians  (novocain, urethane).  In other words,
generalized inhibition of the central nervous system, resulting from narco-


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

    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

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


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傭y
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).


  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,

  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.


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,

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.

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,

31.  Lukyanenko, V.I.  In:  Experimental Water Toxicology.  Riga, Issue 4,
        1973, 9-29.

                                 SECTION 6


                              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

    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


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


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

    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


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-

    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.

    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

    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

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


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


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


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.

    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.

    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


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.


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.

Doudoroff, P-  1969.  Discussion of paper by D.I. Mount, "Developing Thermal
    Requirements for Freshwater Fishes,".  Irア 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.

                                  SECTION  7

                             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葉hat 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.

    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葉hat 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.

Figure 1.  Platform from which oyster trays are suspended.



Figure 2.  Oyster tray.

    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擁n  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


                       40  p
                  I I




     B   15



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
                                       Figure 3.   The structure of a  natural  diatom community.

               OCTOBER 1953 TO JANUARY  1958

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





(Apr. 1954-1958
Species in






    27 '
Species ,






Species in theo-
retical universe







                 I I
                      40  r

            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
                     Figure  4.  The  structure of a  diatom community  under  the effects of pollution  high
                                 in nutrients.







             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-


        cinder block
Figure 6.   Invertebrate sampler.

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).

    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

    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.



Beak, T.W., T.C. Griffing, and A.6. Appleby.  1973.  Use of artificial sub-
    strate samplers to assess water pollution.  Irア 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-

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-

                                  SECTION 8

                               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


    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


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.

    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-

    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.

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).

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).

                                 SECTION 9


          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

    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.


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庸rom both point and non-point sources幼ould 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.

    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輸 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 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-

    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).

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.
 Organization for Economic Cooperation and Development.

    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

    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






              vs. LAND USE
         悠	1	T	1	T~

Figure 1.  The relationship between  mean total  nitrogen concen-
trations in streams and  land use in  the Eastern United States.

                      vs.  LAND  USE





             0,05            OJO
 Figure 2.   The relationship between mean total phosphorus concen-
 trations in streams and land use in the Eastern United States.

     100 ^
"I	1 I  I (Hl|	1	1 Mill!)	1	1 I  II II Tj	1	1 I  I I III
         :   A MESOTROPHIC
      10 -    OLIGOTROPHIC
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

    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.

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.

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,

    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

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.


    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

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佑hemical 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-

    Mechanical harvesting輸 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

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輸nother 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悠t 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-

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輸 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

    Drawdown祐ediment 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裕he 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.

    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.


    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

    Minnesota has passed legislation to set effluent standards for phos-
phorus in municipal discharges.  If the discharge enters a lake directly,

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-

    Indiana, Iowa, and Minnesota have adopted various kinds of farm-animal-
waste regulations, aimed chiefly at problems related to large feedlot

    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.


    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

    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-

    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

   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.

    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."


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.

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.

                                 SECTION 10


             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.


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

    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-

    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).

    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

    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-


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擁n particular, non-coordination of  energetic and


tion of
algae (g/1
of live

cal state

ni trogen
ration of
1 fish

per 24 hrs
in organism
of 1 fish (%
of the ave-
rage daily
by one fi

wi th li-
quid ex-

excreted during 24 hrs
sh (% of the average
nitrogen ration)

with ex-




                     YEARLINGS (mg/g of Fresh Tissue)
Ami no







Lei cine

M ア m











M ア m
C. 13+0. 02
0.01 ア0.00











M ア m












Pike perch

Pike perch
M ア m




Acti vi ty
M ア m

709. 23ア 8.50

517. 64ア 9.54
M ア m
1.90 ア0.10
of paraly-
M ア m




.62+ 1
.41 ア28


Death stage
M ア m




M + m

1095. 80ア 8

777. 69ア 1



                     TABLE  4.   EFFECT  OF  BLUE-GREEN  ALGAE  ON  CONTENT  OF TOTAL THIAMINE  (yg/g)
                               THIAMINASE ACTIVITY  (ng/hr)  IN LIVER AND INTESTINE OF  FISH
erythrophtalnus L.

erythrophtalnus L.


(no "bloom")
("bloom" period)
2. 3U0.18


plastic processes and changes in the above mentioned factors  including
blood analysis妖evelop 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-

    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.


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"

Metelev, V.V.   1971.  Toxicity of pesticides for fish, effect mechanism,
    indicator methods.  Ekspep. Vodn. Toksikologiya, Vol. 2.  "Zinatne"

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.

                                 SECTION 11

                       TOXIC ORGANIC RESIDUES IN FISH

                             Howard E. Johnson

    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.

    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


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.


    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.

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-

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

    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.

    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.


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.

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.

                                 SECTION 12

                      BALANCE OF ORGANIC MATTER IN THE

                               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).


    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.

          Breytovo  )
Fig. 1.  Map-diagram of Rybinskiy  reservoir.  Figures  indicate
        number and location of standard stations.

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

   w  IOO
Fig.  2.  Average values  for phytoplankton production for 13 years
        in the Rybinskiy  reservoir.

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

    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).

              If) (0
Fig.  3.  Average values of the number of  bacteria  according  to
         data for 2 years of standard observations.

                                   (in C)
During the navigable period (May-November)



In 1
1 i ter ,

1 mg2,
For the
1 1 7000
1 74000
1 1 7000
Mean for 24 hours

In 1

1 mg2,
For the
    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.
                                   (in C)
During the navigable period (May-November)
In 1
1 i ter ,
1 mg2,
For the
Mean for 24 hours
In 1
1 i ter ,
1 mg2,
For the

    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


                        MATTER IN THE RYBINSKIY  RESERVOIR
For the whole Under
Element reservoir x 103, 1m2, % Note
ton C g C
Gross production of
Production of
Bacterial assimila-
tion of C02
Input of organic matter
from the watershed and
atmospheric precipita-
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
Difference between input
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,
2.4 1.2 Mean for 5 years
116 58 According to data
for 1965
129 62 Mean for 5 years
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
and expenditure of
organic matter

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.


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.


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.

                                 SECTION 13


                                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-

    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.


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

    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

    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


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).


    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.


Time of establ
steady state


    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熔rganic matter-bacteria:
            =D(Si-S) -klSH
           - = k2SH - DH
    For the system熔rganic matter-bacteria-infusoria
            = D(SX-S) - kjSH - k3SP
           -= k2SH - DH -
         J= k5HP + k6SP - DP
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熔rganic 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):

         ^r= D(S,-S) - kjSH - k3SP
   = k2SH - DH -
     k HP + k」SP - DP
      5      6
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熔rganic matter-bacteria:
:}ァ = k,NH - DH
dt    2
                      - k3NH
    For the system--organic-matter-bacteria-protozoa:


   = DINj-N) - kjNH +
     k2NH - DH - k5HP
     k6PH - DP
         ー|= D(SrS) - k3NH
where N: is the concentration of nitrogen  in the  inflowing medium, N is the
concentration of nitrogen  in the reactor,  and the remaining designations

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.

Initial con-
of nitrate

rate of

of bacteria
mg/liter of
dry weight
Growth of
in dry
weight for
10 hrs,

for 10
    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"

(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


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.

 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.


 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),

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.



15  I
10  S

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.


                   ft \

                      / ^.
                        *   V          \>*'"*'       s-  -*
                      v.    /
      2    4   6   8   10   12   14   16   18   20  22  24   26  28  30

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.

Fig.  3.   Change  in  component values in time  experiment 3.
         Symbols are  the  same as for figure 1.


                  廟 .-' '
         2   468    IO   12   14   16    18   20   22  24

Fig.  4.   Change  in  component values in time  experiment 4.
         Symbols are  the same as for figure 2.

Q. 022
                              8     10     (2     14

2O    22    24
    Fig, 5.  Change  in component values in time   experiment 5.   Symbols are the
            same as for figures 1,  3.



" x/
x^r^ 	 	 ー

" ***"*'^~^^-~^*^~^
^^ <^^ ォ^^


         2   4   6    8   10   12    14   16   18  20   22  24
Fig.  6.   Change  in  component values in time -- experiment 6.
         Symbols  are  the same as for figures 2, 4.

Fig.  7.  Change in component values in  time  experiment  7.
        Symbols are the same as figures  1, 3, 5.

         -9  -

                                                       H    P
*"""風6  -
                                                       -3   -
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.

           I      2345678
 Fig.  9.   Change  in component values in time   experiment  9.
          Symbols are the same as for figures 1, 3,  5,  7.

                                               H   P
                                                 12  -
                                                 9  1
                                                6  -
                                                 3  -
    12345      23    24   25
Fig. 10.  Change in component values in time -- experiment 10.
         Symbols are the same as for figure 8.

  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

                                 SECTION 14


                               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.
Biological Faculty of Moscow State University


    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-

    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,

    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-


                Invertebrates |*
Fig. 1.  Block-diagram of the basic functional
         links in the reservoir
e. Invertebrates
f. Algae

Figure 2.   Change in  numbers of a species with exposure to a
toxic substance.   Designations: K--control; l--low concen-
tration耀timulation;  2--concentration of substance at which
95-97% of individuals  dies.  The individuals which survive
re-establish the  population; 3幼omplete death.

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

    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


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

    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


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葉he 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.


   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


Chlorella vulgaris
Scenedesmus quadri-
Anabaena variabilis
Elodea canadensis
Lemna minor
Limnaea stagnalis
Planorbis planorbis
Planorbis planorbis
Planorbis planorbis
Tubifex tubifex
Daphnia magna
Daphnia magna
Daphnia magna
Daphnia magna
Daphnia magna
Chironomus plumosus
C. plumosus larvae
C. plumosus larvae
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
Number of eggs laid
Survival rate of
Survival rate
Survival rate
Reproduction of
adults of initial
Number of genera-
Number of offspring
Survival rate
Embryonic develop-
Hatching of fry
Survival rate of

Salmo gairdneri
iridens (yearlings)
S. gairdneri iridens
Toxicants:   l~salicylanilide;
2耀odium 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

                                 SECTION 15


                              William C. Klein
    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 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,


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.


    Various chemicals  (rotenone,  toxaphene, creosol, copper  sulfate, and
sodium cyanide) have  been used  in fish sampling.  The most  acceptable  of


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

    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.

    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-


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葉he gizzard shad, emerald shiner, fresh-
water drum, channel catfish, bullhead, and the drum謡ere 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

    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


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.

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.

                                 SECTION 16

                        REGISTRATION OF PESTICIDES:

                           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葉hat 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-


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

    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

    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;


                            RESEARCH SYSTEMS
            ACUTE TOXICITY
            [Static tests}
            [Flow-through tests
             [pro w th_ & reprod uction
             [Adaptabi|1ty_ _*****3
    | Methods'!
                                        MUPTAKE& EXCRETION
  Physiology |
 [Pathology J
 ォPhyjKal_ _Che_miaBJ__{
Figure  1.  Organization of investigational  divisions at the  Fish-Pesticide
          Research  Laboratory,  Columbia, Mo.

    (3)  biological uptake, storage, and elimination;

    (4)  sublethal effects on growth, reproduction,  and

    (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.


    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-


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


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.

    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.


    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輸lthough  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


         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

    Biochemical effects佑ollagen  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.

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.

    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

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

    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


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.

                100 r
                                             TFM Concentration in Water
        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.

    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.

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.

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.

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

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.

                                 SECTION 17


                       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.


    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.

    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.

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

    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.


    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様arvae 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

    There are highly resistant invertebrates様arvae 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.

Degree of
                      48 hrs
                    LC50 mg/1
 Limnophilus flavicornis  Ebr.
 .Leptocerus aterrimus Steph.
 Phryganea striata L.
 Limnophilus stigma Curt.

 Baetis sp.
 Cloeon dipterum L.
 Siphlonurus linnaeanus Eat.

 Nemura marginata Pict.

 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.
 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.
           lappom'ca Mart.
           culiciformis De
                       Orthocladius  sp.
De Geer.

Degree of

  48 hrs
LC50 mg/1
  Sialis  flavilatera L.

 HEMIPTERA (image)
  Sigara  striata L.
  Gerris  lacustris L.
 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.
            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.


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

    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.


    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幼rucian carp, blue bream,  bur-
bot, bream, perch, pike,  ruffe, roach,  trout.  Regarding resistance,  the
following descending order is observed幼rucian  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


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.

    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

    One series of experiments performed in 6 aquaria may be cited as an ex-

    Aquarium 1 - without aquatic organisms

                                                               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

                                              O.I sec
                  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.

             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.

         50   100
         i   n
150  2OO  250  300  350  4OO
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.

    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

    When the toxicant is no longer introduced,  the ecosystem quickly loses
its ability to decompose it.


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.

                                 SECTION  18


                                John F. Carr

    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


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).

                           Figure  1.  The  St.  Lawrence Great Lakes with interstate and
                                 international  boundaries.


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.


                              TABLE 1.   DIMENSIONS OF THE GREAT LAKES
St. Clair
31 ,820
Average surface
elevation above
mean sea
level since 1860
          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).
                      PRIOR TO 1900  (EXPRESSED IN MG/LITER)9
      From Beeton,  1969.

    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).


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


2 ao

Q 7.0

8 6-ー

i 5.0

it 4.0
   PERIOD o)


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,

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


             5/ Lawrence
             Great Lakes
                                                   swego R
                                                   Oneido L
                                                     Mohawk R

                                                   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.
                    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.


    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

    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.


Lake trout

Lake whitefish
Cisco (chubs)
Lake herring
Erie fa
Northern pike
Blue pike








Product first
below 100,000









(1 ,000's pounds)

















L -

recovery (H,M,L)a
and reason

stocking and sea
lamprey control (s.l.c.)
stocking & s.l.c.
stocking & s.l.c.
stocking & s.l.c.







 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.

    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.
58 (-3)
143 (+12)
108 (0)
146 (+3)
149 (+6)
13 (0)
34 (0)
24 (0)
33 (+6)
34 (+10)
4 (0)
13 (+160)
9 (+50)
16 (+23)
18 (+20)
2 (0)
4 (+100)
4 (0)
11 (57)
11 (57)
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

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

    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


                      CURRENT PRODUCTION
Sea lamprey

Al ewi f e

Gizzard shad
Coho salmon

Chinook salmon

Rainbow smelt


White perch
First reported in
First recorded commercial catch
Lake Year Lake
Ontario ?
Erie ?
Huron ?
Michigan 1956 Michigan
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
1929 Erie
Ontario 1955 Ontario
commercial ,
Year Lake








1974 Production
1 ,000 pounds








Production exceeded 100,000 pounds.
bYear of record from Smith, 1972.
cYear of record from Christie,  1972.

  lakes affected
           Primary cause(s)
             of decline
Atlantic salmon     Ontario

Lake trout

Northern pike

Lake herring


Cisco (chubs)


Lake trout


Blue pike

  All lakes

  Erie, Ontario

  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,
Emerald shiner
Deterioration and blockage of
streams, exploitation

Exploitation, destruction of spawning


Destruction of spawning areas, ex-

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

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.
56 (-3)
150 (5)
110 (2)
170 (16)
172 (15)
13 (0)
34 (0)
24 (0)
38 (15)
38 (12)
4 (0)
17 (31)
13 (44)
23 (44)
25 (39)
2 (0)
5 (25)
6 (50)
19 (73)
19 (73)
3 (0)
4 (0)
9 (29)
10 (43)
        From Beeton, 1969.

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優emise 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.


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


    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).


    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


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


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.
55 (-2)
155 (3)
115 (3)
206 (21)
210 (22)
13 (0)
34 (0)
27 (12)
38 (0)
40 (5)
4 (0)
20 (43)
17 (31)
27 (17)
30 (20)
2 (0)
7 (40)
7 (17)
21 (42)
29 (53)
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

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,


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.


    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.

    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.

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.

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.

                                 SECTION 19

                         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.

                         -12-20 MONTHS
Figure 2.  Life cycle  of the sea lamprey (from Crowe, 1975)

"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).


                                        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

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

     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

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

     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.

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.



    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.


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.
    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ア M.W. Hardisty and  I.C. Potter  (ed.) The  biology of lampreys, Vol. 1.
    Academic  Press, New York.

                  VOLUME II

               SYMPOSIA ON THE
               June 22-26, 1976
           Borok, Jaroslavl Oblast
                  Edited by

               Wayland R.  Swain
               Nina K. Ivanikiw
           DULUTH, MINNESOTA  55804

    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.

                          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.

    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

   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


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 ーC and  16 ー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

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

  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 ー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 ー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 ーC    	118

  11     Growth of fathead  minnow in 112-day exposure to  hydrogen
         sulfide at 23  ー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

    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.

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

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

                                SECTION  1

                            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,


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


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.

    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.   '

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.


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.

                                 SECTION  2


                              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  ー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 ー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

    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-

                                 SECTION 3


                 Richard A. Schoettger and Wilbur L. Mauck
    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


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.
    Six experimental forest insecticides,3umithion , carbaryl, Dylox  ,
Matacilョ, Dimilinゥ (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-


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.


Acute Toxicity

    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 ー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 ーC than at
7 ー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 ーC than at
12 ー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 ーC than at 7 ー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.


                                      DIFFERENT TEMPERATURE, HARDNESS, AND pH.
(% active
(Technical, 99.5%)

(Field formulation,
49%, Sevin-4-Oilョ)

Test Water
(mg/1 as CaCOs)

40 7.5
96h-LCt^Q and 95% confidence interval
___ 2/

TABLE 1 (Continued)
(% active
(Technical, 98%)

(Field formulation,

Test Water
(mg/1 as CaCCh)

96h-LC5o and 95% confidence interval

TABLE 1 (Continued)
Test Water
(% active
(Technical, 99%)

(Field formulation,
(Technical, 94%)

Water pH
(mg/1 as CaCOs)
96h-LC5Q and 95% confidence interval


(% active
(Technical, 94%)

(Field formulation,
(Technical, 95%)
(Field formulation,

Test Water
(mg/1 as CaCQg)
1 (Continued)


96h-LCso and 95% confidence interval

                                          TABLE 1 (Continued)
(% active

(Field formulation,

(Field formulation,




Test Water
(mg/1 as CaCO^}










96H-LC50 and 95% confidence interval

 Toxicity based on active ingredient in static tests.
"Not determined.

                 STONEFLY NAIADS TESTED AT 18 ーC and 16 ーC
                                     Invertebrate and Test Temperature
Compound                             Scud                     Stonefly
(% active
ingredient)	18 ーC	16 ー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

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.

    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


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).


    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

    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

                                             CANDIDATE FOREST INSECTICIDES
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
+0.70 .

1 1 1 1 1 1 1 1 1 1 1 1
i i
i i
I i
1 i

      Tests conducted in standard reconstituted water at 12 ー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.

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 ー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.

                            Residues of                 96-h LCso  and 95%
                      Aroclor  1254 in fish             Confidence Interval
 Insecticide                   (yg/1)                           (mg/1)

 Carbaryl                     ^2                             ^


                             0.4                               5.0

                             2.3                               2.5

 Matacil                      Q>082                              n


                             0.4                                10

                             2.3                               6.5

 DylOX                        0.82                              3.8

                             0.4                               4.6

                             2.3                               3.9

 Sumithion                    Q>082                             K6


                             0.4                               1.5

                             2.3                               1.7

  Toxicity  based on active  ingredient; tests conducted with 0.3-g fish at
  12  ォC.
  Background Aroclor  1254 residues.

    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.

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):

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.).


Litchfield, J.T., Jr., and F. Wilcoxon.   1949.   A simplified method of
    evaluating dose-effect experiments.   J.  Pharmacol.  Exp.  Ther.   96:

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ョ.   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涌regon  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.).

U.S. Department of Agriculture.  1975.  Draft environmental  statement  for
    cooperative Spruce Budworm suppression project柚aine 1976. U.S.
    Forest Service, Upper Darby,  Pa.  137  p.  (Unpubl.  Rep.).

                                 SECTION 4

                              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-


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-

    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


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


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-

    The second route of influence of the abiotic factors of environment
upon the resistance of the aquatic organism to poisons  is the indirect


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,


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

    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


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


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

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.

    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


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

    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.

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.


                                SECTION 5

                            ON FRESHWATER FISH

                           Lloyd L. Smith, Jr.

    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 ー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.

    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


	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.

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 ー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 ー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 ー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 ー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.

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-


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
     oc  1000

                      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.



Mean HCN**
( g/D
Percentage Survival***
Mean Weight
  *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.


Fry/ 100
Mean Weight
Surviving Adults

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 ー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

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 ー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-


Control A3
Control B3
Control C

Mean Total Length
28 Days
56 Days
Mean Weight
56 Days
^Except where noted, chambers originally contained 80 larvae.   Numbers were
 reduced to a maximum of 40 per chamber after 28-day measurements.
 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).

                    144 DAYS BEFORE START OF SPAWNING.

Start of
Total Eggs/100 g
of Female
(12 Days)
 Formation of neural keel.

8  160

8  120


S   80
*   40
          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.

of Alevins

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.
20    30    40    50    60

    Figure 3.   Growth of brook trout in  various  concentrations  of HCN.
                   Expressed as percentage of controls.


                        TO VARIOUS LEVELS OF HCN.
Mean Length (mm)
Mean Weight
90 Days
% of
*Significantly different than controls.


    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 ー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 ー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.

    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.

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,


Fathead Minnow
Brook Trout

96 h
IpH 8.0.

Izatt, R.M., J.J. Christenson, R.T.  Pack, and R.  Bench.   1962.   Thermo-
    dynamics of metal-cyanide coordination.   I.   pK,  AHー,  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,

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.

                                SECTION 6

                          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).


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,



                                   PRESENCE OF MICROORGANISMS
                                                                                                           MICROBIOLOGICAL PROCESSES



m K










II _J_



















a co
x "~
O uj

5 5



4 MI
< H
CJ ^
-J _


                                              CHARACTERIZATION MEDIUM FOR THE PRESENCE OF
                                                SPECIFIC POLLUTANTS (FECES, HYDROCARBONS,
                                            CELLULOSE, SULFIDES, POLYTHIONATES, PERCHLORATES,
                                                    OXIDES OF CHROMIUM, SULFATES)

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.


   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

 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-

    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

    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

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-

    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.

cally controlled incubator at a preselected temperature, usually  26  ー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ー 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  ー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.


    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.


     Ratio, %            Water                 Water-Body

 0.003 or less        Very clean       Lakes Onega, Ladoga, Baikal

 0.03                 Clean            Reservoirs: Rybinskoe, Sheksninskoe
3 or greater
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-


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
 Figure 3.  Position of the test-tube (N 2) with ethyl alcohol  in the row when
             determining the spores of bacteria in mud deposits.

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

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 ー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
    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

    Presence of Perchlorates and Chlorates悠n 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サ 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佑hromates 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

    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-


    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-

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

    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


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:

         ーH ~ ーk = ーn  rag/1/day

         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.


  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 ーC.

     Water from the bottom  sampler is poured into  a flask,  its tempera-
ture is adjusted to  20  ー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:

         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.

                       WATER BODIES OF VARIOUS TYPES

  Type of Water-Assimilation of CO2,Character of
      Body	g C/1/day	Water
Very Clean
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


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

    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,

    The quantity of carbon dioxide assimilated by microorganisms is cal-
culated by the formula:
         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.

    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.




               Bacillus  S
             10-9   io~8    icr7
 Figure 4.  Effect of Ag ions on bacteria.  Analysis using the
            heterotrophic assimilation of C02.

    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 -
                             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.


         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.


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.


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.
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    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.

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                                 SECTION 7


                  Rosemarie C. Russo and Robert V. Thomann


    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

    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.


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ーU) - NH3-nH2ー(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 ーC results in a nearly tenfold increase in the concentration of NHs;
a temperature increase of 5 degrees between 0-30 ー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

pH Value
*[condensed from Thurston, at at. (1974)]

    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

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$Q values ranging  from 2.4-3.8  mg/liter  NH3
for channel catfish (IctaJtiuJwA punctatuA), and  LCァo values ranging from
1.9-3.4 have been observed for fathead minnows  (pAjme.ph.alu piome」o6)
(Thurston, unpublished data).  A 17-hour LCso of 1.3 mg/liter NH3  was re-
ported for Gambusia (Gambu-6-ca a.^i.nu>) (Hemens, 1966), and a 24-hour
of 2.9 was reported for channel catfish (Robinette, 1976).   Lower  LC5Q
values between 0.35-0.50 mg/liter NH3 have been reported  for 5-  to 7-day
tests on bream (Afc-towu^ biama), roach (Rotc」u4  fwutituA),  perch (PeAca
j$」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


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  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


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.

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.


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.


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.

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-

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.


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:

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

                           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-

    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):

    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-

    4.  To use in the experiment various organisms, since  different
        organisms respond to the action of substances in different

    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).


              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.
the chronic lethal
lysis of the dependence of
of toxic substances (Jones
estimation of the duration
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
10 days
1 month
3 months
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).

(Boundary concentrations are estimated on the basis of regression equation)
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
Days of
18-21 21-30
0.09 0.02

    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.


    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-

Type of Influence
  Character of Influence
On properties
of pollutants.
On the time and
condition of the
contact or orga-
nisms with pollu-
On sensitivity of
organisms to the
action of toxi-
COcontent in water.
                    Content of mineral
                    substances in water.
                    Content of organic

                    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-

                    Weakening of organisms
                    due to starvation, par-
                    asites, infections, etc.

                    Stress conditions of
Rate of oxidation of pollu-

Coagulation of some substances,
neutralisation of acids and

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.	

Group of Organisms



Taxonomic 01 i go-
Group toxobity
Salmon All species
Whitefish All species
Perch Zander

Sturgeon All species

Sheatf ish

Crustaceans Gammarids
mysids, coro-
phiids, cray-
Aquatic Ephemerop-
Insects terans



Caspian zan-
der, perch,

Bream, white
bream, roach



Si das, pre-
datory clado-
All species
All species

All species
Caddis Flies


Carp, cru-
cian carp,
tench big-
head, amur


All species

All species
Bdel lids
Mobile forms

All species

All species
Beetles, Bugs

Mobile forms

All species




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.

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.

                                 SECTION 9

                            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,

    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


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.


    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 ーC.


Species and Concentration
Life Stage (ng/1 )
Brook trout
Eggs 39-502
Fathead minnows
Experiment 1 94-727
Experiment 2 13-1/3
Channel catfish
Adults 49-630
Age At
Water Initiation
Temperature of Exposure
(ーC) (Days)




eyed eggs


2.5 years
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
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

                        IN FISH EXPOSED TO TOXAPHENE
                                                            P Value of
                  Coefficient of Determination (r2)         Correlation
Species Hydroxyproline
Brook trout
Fathead minnow
Channel catfish
Coefficient (r)

Species, Life Stages
and Exposure Period
Brook Trout
Fathead Minnows
Channel Catfish
Exposure Concentration (ng/1) of
Toxaphene, and Statistical Significance
(1) or Nonsignificance (0) of Effects
on Growth (Left Column) and Hydroxy-
proline Concentration (Right)
0 -
- 0
- 0
0 1
1 1
1 1

0 0

0 0

0 0

- 0

0 -
0 0
0 -
0 0

0 -
- 1
- 1 -
0 1
1 1
1 1

0 0

0 0

0 0

- 1

0 -
0 1
0 -
0 1

0 -
- 1
- 1
1 1
1 1
1 1

0 1

1 0

0 0

- 1

0 -
0 1
0 -
0 1

0 -
- 1
- 1
1 1

1 1

1 1

0 0

- 1

0 -
0 1
1 -
1 1

0 -
- 1
- 1

1 1

1 1

0 0

- 1

0 -
0 1
1 -
1 1
Statistical  significance, P=0.05.
 Not determined.
CA11 fish had died.

growth rates of the surviving catfish fry may not have been fully representa-
tive of the original populations.


    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;

        Liver <^ Cl
r+ c*
-t <
n> n>
c n>
-i. 3
-S N

< tti
                             Hydroxylating enzymes
                                       vitamin C
                                                                   Storage, further


      (Prolineand   HVdroxylating enzymes    Hydroxyproline

Bーne)    lysine    	       anH
                                       vitamin C




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.

    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.

    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).

Axelrod, J., S. Undenfriend, and B.B. Brodie.  1954.  Ascorbic acid in
    aromatic hydroxylation.  III.  Effect of ascorbic acid on hydroxyla-
    tion of acetanilide, aniline and antipyrine -in v-tvo.  J. Pharmacol.
    Exp. Therapeut.  3: 176-181.

Barnes, M.J.  1969.  Ascorbic acid and the biosynthesis of collagen and
    elastin.  In J.C. Somogyi and E. Kodicek (ed), Nutritional Aspects of
    the Development of Bone and Connective Tissue, S. Karger A6, Basel,
    Switzerland,  p. 86-98.

Barnes, M.J., B.J. Constable, L.F. Morton, and E. Kodicek.  1970.  Studies
    in v-ivo on the biosynthesis of collagen and elastin in ascorbic acid-
    deficient pigs.  Biochem. J.  119: 575-583.

Benoit, D.A., and F.A. Puglisi.  1973.  A simplified flow-splitting chamber
    and siphon for proportional diluters.  Water Res.  7: 1915-1916.

Drummond, R.A., and W.F. Dawson.  1970.  An inexpensive method for simu-
    lating diel patterns of lighting in the laboratory.  Trans. Am. Fish.
    Soc.  99: 434-435.

Grant, B.F., and R.A. Schoettger.  1972.  The impact of organochlorine con-
    taminants on physiologic functions in fish.  Proc. Tech. Sessions 18th
    Annu. Meeting Institute of Environmental Science, New York.  p.

Green, H.B., B. Goldberg, M. Schwartz, and D.D. Brown.  1968.  The synthe-
    sis of collagen during development of Xenoptt* 」aeu^s.  Developmental
    Biol.  18: 391-400.

Harrington, W.F., and P.M. Hippel.  1961.  The structure of collagen and
    gelatin.  In C.B. Afinsen, jr.s, M.L. Ansbn, K. Bailey, and J.T. Edsall
    (ed), Advances in Protein Chemistry - Vol. 16, Academic Press, Inc.,
    New York.  p. 1-138.

Mayer, F.L., and P.M. Mehrle.  1974.  Fathead minnow growth and reproduc-
    tion as affected by 2,4-D DMA.  Proc. Midwest Fish and Wildlife Conf.
    36: 64-65.

Mayer, F.L., and P.M. Mehrle.  1976.  Vitamin C distribution in channel
    catfish as affected by toxaphene.  Toxicol. Appl. Pharmacol.  37:
    168-169.  (Abstr.).

Mayer, F.L., P.M. Mehrle, and W.P. Dwyer.  1975.  Toxaphene effects on re-
    production, growth, and mortality of brook trout.  Ecological Research
    Series EPA-600/3-75-013, U.S. Environmental Protection Agency, Wash.,
    D.C.  51 p.

Mayer, F.L., P.M. Mehrle, and W.P. Dwyer.  1977.  Growth, reproduction, and
    mortality of fathead minnow and channel catfish as affected by toxa-
    phene.  Ecological Research Series, U.S. Environmental Protection
    Agency, Wash., D.C. (in press).

McAllister, W.A., W.L. mauck, and F.L. Mayer.  1972.  A simplified device
    for metering chemicals in intermittent-flow bioassays.  Trans. Am.
    Fish. Soc.  101: 555-557.

Mehrle, P.M., and F.L. Mayer.  1975a.  Toxaphene effects on growth and
    bone composition of fathead minnows, p^nep/io」e^ piome^ai-  J. Fish.
    Res. Board Can.  32: 593-598.

Mehrle, P.M., and F.L. Mayer.  1975b.  Toxaphene effects on growth and
    development of brook trout (SalveLinuA ^ontinaLu,).  J. Fish. Res.
    Board Can.  32: 609-613.

Mehrle, P.M., and F.L. Mayer.  1976.  Bone development and growth of fish
    as affected by toxaphene.  In I.H. Suffet (ed), Fate of Pollutants in
    Air and Water Environments, Wiley Interscience Publishers, New York
    (in press).

Mount, D.I., and W.A. Brungs.  1967.  A simplified dosing apparatus for
    fish toxicology studies.  Water Res.  1: 21-29.

Mussini, E., J.J. Mutton, and S. Undenfriend.  1967.  Collagen proline
    hydroxylase in wound healing, granuloma formation, scurvy, and growth,
    Science.  157: 927-929.
National Academy of Sciences.  1973.  Nutrient requirements of
    animals.  No. 11.  Nutrient requirements of trout, salmon,
    Nat. Acad. Sci. Wash., D.C.  57 p.
                and  catfish.
Nusagens, B., A. Chantraine, and C.M. Lapiere.
    matrix of bone.  Clin. Ophth. Rel.  Res.  8!
 1972.   The  protein  in  the
;  252-274.
Peterkofsky, B.  1972.  The effect of ascorbic acid on collagen polypeptide
    and proline hydroxylation during the growth of cultured fibroblasts.
    Biochem. Biophys. Res. Commun.  152: 318-321.

Piez, K.A., and R.C. Likins.  1958.  The nature of collagen.  II.  Verte-
    brate collagens.  In Calcification of Biological Systems, Assoc.
    Advance. Sci. Publ.  No. 64, Wash., D.C.  420 p.

Rollins, J.W., and R.A.  Flickner.  1972.  Collagen synthesis in Xe.nopu4
    oocytes after injection of nuclear RNA of frog embryos.  Science.
    178: 1204-1205.

Street, J.C., R.C. Baker, D.J. Wagstaff, and P.M. Urry.  1971.  Pesticide
    interactions in vertebrates:  Effects of nutritional and physiological
    variables.  Proc. 2nd. IUPAC Int. Congr. Pesticide Chem., Tel Aviv,
    Israel,  p. 281-302.

Wagstaff, D.J., and J.C. Street.  1971.  Ascorbic acid deficiency and in-
    duction of hepatic microsomal hydroxylative enzymes by organochlorine
    pesticides.  Toxicol. Appl. Pharmacol.  19: 10-19.

Whitehead, R.G., and D.G. Coward.  1969.  Collagen and hydroxyproline
    metabolism in malnourished children and rats.  In J.C. Somogyi and
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                                 SECTION  10


         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-

    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.,

    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).

                            5-DAY EXPERIMENT)



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

    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-


of Ions, mg/1

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.

Days of
(Diurone 0.2 mg/1 )
02, Percent of
mg/1 hr-1
02, Percent of
 1 hour after
   addition      28
































     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-


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).

 Species of
Fish, Tested
 DDT Content
   Toxic Effect
   On Image of
  Intestinal      4.0

  Liver           4.0
                  Instant death of       ++++
                  all  the insects.

                  Clear symptoms of       +++
                  intoxication.  Agony.
                  Death in 20-40 min.
30-40 percent are
dead in 1  hour.
  Muscle Tissue   4.0

  Muscle Tissue   2.0

                  Death in 1  hour.
                  All  are living for
                  6-8  hours.

    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-

                      OF AQUATIC ORGANISMS AT 30 ーC
Lethal Concentration, mg/1
Test Organisms
Vaphvua. magna
(females with eggs)
Vaphnia magna
Aie」」a4 aqucuttcuA
Aie」」o6 aqao^ccoa
at 18 ーC
at 30 ーC

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.


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.

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.

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.

                                 SECTION 11

                             ON  FRESHWATER FISH

                            Lloyd L.  Smith,  Or.

     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


to determine "no-effect"  levels of hydrogen sulfide based on  long-term
tests, and to relate them to acute toxicity levels.


    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 ーC to 26 ー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 ーC with fathead minnow sub-adults, to 0.007 mg/1 H2S at 24 ー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$Q concentration increased from 0.021 mg/1 at 24  ーC to 0.515
mg/1 at 6.1 ーC (Table 2).  This twenty-five fold change in tolerance
occurred principally between 10 and 6.1 ーC.  Goldfish tested  at tempera-
tures between 14.1 and 26 ーC showed 96-hr LC$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

	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.





No. of
Stage Tests
Sac fry
Swim-up fry
Sac fry
Swim-up fry
96-hr LCso
Mean (mg/1)
_ ォ

Mean (mg/1)






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 ーC with different species.  The "no-effect" concentration varied from a
minimum concentration of <0.001 mg/1 ^S at  24 ーC in bluegills,  to 0.010
mg/1  H2S at 18.6 ー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




(5 g)






 Two tests.
^Reproduction inhibition at 0.002 mg/liter.
 R, reproduction; G, growth;  S,  survival.

 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


  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

                                         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
__                                  _

         SULFIDE AT 23  ーC, pH 7.7, Q2 6.4-7.3 MG/LITER  - EXPRESSED
                        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
Total % Increment
Survival %
No. Eggs/Female
No. Spawnings
No. Females






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 ー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.


    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


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 ーC.  In bluegills, the level is significantly
lower, and this pattern may be followed in other species not tested.

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.

                                 SECTION  12


                                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


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 TIME, days
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.

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.

                        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;

 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.

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.

    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


 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

Exposure 48 hrs.
Maximum Tolerated
Toxicants Concentrations LC5Q
No avoidance

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.


Abramson, H.A. and L.T. Evans.  1954.  Lysergic acid diethylamide  (LSD 25)
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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-
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Alekseev, V.A., and B.A. Flerov.  1972.  Effect of phenol on photoreac-
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Alekseev, V.A., and B.A. Flerov.  1972.  Reaction of avoidance of  toxic
    solutions of phenol by some aquatic insects and arachnids.  Inform.
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Alekseev, V.A., and B.A. Flerov.  1972.  Some peculiarities of behavior
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Anderson, J.M., and M.R. Peterson.  1969.  DDT:  Sublethal effects on
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Anderson, J.M., and H.B. Prins.  1970.  Effect of sublethal DDT on simple
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Braginski, L.P.,  I.L.  Burthaya,  and  E.P.  Shcherban.   1972.   Some peculiar-
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Davy, F.B., H. Kleerekoper,  and  P. Gensler.   1972.  Effects  of  exposure to
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European Inland Fisheries Advisory Commission.   1965.  Water quality cri-
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Evans, L.T.,  L.H. Geronimus, C. Kotnetsky, and H.A.  Abramson.   1956.
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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
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Flerov, B.A.  1973.  Experimental investigations of  phenol  intoxication
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Foster, N.R., A. Scheier, and J. Cairns.  1966.  Effects of ABS on feeding
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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-
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Hansen, D.J., S.C. Schimmel, and J.M. Kelther.   1973.  Avoidance of pesti-
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Hansen, D.J., S.C. Schimmel, and E. Matthews.  1974.  Avoidance of Aroclorョ
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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
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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.

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.

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.


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
    patterns to monitor zinc in water.  Water Res., V. 6, N.. 1.

                                SECTION 13


                         Albert B. Dickas, Ph.D.



    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.

    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

*That particular amphibol
Mining Company tailings
**That particular amphibol
Hollow Curved Fibers:
00^250 A; IDR50 A; of-
ten occur in bundles
and change shape in
lung liquids.
do not oc-
cur in bun-
dles and do
not split or
change shape
in lung liq-
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

    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.


    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 ーC (2400 ー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 of asbestos containing waters falls into two principle

    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

 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

     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).

    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


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  ー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


    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 ^
Chemical Oxygen Demand
> 42
> 50
> 15
> 3
> 1.5
> 19
> 0.04
(Mg 0/1 )> 16
^ 0.5 ppm
^ 40 ppm
'\, 46 ppm
表- 1 4 ppm
表> 2.7 ppm
^ 1.3 ppm
夫 4.0 ppm
夫 0.02 ppm
封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

    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.

    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.

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.


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.

                                SECTION 14


                               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.

     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

     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.

    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.

     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-

     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


                                  IZHORA RIVER STATIONS
                                 A MOSKVA RIVER STATIONS
                                 O PREGOLYA RIVER STATIONS
i    i
i    i
                                4        6
                               BIOTIC INDEX
Figure  1.   Correlation  between Woodiwiss' biotic index and  BODC

   O 30


   o 10
                     MOSKVA RIVER
                        STA TIONS
  I      I     I     I      I     I      I
Figure 2.  Correlation between Woodiwiss' biotic index  and bichromate
                      oxygen consumption.

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:
       ー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

    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.


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.

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.

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.

                                 SECTION  15

                      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-

    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.

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-

    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

                         POLLUTED PARTS OF RIVERS

                                        Chlorophyll  in Seston, %          ~~
  River                         Clean  Part                  Pol luted  Part
Western Dvina


O I-

> H 0.2
H- 2
< O
-I o

"    0.1





                   2345        6

                     SVISLOCH RIVER STATION
       Figure 1.  Distribution of seston in  the River Svisloch
                        in August of 1973.

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

O I-

> K 0.2
h- Z
< O
-1 O
                     234       5         6
                       SVISLOCH RIVER STATION
Figure  2.  Relative  chlorophyll  content  in  seston of  the River  Svisloch
                             in June of 1973.

    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

    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-

      o <=
      2 g 0.04
      LU O)
      CD o
      ^~ 01

      Ho 0.02
                      Lake Naroch
Lake Myastro    Lake Batorin
 (eutrophic)   (highly eutrophic)
    Figure  3.   Specific Oxygen Consumption (SOC) by seston  in  three
                          representative lakes.

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.
3 10
^ 5'

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 サ\j \.^i
_ _ en
0.8 E
nォ Q
123456 7 8 9 10 11 12
   Figure 4.  Relationship of photosynthesis  ($)  and  distruction  (D)  in
                    the Neman River in August of  1975.


    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.
           12   345    6
 7     8    9   10  11   12

        Figure 5.  The ratio of */D and $/R on the  Pripyat  River.


    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.

                                    TECHNICAL REPORT DATA
                             (Please read Instructions on the reverse before completing)
!僊NDSUBTITLE 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.
                                                            5. REPORT DATE
                                                               August 1978 issuing date
                                                            6. PERFORMING ORGANIZATION CODE
   Environmental  Protection Agency  -  USA
   Soviet  Academy of Sciences - USSR
                                                           8. PERFORMING ORGANIZATION REPORT NO.
   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.

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                                                        11. CONTRACT/GRANT NO.
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                                                        14. SPONSORING AGENCY CODE
   Prepared in cooperation with the Institute  for  the Biology of Inland Waters,
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        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.
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