13040EYX 11/69
    Agricultural Practices

    and Water  Quality

          NOVEMBER 1969

The Water Pollution Control Research Reports describe
the results and progress in  the  control and  abatement
of pollution of our  Nation's Waters.  They provide a
central source of information on the  research, develop-
ment and demonstration activities of  the  Federal Water
Pollution Control Administration,  Department  of the
Interior,  through inhouse research and grants and
contracts with Federal,  State,  and local  agencies,
research institutions,  and industrial organizations.

Water Pollution  Control Research Reports will be
distributed to requesters as  supplies  permit.   Re-
quests should be sent  to  the Publications Office,
Dept.  of the  Interior, Federal  Water  Pollution
Control Administration,  Washington,  D.  C.  20242.

          WATER  QUALITY
      Proceedings of a conference concerning

      The Role of Agriculture in Clean Water

    held at Iowa State University, November 1969



            Iowa State University
               Ames, Iowa

            Grant No. 13040 EYX

              November 1970

          Ted L. Willrich and George E. Smith, editors
Laws, public interest, and political motivation all point to clean
water as a major national concern. When sources of water pollution
are mentioned, agriculture is often listed as a major contributor.
Existing knowledge does indicate that agricultural operations can
contribute to the deterioration of water quality through the release
of sediments, pesticides, animal manures, fertilizers, and other
sources of inorganic and organic matter into  the water.

Streams,  lakes , and groundwater that are now polluted must be
cleansed. Future pollution must be prevented. However, prevention
cannot be attained without adequate knov/ledge concerning the causes
and sources of pollution. The opinions of the uninformed that in turn
misinform the public can only retard effective progress to assure an
adequate  supply of clean water.

This volume, an outgrowth of a 1969 conference held at Iowa  State
University, should prove both informative and useful.  The papers by
knowledgeable scientists from many disciplines present and evaluate
the existing body of facts on agriculture's contributions to polluted
water and reveal alternative solutions to providing clean water.

This report was submitted in fulfillment of Grant  No. 13040 EYX
between the Federal Water Pollution Control Administration and
Iowa State University.

MINORU  AMEMIYA,  Associate  Pro-
fessor   and   Extension  Agronomist,
Department   of   Agronomy,   Iowa
State  University,  Ames,  Iowa.

D. E.  ARMSTRONG, Assistant  Profes-
sor of  Water Chemistry,   University
of Wisconsin, Madison,  Wisconsin.

E.  R.  BAUMANN, Professor,  Depart-
ment of Civil  Engineering,  Iowa State
University, Ames,  Iowa.

HAROLD  BERNARD,   Chief, Agricul-
tural  and  Marine Pollution Control,
Office of Research and  Development,
FWPCA, Washington, D.C.

C. A.  BLACK, Professor, Department
of Agronomy, Iowa State  University,
Ames,  Iowa.

C.  S.  BR1TT,  Assistant  to  Director,
Soil and Water Conservation Research
Division, ARS, USDA, Beltsville, Mary-

G.  M.  BROWNING,  Regional  Direc-
tor,  North  Central  Agricultural  Ex-
periment   Station  Directors,   Iowa
State  University,  Ames,  Iowa.

R.  S.  CAMPBELL, Professor  of  Zo-
ology,  University  of Missouri,  Colum-
bia, Missouri.
R. J. DEMINT, Research Chemist, Crops
Research  Division,  ARS, USDA, Den-
ver, Colorado.

S.  L. DIESCH, Associate Professor,  De-
partment  of  Veterinary  Microbiology
and Public Health,  University of Min-
nesota, St. Paul,  Minnesota.

R.   H.  DOWDY,  Research  Soil Scien-
tist, SWCRD,  ARS, USDA, Morris, Min-

W.  E.  FENSTiR, Assistant  Professor
and Extension Specialist in Soils,  De-
partment  of Soil Science,  University
of  Minnesota,  St.  Paul, A^innesota.

P.   A.  FRANK,  Plant  Physiologist,
Crops  Research  Division, ARS, USDA,
Denver, Colorado.

L.   R.  FREDERICK,  Professor, Depart-
ment of Agronomy,  Iowa State  Uni-
versity, Ames, Iowa.

M. C.  GOLDBERG,  Research Hydrolo-
gist,  U.S.  Geological  Survey,  USDI,
Denver, Colorado.

L.   D.  HANSON, Associate  Professor
and Extension Specialist in  Soils, De-
partment  of  Soil Science,  University
of  Minnesota,  St.   Paul,  Minnesota.
P. A.  DAHM,  Professor, Department   T. E.  HAZEN,  Professor,  Department
of  Zoology  and  Entomology, Iowa   of  Agricultural  Engineering,  Iowa
State University, Ames,  Iowa.          State  University, Ames,  Iowa.

 H.  G.  HE1NEMANN,  Director,  North   W. C. MOLDENHAUER,  Research Soil
 Centra!  Watershed  Research  Center,   Scientist,  SWCRD, ARS,  USDA,  Ames,
 SWCRD, ARS,  USDA,  Columbia,  Mis-   Iowa.
 N.  W. HINES,  Professor of Law,  Col-
 lege  of  Law,  University  of  Iowa,
 Iowa City, Iowa.
 J. A.  MOORE,  Instructor, Department
 of  Agricultural  Engineering,  Univer-
 sity of  Minnesota, St.   Paul,  Minne-
 R. F.  HOLT,  Director,  North Central   H. P. NICHOLSON, Chief, Agricultural
 Soil  Conservation   Research Center,   and  Industrial  Waste  Control   Pro-
 SWCRD,   ARS,  USDA,   Morris,  Min-   grams,  Southeast  Water Laboratory,
 nesota.                                FWPCA,  USDI,  Athens,  Georgia.

 H. P.  JOHNSON,  Professor, Depart-   R. A OLSON, Professor, Department
 ment  of  Agricultural  Engineering,   Of   Agronomy,   University   of   Ne-
 lowa State University,  Ames,  Iowa.   braska, Lincoln, Nebraska.
 SHELDON KELMEN,  Assistant  Profes-
 sor,  Department  of Civil  Engineer-
 ing,  Iowa  State University,  Ames,

 H. E. LE GRAND, Research  Hydrolo-
 gist.  Water Resources  Division, U.S.
 Geological   Survey,   USDI,  Raleigh,
 North Carolina.

 R.  I.  LIPPER,  Professor,  Department
 of  Agricultural  Engineering,  Kansas
 State University, Manhattan,  Kansas.

 T.   M.   MC CALLA,   Microbiologist,
 USDA,  Lincoln, Nebraska.

 R.  E. MC KINNEY, Professor,  Depart-
 ment of Civil  Engineering, University
 of  Kansas,  Lawrence, Kansas.

W. P. MARTIN,  Professor and  Head,
 Department  of Soil  Science,  Univer-
sity of Minnesota,  St.  Paul,  Minne-
 G.  L. PALMER, Instructor,  Department
 of  Agronomy, Iowa  State  University,
 Ames, Iowa.

 J. T.  PESEK, Professor and Head, De-
 partment of  Agronomy,  Iowa  State
 University,  Ames, Iowa.

 D.  C.  PETERS, Professor,  Department
 of  Zoology and Entomology,  Iowa
 State University, Ames,  Iowa.

 H.  B. PETTY, Professor and  Extension
 Entomologist,  University  of  Illinois,
 Urbana,  Illinois.

J. M. RADEMACHER,  Regional Direc-
tor,  Missouri  Basin  Region,  FWPCA,
USDI, Kansas City, Missouri.

G.  A. ROLICH,  Director,  Water Re-
sources  Center, and  Professor,  Sani-
tary Engineering, University of Wis-
consin, Madison,  Wisconsin.
J.  R.  MINER,  Assistant Professor, De-    F.  J.   STEVENSON,  Professor,  Soil
partment  of Agricultural  Engineering,    Chemistry,  Department of Agronomy,
Iowa  State  University, Ames, Iowa.     University of Illinois, Urbana, Illinois.

                                                        LIST OF AUTHORS /  vii
D. R. TIMMONS, Research Soil  Scien-
tist,  SWCRD, ARS, USDA, Morris, Min-

f. I. TIMMONS, Research Agronomist,
Crops  Research Division, ARS,  USDA,
Laramie, Wyoming.
J.  F.  TIMMONS,  Professor,  Depart-
ment of  Economics,  Iowa  State  Uni-
versity, Ames, Iowa.
JACOB VERDUIN,  Professor,  Depart-
ment of Botany, Southern Illinois Uni-
versity, Carbondale, Illinois.
C. H. WADLEIGH,  Director,  Soil  and
Water  Conservation  Research  Divi-
sion,  ARS,  USDA,  Beltsville,  Mary-

G. H. WAGNER, Associate Professor,
Department of  Agronomy, University
of Missouri, Columbia, Missouri.
J. R.   WH1TLEY,  Supervisor,  Water
Quality  Investigations,  Missouri  De-
partment of Conservation, Columbia,
T. L. WILLRICH,  Professor,  Department
of Agricultural  Engineering,  and  Ex-
tension   Agricultural Engineer,  Iowa
State University, Ames, Iowa.

    Foreword	xiii
    Preface	xv
    Introduction: Issues in Food Production and Clean Water -    -   .  xix
            Cecil H. Wadleigh and Clarence S. Britt

     1.  Pollution by Sediment:  Sources and the Detachment and
        Transport  Processes	3
            H. P. Johnson and IF. C. Moldcnhauer
     2.  Chemistry of Sediment in  Water	21
            R. F. Holt, R. H. Dowdy, and D.  R. Timmons
     3.  Land and Water Management for Minimizing Sediment  .   .   35
            Alinoru Amemiya
     -1.  Workshop  Session	16
            G. M. Browning, Leader; H. G. Hcinemann, Reporter

     5.  Significance of Phosphorus in Water Supplies	63
      i      Jacob  Verduin                               '
  I   6.  Behavior of Soil and Fertilizer Phosphorus in Relation  to
  I      Water Pollution	72
            C. A.  Black
     1.  Sources of Nitrogen  in Water Supplies	94
            Mai~'in C. Goldberg
     8.  Chemistry of  Nitrogen in Soils	125
            F. J. Stevenson and G. H. Wagner                        ,.,_
 •   ; 9.  Fertilizer Management  for Pollution Control	H2
            (F. P. Martin, IV. E. Fenster, and L. D.  Hanson
    10.  Workshop  Session	159
            J. T. Pesek, Leader; R. A. Olson,  Reporter

    11.  Chemistry and Metabolism of Insecticides	167
            Paul A. Da/im
    12.  The Pesticide Burden in Water and Its Significance .   .   .183
            H. Page  Xicholson



     13. Herbicide Residues in  Agricultural Water from Control of
        Aquatic and Bank Weeds	194
             F. L. Timmons, P. A. Frank, and R. J. Demint
     14. Pesticides and Pest Management for  Maximum Production
        and Minimum Pollution	209
             Don  C.  Peters
     15. Workshop Session	224
             Don C. Peters, Leader; H. B. Petty, Reporter

     16. Livestock Operations and Field-Spread  Manure as
        Sources of Pollutants	231
             J. R. Miner and T. L. Willrich
     17. Manure Decomposition and Fate  of  Breakdown
        Products in Soil	241
             T. M. McCalla, L. R. Frederick,  and G. L. Palmer
     18. Manure Transformations and Fate of Decomposition
        Products in  Water	256
             Ross  E.  McKinney
     19. Disease Transmission of Water-Borne Organisms of
        Animal Origin	265
             Stanley  L. Diesch
    20. Animal Waste Management to Minimize Pollution   .   .    .  286
            /. A. Moore
    21. Workshop Session	298
             T. E. Hazcn, Leader; R. I. Lipper, Reporter

    22. Movement of Agricultural Pollutants with Groundwater  .    .  303
            Harry E.  LeGrand
    23. Effects of Agricultural Pollution on Eutrophication   .   .    .  314
            D. E. Armstrong and G. A. Rohlich
    24. Effects of Agricultural Pollutants on  Recreational  Uses
        of Surface Waters	331
            Robert S. Campbell and James R.  Whitley
    25. Effects of Surface Runoff on the Feasibility  of Municipal
        Advanced Waste  Treatment	344
            E. Robert Baumann and Sheldon Kelman

    26.  Legal Aspects	365
            N. William Hines
    27.  Economic  Aspects	377
            John  F.  Timmons

                                             TABLE OF CONTENTS / xi

28.  Alliance  for Action	390
        John  M. Radcmacher
29.  Accomplishments  and  Goals	397
        Harold  Bernard
Index	409

      I HE Water Resources  Research Act of  1964 (Public Law 88-
379, as amended by 89-404) provided for the investigations of water
problems  through organizations at  the  land-grant universities.  In
the midcontinent area, which produces the major portion of the na-
tion's grain and meat, a substantial portion of the research con-
ducted  by  these  organizations  has been concerned  with pollutants
that could originate from farmland.
    At a regional meeting of the organization directors and research
workers from state universities  in April 1968, "Pollution of Water by
Agriculture" was the subject for discussion.  It proved to be a topic
of intense and widespread interest. It was  apparent  that the subject
was so broad and complex that well-trained specialists in specific  re-
search fields were not communicating with their research associates
in other departments.
    This  group agreed that there would be an increasing interest
in water quality  in the Midwest, particularly in  those  areas where
agricultural production is a major portion of the economy, and where
crop and  livestock enterprises  might prove to  be  important and
growing sources of pollution. It was decided that there was need for
an  exchange of ideas and an understanding of basic chemical and
biological  processes by  those most knowledgeable in specific fields
relating to agriculture as  a  source of water pollutants.  It was
further agreed that only fundamentals and established research facts
(not opinions) should be  considered and presented  at  a level that
could be understood by representatives from  other disciplines.
    A committee consisting of Professors  Don Kirkham, Robert L.
Smith,  and George E. Smith was appointed to arrange a regional
conference on "Agriculture  and Clean Water."  Subsequently a con-
ference was held on  the  Iowa State University campus November
18-20,  1969.  Dr. T.  L.  Willrich of Iowa  State University and Dr.
George E.  Smith  of the University of Missouri served as cochairmen
of the conference.  Professor  R. L.  Smith (University of  Kansas),
Doctors Don Kirkham, Lee  Kolmer,  E.  R.  Duncan,  H. P.  Johnson,
E. R.  Baumann,  J.  R. Miner,  and -D. C. Peters  (all of Iowa State
University) assisted with the initial  planning.  In addition to spon-
sorship  by Mid-Continent State Research  Organizations and Iowa
State  University, the  Federal  Water Pollution Control Administra-
tion also cooperated.  The conference was also funded in part  by
Iowa Community Services under Title I of the Higher Education Act
of 1965.


     The participants in the conference were some of the most out-
 standing research workers in their respective fields.  The conference
 was attended by about 250 individuals from 32 states. Many  disci-
 plines,  including  engineers; agricultural, biological and  social sci-
 entists;  geologists; hydrologists;  and specialists (including  legisla-
 tive  representatives)  from other fields, attended.
     The planning committee  was pleased with  the interest, the
 scientific soundness of the presentations, and the participation. The
 conference  accomplished  the  original objectives  of the water re-
 sources  research directors.  The information published in this pro-
 ceedings is probably the most factual material in one volume on the
 chemical and biological reactions in soils and on crop and livestock
 production as they may be potential contributors to  the degradation
 of water quality.   This proceedings  provides  valuable information
 to those persons genuinely interested in the relation of modern agri-
 cultural technology to the water environment.

 DON KIRKHAM,  Director
  Water Resources Research Institute
  Iowa  State University
  Ames,  Iowa

 G. E. SMITH, Director
  Water  Resources Research Center
  University of Missouri
  Columbia, Missouri

R. L. SMITH, Chairman
  Department of  Civil Engineering
  University of Kansas
  Lawrence, Kansas

       • NVIRONMENTAL pollution is a major concern to many people.
When sources of water  pollution are enumerated, agriculture  is,
with increasing frequency,  listed  as a  major contributor.
     Except for chemical  pesticides, many materials now designated
as water  pollutants,  such as sediments, nitrates, phosphates, and
organic materials, have entered streams  and lakes since the first
sod was plowed, and even prior to that time. However, the concen-
tration of these pollutants in water has generally  increased  with
time. A portion is from agricultural  lands.  The remainder is from
nonagricultural operations.
     Movement of pollutants into water is controllable if it results
from man's activities. However, water quality degradation by natural
causes also occurs, and this may  not be controllable.
     As the nation makes an effort  to correct abuses  to its water
resources,  there is a  need to determine the causes of water quality
degradation and to quantify pollution contributions from the many
     Until such time as adequate facts are made  available  through
research  to delineate causes and sources,  conflicting  opinions will
continue  to flourish  and programs to  control  and  abate  pollution
will  be less effective and  efficient in the use of limited  resources.
     Existing  knowledge  indicates that agricultural operations can
contribute to water  quality  deterioration through  the release  of
several materials into water: sediments, pesticides, animal manures.
fertilizers, and other sources of inorganic and organic matter.
     Sediment  from land erosion can  be a pollutional material  in
surface water.  Although  soil loss from cultivated land is the major
source of  sediment in streams and reservoirs in most areas, highway
construction, rural roads,  stream bank erosion, gully erosion, housing
developments, strip mines, and logging  operations are also important
contributing sources to production of sediments.   Sediment reduces
the storage capacity of reservoirs  and lowers their value for recrea-
tional uses. Sediment, depending on origin, contains different inor-
ganic minerals and organic compounds. Both supply plant nutrients.
These  nutrients can stimulate  the growth of undesirable  aquatic
plants that on  decomposition can cause eutrophication and  increase
costs when surface water  is treated for domestic use or for industry.
Decomposition of the organic material can utilize dissolved oxygen
in water.   Residues of slowly  degradable  pesticides used in agricul-

 xvi / PREFACE

 tural production are adsorbed on sediments. These may serve as a
 reservoir to be taken up by aquatic plants and eventually  enter the
 food chain.
      Production economics and a shortage of farm labor have caused
 some livestock and poultry operations to develop  as an agribusiness
 with large numbers of animals concentrated on small land tracts.
 Concentrated  animal wastes have  created  problems  of waste  dis-
 posal and water pollution  through runoff and  leaching.  Animal
 manures for soil fertility maintenance  are no longer  considered as
 valuable as they once were.  Nutrients  required in crop production
 can frequently be applied in chemical fertilizers at a lower  cost than
 the cost of hauling and  spreading animal wastes. However, land
 application of animal manures continues to be the least-cost disposal
 alternative in most situations and the preferred method to  reuse the
 plant nutrients they contain.  Lot runoff has less nutritive value than
 concentrated manure, but the  plant  nutrients and organic matter
 that it contains can pollute a receiving body of water.  Consequently,
 many states have or are in the  process of regulating feedlot runoff
 discharges to prevent water quality degradation.
     The use of chemical fertilizers, with  other practices, has pro-
 vided ample crops for domestic consumption and  export to meet the
 needs of an exploding world population.  At least  one-third of this
 nation's food production can be attributed to the  use of  chemical
 fertilizers.  Therefore, fertilizer use is  essential  to prevent mass
 starvation.  However, a portion  of the applied fertilizer may be re-
 moved from the soil by leaching or  runoff and thus enter a ground-
 water or surface-water body  along with plant nutrients from other
 sources. Inadequate data are available to separate nutrient contribu-
 tions from the many  sources:   chemical fertilizers; weathering of
 soil minerals;  mineralization of  nature's storehouse of humus; crop
 residues; animal  manures;  atmospheric contributions of nitrogen
 through rainfall,  soil adsorption, and legume  fixation;  and  many
 domestic, municipal, and industrial wastes.
     The problem of pesticides  as pollutants is complex.  The  use
 of these  compounds, combined with other management  practices, has
 permitted the production of an abundance of a wide variety of foods.
 However, some of the more effective pesticides degrade only slowly.
 Some may dissolve in water or be sorbed on sediments. The more
 resistant materials may accumulate  and enter the food chain. Given
 a  choice, few consumers would buy foods which  contain insects,
 are affected by disease, or are contaminated by rodents—no matter
 how low the price. It is doubtful if sufficient food  could be  grown,
 stored, and processed that would meet the requirements of the Food
 and Drug  Administration without  the use of pesticides.  There is,
 however, a need to develop effective new  compounds that  do  not
 persist in soil, water, or plant or animal tissue.
     Laws, public interest, and  political motivation point  to clean
 water as a major national issue in coming years. Streams, lakes, and
 groundwater that  are polluted must be cleansed.   Future  pollution
must be  prevented. However, prevention cannot be  attained without
 adequate knowledge concerning the causes and sources of pollution.

                                                      PREFACE / xvii

The opinions of the uninformed that misinform the public can only
retard effective progress to assure an adequate supply of clean water.
     The materials contained in this volume should he both informa-
tive and useful.  It was assembled by knowledgeable scientists, rep-
resenting many disciplines, to identify the role of agriculture in clean
water; more specifically,  to present and evaluate  the existing body
of facts as they identify agriculture's contributions to polluted water
and reveal alternative solutions to provide clean water.

  Extension Agricultural Engineer
  Iowa State  University
  Ames, Iowa

GEORGE E.  SMITH, Director
  Water Resources Research Center
  University of Missouri
  Columbia. Missouri


          E AMERICANS are carnivorous.
    The average person's dinner plate accounts for 238  pounds of
flesh per year. Three-fourths of this consumption is of red meat, and
nearly half of the total is beef.  Poultry now provide one-fifth of our
total meat consumed, while lamb and fish account for a rather small
percentage of the total.
    We are eating just twice as much beef per capita today  as we
were 30 years ago. In fact, if a visitor from outer space were to enter
one of our wonderful restaurants, he would  gain the impression  that
the favorite indoor sport of Americans is that  of attacking a juicy
    I am delighted to be an American!
    Our inventory of beef cattle has been increasing at about twice
the rate of our population. Since our numbers of dairy cattle have
been decreasing rather markedly during the past 20 years,  total cattle
population appears to be leveling off at about 110 million.
    A big Holstein cow will  produce 75 pounds of fecal wastes a
day, along with 20 to 30 pounds of liquid wastes.  A little  effort with
a slide rule will tell you that a 100-head dairy produces 1,800 tons of
wastes a year, exclusive of bedding. Obviously, every dairyman has
no small problem in working out a system of materials handling.
    Beef  steers out on the range may produce only 30  pounds of
fecal wastes a day, and 15 to 20 pounds of liquids. Cattle being fat-
tened in feedlots daily  produce between 35 and 60 pounds of fecal
wastes and between 18  and 25 pounds of liquids.
    The poultry industry also  faces major problems  in  waste dis-
posal. Per  capita consumption of fryers has increased 26-fold in the
last 30 years.  Furthermore, in terms of 1968 dollars, the  farm price
per pound of broilers is only one-fourth of what it was 30 years ago.
How could one have  a  more vivid picture of what improved agricul-
tural technology means to the food consumer?

     CECIL H. WADLEIGH is Director of the Soil and Water Conservation
     Research Division,  ARS, USDA. CLARENCE  S. BRITT is Assistant to
     Director of the Soil  and Water Conservation  Research Division, ARS,


     Now let us consider a few of the problems that have developed
 because of increased production of animal wastes  and the demands
 of a burgeoning  population  in  suburbia  that voices loud concern
 about the quality of its environment, with special emphasis on air and
     Many of these suburbanites have emigrated into the rural fringe
 in order  to live in the pastoral delights  of a rural  atmosphere—and
 then have vigorously complained about some of the rural atmosphere
 they have received. Many of  them have problems in water  supply
 and waste disposal.
     We  can  now recognize that our current animal waste disposal
 problems were  markedly  affected  by two developments  that took
 place in 1912. We can assume beyond all  reasonable doubt that the
 distinguished German chemist, Fritz Haber, had no idea that he was
 sowing the seeds of a massive  manure disposal problem in the United
 States when in 1912 he succeeded in synthesizing ammonia by pass-
 ing H2 and N2 over hot iron filings at high temperature and pressure.
 The seeds were not long in sprouting.  Haber's process, with the de-
 velopmental work of Karl Bosch, was a tremendous contribution  to
 Germany's armed might during World War I by making Germany in-
 dependent of Chilean nitrate (Taylor, 1953).
     This synthesis of ammonia was first  performed in the United
 States  in  1920 at  the Fixed Nitrogen Laboratory set up  by the War
 Department in  1919.  By the late  1920s, synthetic  ammonia for
 fertilizer  use was  in commercial production. Figure A shows that
 during the past 50 years, use of fertilizer nitrogen has doubled about
 every 10  years until in 1969  we used nearly 7  million tons.  This
increase in usage has been abetted by the relatively  low cost of nitro-
 gen. For  example, during this past year  many a ton of nitrogen was
 applied to fields at a cost of less than 5 cents a pound.
                           70OO -
FIG. A.  Use of  plant nu-
trients, 1920 to 1968.
                                  1930   I94O
                                           1950  I960

                                                 INTRODUCTION / xxi
                                   - 200
                                          FIG.  B.  Acreage  of  59
                                          principal  crops harvested,
                                          plus   acreages  in  fruits,
                                          tree  nuts,  and farm  gar-
                                          dens. Total  United States
                                          population, including  per-
                                          sons  in our military forces
                                          in   this   country   and
   1920  1910  I94O 1950  I960  1970  1980  IDSO 2000
     This rapid increase in the use of chemical nitrogen, along with
other purchased inputs, is closely associated with the fact that al-
though the  United States has had a sharp increase in  population,
there was a marked decline in  acreage of cropland  harvested as
shown in Figure B.  The rapid increases in use of weed killers and
farm machinery are other inputs that have contributed  to  this tre-
mendous increase in efficiency of land use to meet our population
growth.  Our corn crop, which is a primary source of feed in animal
production,  provides  an example of the  effect of enhanced  farm
technology  on land requirements in producing an  abundance of
animal feed at low cost.
     During the past 20-odd years, the corn acreage harvested has
decreased markedly, so it is only  about two-thirds of what it was in
1945.  Nevertheless, our  corn production  over  this period has  in-
creased very significantly, to double the total bushels produced in
1940.  These trends took place because of the rapid increase of aver-
age corn yield per acre  since 1940 due to the increased use of fertil-
izer  and other purchased inputs upon corn land. As a consequence
the price of corn  to the farmer in 1968 dollars has shown a very
significant general decline since 1940, which,  in turn,  has  its impli-
cations on the price of the tremendous amount of meat we eat.  The
corn crop has received a high degree  of sophisticated mechanization,
including eight-row planters that  sow seed, distribute fertilizer, and
apply pesticides all in one fell swoop.
     This rapid increase in field mechanization reminds us of another
very significant event that took place prior  to World War I that now
has a very marked bearing on our animal  waste disposal problems.
During 1912-14, Henry Ford started mass production methods, in-
cluding continuously moving assembly lines. This immediately  made


 possible the assembling of a Model-T in 93 minutes. This innovation
 in materials handling sparked the American industrial revolution.
 The value of man's labor was greatly increased. With the adoption
 of electronic controls to assembly-line methods, industry made further
 advances in efficiency of  production. With the advent of World War
 II, farm labor became in short supply.  Innovative poultrymen and
 cattle feeders  began installing labor-saving devices involving auto-
 mated  silos, conveyor belts, mixers, and all manner of schemes to
 expedite the handling of feed.  Thus one man became  able to feed
 thousands of cattle and tens of thousands  of birds.  In due course,
 feedlots came into being that carried 50,000 head, and poultry enter-
 prises developed that involved over 100,000 birds. Waste production
 was concurrently concentrated in large masses.  Unfortunately, effi-
 cient methods  of handling waste materials by no manner kept pace
 with improved  efficiency in feeding operations. In many instances the
 stuff accumulated as miniature mountains.  A number of economic
 studies indicated that the value of the manure  to the land was  so
 low that it was cheaper for the farmer to get his plant nutrients from
 the fertilizer bag than to haul manure from feedlots to the field.
     There  is indeed a vast materials handling  problem when you
 consider that total cattle  wastes alone amount to 1.4 billion  tons a
 year. Wastes from all of our domestic livestock  come to 1.6  billion
 tons; and when bedding, dead carcasses, and the offal from slaughter-
 ing are added, the total is close to 2 billion tons a year. Obviously,
 large feedlots carrying 100,000 head of cattle produce  tremendous
 masses of material.
     To consider the magnitude of the problem, we have  to recognize
 that a fattening steer will give off in the neighborhood of 110 pounds
 of nitrogen, 125 pounds of potassium, and 365 pounds of biochemical
 oxygen demand (BOD) in its excrement each year.  This means that
 a feedlot with  a stocking rate of 200 head  per  acre will deposit a
 really tremendous amount of plant nutrients and readily  oxidizable
 organic matter on each acre.
     The Monfort Feedlot  near Greeley,  Colorado,  carries 90,000
 head on 320 acres.  The feeding and waste handling procedures are
 reallv something to see.
     We know  that most of the nitrogen deposited on a  feedlot goes
 into the atmosphere as a result of denitrification processes, but we
 also know from soil  cores taken from feedlots in  Colorado  that there
 can  be  deep percolation  of nitrate-nitrogen in  comparatively high
 amounts.   We   also know  some  of these  nutrients and organic
 materials can enter into runoff water.
     As a matter of fact, it has been common practice to locate feed-
 lots on a hillside  above a waterway in order to provide good drainage
 and  a disposal  area for the runoff.  Some feedlots traverse a stream
 course with  the  assumption that  the stream will  carry  away  the
     On a smaller scale,  we find  that even  some of the best dairy
 farmsteads in the Northeast are built along stream courses so that the
 barnyard will drain into the brook. Farmers running dairy farms or
feeder operations in the North prefer  to apply manure  to the land
 during the  wintertime  when labor demands for other tasks are at a

                                                INTRODUCTION /  xxiii

minimum.  This means  that much  of  this manure may be spread
right on top of an  accumulation of snow.  It also means that if there
is rapid snowmelt in the spring, with a comparable rate of runoff,
there will be  appreciable amounts of the wastes with plant nutrients
and BOD moving  off the fields and into streams.  Studies  on Lake
Mendota near Madison, Wisconsin, indicate that  much of the pollu-
tion coming into that lake during the spring months has its source in
runoff from barnyards or fields on which manure  was applied during
the winter.
    We must recognize that runoff carrying manure can be a cause
of major fish kills.  In fact, the Federal Water Pollution Control  Ad-
ministration  reports that of 8  major fish kills in 1967, 3  of  them
were due to manure drainage (U.S. Department of the Interior, 1968).
It is also of interest that the one really serious fish kill was caused
by food products, which frequently have an effluent exceedingly high
in BOD. It is also of interest that not one of these listed was caused
by pesticides.
    Studies  in Kansas  (Smith and Miner,  1964) show what may
happen in a stream receiving drainage from a feedlot. Their studies
on the Fox River were made at a point about  a. mile below  a feedlot
during November.  Figure C shows  that in 20 hours  after a 1-inch
storm the water in the stream  1 mile below the feedlot contained 90
ppm of BOD  and just about zero ppm of oxygen. Fish cannot survive
if oxygen content  of water falls below 4 ppm. At the point of sam-
pling, the BOD dissipated rapidly and the oxygen content of the water
    It is also important to note that in this study Smith and Miner
found the pollution from fecal coliform bacteria rose to a tremendous
level 20 hours after the storm and then dissipated rather rapidly.  By
contrast, the  count for fecal streptococcus bacteria rose to  an enor-
mous  count and continued  at  that high level  of infestation for  the
duration of the sampling period (Table A).
    As with cattle, the tremendous  growth of the broiler industry
in concentrated chicken factories has resulted in the production of
large quantities of manure in local areas.
                                      ~]   FIG.  C.  Water quality pa-
                                          rameters. (From Smith and
                                          Miner, 1954.)
                 40  70  80


TABLE A.  Fox creek bacterial pollution (average counts per TOO ml sample,
          Nov. 1962).
Mean dry weather 	
After rainfall (11-27) 	
After rainfall (11-28) 	
After rainfall (11-29)
After rainfall (11-30) . .

	 542 000
23 000
	 7 900

1 600 000
1,410 000
1 600 000
1 600 000

Source: Adapted from Smith and Miner (1964).

     In the Southeast alone,  9 milh'on tons of chicken litter are pro-
duced annually.  When this  chicken litter is spread on fescue grass
in September at the rate of 16 tons per acre, it supplies 640 pounds
of nitrogen and 250 pounds of potassium per acre.  If it is applied
at higher levels, it may even kill the fescue grass and allow useless
weeds to take over.  Even a modest application of chicken manure to
fescue pastures will produce fescue hay in November that contains
0.6% nitrate-nitrogen.  This grass may not only induce nitrate poi-
soning in the cattle but may also be "grass tetany prone"—that  is,
too high in potassium and  too low in magnesium.
     High applications of chicken litter to oats in late winter will en-
courage a lush  growth that is very susceptible to lodging and danger-
ously high in nitrate-nitrogen.
     It is not unusual  to  see cattle  that have  been eating a high
nitrate forage gather in farm ponds  to keep cool.  They often pant
continuously. They  do this because the nitrate in the grass becomes
nitrite in the rumen, and this causes some of the hemoglobin in their
blood to be changed to methemoglobin, which does  not carry oxygen.
In other words, the cattle  pant and try to keep  cool because of  an
oxygen deficiency in their systems.
     We hear much about agricultural endeavor causing the eutrophi-
cation of our surface waters. Eutrophication is nutrient enrichment
enabling the growth of plant life in water. It is  nature's way of pro-
viding fish food.  As already documented, feedlots can be a good
source of plant nutrients.  We can see tremendous  eutrophication in
drainage ditches in  western  Minnesota or streams  in eastern Mary-
land. Lakes in Minnesota that are  completely  removed from any
agriculture may accrue an excess of water plants.
     We ought  to look at this movement of nutrients into water and
recognize that it is more complex than just the supplying of nitrogen
and  phosphorus that causes an excess of plant  growth. There  are
any number of  publications on fish production in farm ponds, or fish
farming, that indicate  the need for the addition  of 800  to  1,200
pounds of 8-8-4 fertilizer per acre of pond surface  per year in order
to assure  good  fish production.  This fertilizer is essential to enable
good growth of water plants that provide food for the fish.
     Studies on agricultural  and wooded watersheds at the  U.S. Hy-
drologic Field Station, Coshocton, Ohio, show the phosphorus delivery
per acre per  year is only 0.03 to 0.06 pounds. Nitrogen yield from

                                                INTRODUCTION / xxv

the watersheds is appreciably higher but still low in terms of nitrogen
needs of plants for fish food.
     If there is an excess of organic matter high in BOD going into
the water along with the nutrients, the depletion of oxygen will kill
the fish—as one can sometimes see along the banks of the Potomac
River.  In fact, there are places in the Potomac estuary, below the
Blue Plains  sewage disposal facility that dumps into the river, where
nothing seems to grow but ugly water plants. The oxygen content of
this water is so low that the fish which would normally eat some of
the plankton and other plants are eliminated by oxygen deficiency.
     Fish can  also be eliminated by pesticides.  The heavy fish kills
in the Mississippi River in 1964 were alleged to have been caused by
endrin.  The source of the  suspected  endrin was not land runoff.
There have  been no end of  fish kills attributed to pesticides  moving
into water.  Here  again, if  these  pesticides kill  the fish, ecological
balance is upset; that  is,  there is no curb  on the growth of water
plants.  When an  overproduction of water plants takes place, some
will die and rot, contributing to  further oxygen depletion,  nutrient
release, and initiation of a vicious cycle of an abundant plant growth
incurring water environment inimical  to the growth and survival of
fish and other faunal  life.  The process becomes one of forming a
muck bog out  of the lake.  Yet, we are inclined to the view that plant
nutrients in water  should not be  considered as  pollutants.  Rather,
we ought to look upon such nitrogen  and  phosphorus as potential
     Possibly  we  should  even  paraphrase the words  of  George
Clemenceau when he stated that "war was just too important a matter
to be left in the hands of generals" (Seldes, 1966), and say that water
contamination is just too important a matter to be left in the hands of
sanitary engineers.
     Consider  a few data.
     A good fish pond will produce over 1,000 pounds of fish per acre
per year.  One thousand pounds of fish contain  200  pounds of dry
matter, of which 150 pounds are  protein that contains 24 pounds of
nitrogen.  The conversion factor from plant protein to flesh protein
by foraging fish ranges from 5 to more than 20 to 1.  Thus, at least
120 pounds of plant protein nitrogen  are needed to produce these
1,000 pounds of fish. However, fish biologists often find P deficiency
in surface waters as the main limiting factor in fish production.
     We ought to ask fish and wildlife experts to prescribe  optimal
aquatic ecologies for the production of adequate food for abundant
fish not only for man's food and recreation but also for the benefit of
fish-eating  wildlife.  This will certainly require  minimal  delivery to
our surface waters of  such fish killers as putrescible  matter,  acids,
sediment, insecticides,  and other chemicals.  It may also require use
of herbicides  with  high biochemical specificity on unwanted  water
plants.  It may mean manipulation of fish population to attain proper
balance between foragers and carnivores.
     What   are  we  doing  in agriculture  to  solve  water  pollution
     First of all, we in agriculture feel strongly  that every measure


 feasible must be taken to  minimize or eliminate possible adverse
 effects  from the use of pesticides. We are now using about a billion
 pounds of these chemicals a year. Some of them are very persistent
 in the environment; some of them volatilize and become widely dis-
 persed; some of them can be very toxic to insects, plants, or wildlife
 which we want to protect. And yet we also recognize that it is manda-
 tory that the ominous threat of insect pests, diseases, and weeds  to
 our production of food and fiber must not be ignored.   We must ever
 seek  chemicals  carrying a minimum of danger and  adverse  side
 effects. We must seek improved technology in handling and applica-
 tion of these chemicals, and wherever feasible, seek methods  of
 biological control or nonchemical control.
     Toward minimizing the damages that may occur from pesticides
 and all other pollutants that may occur in runoff from the land, we
 must recognize the long-proved advantages of conservation practices
 that will curb runoff and soil delivery. Water moving across the land
 is completely indiscriminate. It will pick up and move that which is
 movable,  whether it be soil particles, manure, plant residues, pesti-
 cides, fertilizers, or other  chemicals.  Use  of  grass waterways has
 proved  very effective in minimizing the transport of any undesirable
 burdens in the runoff water.
     We need to develop water diversion structures around our farm-
 steads and feedlots so that none of it runs directly into a watercourse,
 but  rather  into  a  storage  lagoon where   oxidation  of degradable
 materials may take place.
     Under some conditions there probably ought to be secondary
 or even tertiary  lagoons to make certain that  runoff finally  entering
 into the watercourse is fairly well reclaimed (less than 20 ppm BOD).
 Lagoons have been used with good success around poultry operations
 in the  South if  they were designed  to be of adequate capacity and
 were operated without intermittent loading.
     Some hog operations use lagoons satisfactorily, yet many in the
 northern  states  are failures. They do not  control the emanation  of
 foul odors.  There is a large hog operation  near Pendleton, Oregon,
 that is of interest. The hogwash is collected in lagoons and then dis-
 tributed through a large sprinkler irrigation system that covers 140
 acres of cropland in  one rotation. In this  particular operation the
 hogwash aids in producing 10 tons of alfalfa hay per acre.  The hay
 is ground and used as hog feed. The operation is a good example  of
 recycling  of wastes—an objective that should be followed whenever
     Cattlemen and dairymen in the northern states,  where restric-
 tions have been imposed on spreading manure on frozen ground, are
using slatted floors with the collection of the excrement in enormous
vats.  The liquified manure is spread upon the land by use of either
movable sprinkler systems  or large mobile tanks.  Many dairymen
have constructed concrete receiving basins for manure  that eliminate
runoff to stream channels while enabling the easy operation of load-
ing equipment to get the manure on the land expeditiously.
    Let us go back to the Monfort Feedlot in Greeley, Colorado.
    The hundreds of thousands of tons of manure produced on this

                                               INTRODUCTION / xxvii

feedlot  are picked  up by  high-efficiency loading  equipment  and
trucked to over 10,000 acres  of land growing corn for  cattle  feed.
Chopped corn so produced is ensiled in the amount  of about 200,000
tons.  The ensilage is  then  mixed with cooked  grain by automated
equipment and fed to the cattle by specially designed trucks.
    This operation is  a very  excellent  example of the recycling of
    Finally, we must stress again that a key contribution in making
beneficial use of agricultural wastes, and minimizing any loss of these
wastes from the farm, can take place  through sound conservation
farming.  It also contributes to beauty of the countryside.
    We  must make sure that every watershed above our water im-
poundments is effectively protected  so the quality of the water in the
reservoir may be used without concern for fishing, recreation, supple-
mental irrigation,  and even for municipal water supply.  Obviously,
we  who are involved in agricultural technology still have a big job
to do.


Seldes, G. 1966. The great quotations, p. 162. New York:  Lyle Stuart.
Smith, S.  M., and Miner, J. R. 1964.  Stream pollution from feedlot
    runoff. Trans. 14th Ann. Conf.  Sanit. Eng. Bull. Engineering
    and Architecture 52. Lawrence,  Kans.:  Univ.  of Kans. Publ.
Taylor, G. V. 1953. Nitrogen production facilities in relation to pres-
    ent and future demand.  In Fertilizer technology and resources
    in the  United States,  ed. K. D. Jacob, pp. 15-61.  New York:
    Academic  Press.
U.S. Department of the Interior.  1968.  Pollution  caused fish  kills,
    1967. CWA-7.





          HILE erosion has been active over geologic time, man has
often altered the process to the detriment of his environment. Con-
sidered  by many people  more  innocuous  than sewage, suspended
solid loads delivered to streams and lakes as sediment in surface run-
off are equivalent by weight to more than 700 times the load from
sewage  (U.S. Department  of Agriculture,  1968).  Sediment  reduces
water quality and often degrades deposition areas. Sediment  pollutes
when it occupies  space in reservoirs, lakes,  and  ponds;  restricts
streams  and drainageways; reduces crop yields in  a given year;
alters aquatic life in streams; reduces the recreational and consump-
tive use value  of water through turbidity; and increases water treat-
ment costs.  Sediment  also carries  other  water pollutants  such  as
plant nutrients, chemicals, radioactive materials,  and pathogens.
    Because the sediment pollution problem is so broad, we do not
attempt to describe the entire  problem but do (1) identify  problem
areas, (2) define present understanding of  the erosion and transport
process,  and (3) indicate  research  needs. We discuss continuum
from field erosion  to streams, but our primary emphasis is  given to
agricultural  aspects of erosion  and sedimentation in the humid cen-
tral region of the United States.  Only mechanical processes  are con-
sidered;  chemical  and biological processes  are  omitted.   Detailed
coverage of  the erosion-sedimentation process is available from sev-
eral sources (Colby, 1963; Einstein, 1964;  Gottschalk,  1964; Wisch-
meier and Smith, 1965; Raudkivi, 1967).
    To express relatively the status of  understanding of the various
     H. P. JOHNSON is Professor of Agricultural Engineering, Iowa State
     University. W. C. MOLDENHAUER is Research Soil Scientist, ARS-
     SWC, USDA, and Professor, Iowa State University.
     Contribution from Agricultural Engineering Department, Iowa State
     University, Ames, and Corn Belt Branch, Soil and Water Conserva-
     tion Research Division, ARS, USDA, Ames.  Journal paper No. J-6393
     of the Iowa Agriculture and Home Economics Experiment  Station,
     Ames. Project Nos. 1266 and 1776.

 TABLE  1.1.   Analysis approaches.
   diagrams of
   system output
 Laws  of
Given little
Selected then
  screened by



Selected data;
None defined;
Defined by


Defined by

Neglects change in
  system with time;
  need representative
  data over long

Need representative
  data; accuracy of
Range of data;
  design require-
  ments of model
Accuracy of
  equations describ-
  ing processes;
  input data

Often difficult to
  relate to entire
  system; often a
  component of
 processes involved in erosion and  sediment transport,  we comment
 on approaches to problems.  As understanding of a problem improves,
 we proceed from  empiricism  to physical  "laws."  This should  not
 condemn empirical approaches; in many instances these are the only
 approaches available to planners and designers.
     As we proceed from empiricism to  laws, however, we  are better
 able to define the factors involved in  a  process (inputs)  and  can
 better  explain  the interaction of  the  factors  (system operation).
 Table  1.1 presents an attempt to describe analysis approaches.  Al-
 though all approaches are used in erosion and sedimentation studies,
 the application of  mathematical models to unsteady  state problems
 is only beginning.  Most design approaches  are based  on field ob-
 servations,  and it is ironical in this time  that most design is based
 on observation and not on Newtonian physics.

     The  ability  to predict on-site  sheet  and gully erosion and  the
transport of eroded material to a point of concern is extremely  im-
portant in planning, design, and economic  analysis. The total  on-
site sheet and rill erosion (gross erosion)  is not delivered to streams.
The amount of sediment that completes the route of travel from  the
point of erosion to a point of concern in  a watershed is termed sedi-


ment yield.  The amount of sediment that travels this route involves
factors  related to sheet and gully erosion.
     Sheet Erosion

     In the early  1930s the  U.S.  Department of Agriculture estab-
lished  ten  soil erosion  research stations.  Using  some of the  data
collected at these  and at state stations, Smith (1941), Browning et
al. (1947),  and Musgrave (1947) attempted to systematize the calcu-
lation  cf erosion losses by using the pertinent causative factors.
     In 1954 a Runoff and Soil Loss Data Center was established at
Purdue University by the Agricultural Research  Service of the U.S.
Department of Agriculture.  All available  data from soil and water
loss  experiments throughout  the  United  States were  assembled for
summarization and analysis.  A major result of  this summarization
was  the so-called Universal Soil Loss Equation (see Wischmeier and
Smith, 1965, for development and use). This  equation is A=RKLSCP,
in which A is the computed soil loss in tons per acre, R is the rainfall
factor  and is the number of erosion index units in a normal year's
rain. K is the soil credibility factor and is the erosion rate per unit of
R for a specific soil in cultivated,  continuous fallow on a 9% slope,
72.6 feet long. L  is the effect of slope length and S is the effect of
slope gradient. C  is the crop  management factor  and  is the ratio of
soil loss from a field with specified cropping and management to that
from the fallow condition on which the factor K is evaluated.  P is the
erosion control practice factor and is the ratio of soil  loss with con-
touring, strip-cropping, or terracing to that with straight-row farming
up and down  slope. All factors are dimensionless except A and K,
which  are in tons per acre, and R which is the number of El units.
     R, the rainfall  factor, is the rainfall erosion index  developed
by Wischmeier (1959) and Wischmeier and  Smith (1958).  It is the
annual summation of El/100, where E is the kinetic energy  of  a
rainstorm and  I is its maximum 30-minute  intensity.  The E and I
values  can be  obtained from recording  rain-gage charts.  Expected
Iccational values were published in 1962 in the form of an iso-erodent
map (Wischmeier, 1962).  The proved  high correlation of Ef  with
soil  erosion has made this equation usable  anywhere in  the world
where  the R and K values can be characterized.
     Values of  K have been determined for 23 major soils on which
erosion plot studies wrere carried out (Wischmeier and Smith, 1965).
Values for many other soils have been approximated by interpolation
and  extrapolation at joint ARS-SCS workshops. Recently, Wischmeier
and  Mannering (1969)  developed  an equation by using multiple re-
gression analyses  which estimates K on the basis of  soil properties
and  their interactions.  This  equation will allow  more accurate de-
termination of K than  can be done by interpolation and  extrapola-
     The slope length  and gradient  factors  (L and S) are ratios to
field slope losses from a 72.6-foot  length and 9^r  slope, respectively.
L may be expressed as (A/72.6)"1, where A is  field slope length in feet


 and m is an exponent determined from field data.  S = (0.43 + 0.30s
 -f- 0.43s2)/6.613 where s is the  slope gradient expressed in percent.
 Together they may  be expressed  as  LS = yX(0.0076 -f 0.0053s
 + 0.00076s2).
     The cropping management factor C is the  ratio of the soil loss
 from a field with  specified  cropping and management to that from
 the fallow condition on  which the factor K is evaluated.  Five crop-
 stage periods are  used that reflect the changes in plant cover and
 surface residues through the year.  Productivity level, crop residue
 management, crop  sequence, plow date,  and  length of  meadow
 periods are all considered  (Wischmeier,  1960).  The  erosion control
 practice factor P is concerned with  only contouring, strip-cropping,
 or  terracing.  Improved tillage  practices, sod-based rotations, fer-
 tility treatments, and greater quantities  of crop residue left on the
 field are included in the C factor.
     The Universal Soil Loss Equation  was  developed from many
 years of plot data assembled from many locations.  In the past 12
 years rainfall simulators have been  used to update the information
 from earlier plot studies and to field test new concepts and prac-
 tices (Meyer et al., 1965). Most of the  plots used were 72.6 feet long
 and 0.01 to  0.025 acre  in  area.  The plot sites represented major
 soil types over a large part  of the United States.  All the plots were
 on  uniform slopes.  Consequently, the more  uniform the  slope in
 the field, the more accurate were the predictions.  In  developing this
 equation for field  use, researchers recognized that data  were most
 lacking for predicting K values  and for  dealing with more complex
 field topography.  Onstad et al. (1967), Young and Mutchler (1969),
 and others have found that  erosion from a concave slope is less than
 that from a uniform  slope because sediment tends to deposit at the
 bottom.  Erosion from a convex slope is greater  than that from a
 uniform slope.  Incorporating this type of informaticn into the Uni-
 versal Soil Loss Equation can improve predictions. Wischmeier and
 Smith (1965) recommend use of the complete slope length with the
 gradient of the  lower one-third  to determine the value  of LS for
 concave or convex slopes.
     The Universal Soil Loss Equation was designed to predict field
losses on an average  annual basis.  When  it is used  to predict sedi-
ment content of streams  and losses from  watersheds, factors must be
added to account for deposition in terraced and bottomland areas ad-
joining streams and for contributions from streambanks and gullies.
It is difficult to check the equation's accuracy on  a field basis.  The
geometry of most fields does not allow  measurement of field soil loss
because of interception above the gaging point.  Hadley  and Lusby
(1967), however, found very close agreement between measured and
predicted erosion (13 vs. 15 tons per acre). In  1965 at the Treynor
Experimental Watersheds in Iowa, Piest and  Spomer (1968) found
measured values in May  and early June were greater than predicted.
After early June, predicted  values were  higher.  It is expected  that
predicted values would always be higher because sediment deposition
on alluvial  and colluvial  areas of the watershed  would remove some
sediment actually  lost from the  hillslopes. Higher measured early
losses may be due  to development of rills in the  plow-through drains


as noted by Piest and Spomer.  Once the rills reach the depth of the
plowed layer, the rill growth seems to slow considerably or stop alto-
gether where the slope gradient is low.  Because of significant inter-
actions of the management and practice factors (C and P) with storm
size and antecedent soil moisture, single-storm or short-term predic-
tions tend to be less accurate than longer-term predictions. Spraberry
and Bowie (1969) correlated total measured sediment from 12 water-
sheds ranging from 243 to 32,000 acres with computed gross erosion.
They found a coefficient of correlation of 0.97 between total measured
sediment and the  sum of erosion computed from active gullies  and
sheet errosion computed by the Universal Soil Loss Equation. The
coefficient  of correlation  was 0.95 when the  Musgrave (1947) equa-
tion was used.  They also found that computed gross erosion from
cultivated land 2% slope and above, and from active gullies, corre-
lated better with total measured sediment yield than the erosion com-
puted from the entire contributing area.

     Some studies in geomorphology are cf interest from an agricul-
tural point of view. For example, Leopold et al. (1966) estimated
slope erosion by using a system  of pins and washers.  They  also
studied deposition in an attempt  to determine a sediment budget.
They obtained an average value cf surface erosion of 0.015 feet per
year on sparse range vegetation in a semiarid area in New Mexico.
This amounts to an erosion rate of 30 tons per acre per year from
10%  slope.  They estimated from their  data that sheet erosion
is by far the largest source of sediment. Channel deposition is only
about half cf the total sediment trapped, and this is only about one-
quarter of the total sediment produced.  They point out that sediment
spread thinly over colluvial areas does not show up in their measure-
ment data. Their sediment accumulation data compare very  favor-
ably to those of  Hadley and Schumm (1961).  Both groups conclude
that sediment accumulation per unit area of basin decreases rapidly
with increasing  drainage area.  Data of Hadley and Schumm were
also collected in  a semiarid area.
     Hillslope erosion resulting from  runoff from a high-intensity
thunderstorm near Matt, Colorado, was measured  by  Hadley  and
Lusby  (1967). They also used pins previously driven in the ground
for measurement of erosion losses. From a 12-acre watershed, they
found  an  erosion of 18  tons per acre during a  0.90-inch storm with
a 0.51-inch runoff. The maximum intensity of rainfall for  a 10-
minute period was 1.98 inches per hour. Here, again, the climate was
arid to semiarid, the average annual precipitation being 8.3 inches.
Although  this would not be considered  an unusual storm in the Corn
Belt, runoff and erosion of this magnitude from this type of storm
would be highly  unusual in the Corn Belt unless antecedent moisture
was very high.
     Ruhe and Daniels (1965) measured deposition that had occurred
over a period of  several  thousand years and for the period from when
the area was first settled until  the present.  Older deposition rates


 were determined by carbon dating,  and deposition rates  during the
 postsettlement period were measured from tree-ring data. These data
 are very interesting, but because of the long periods involved, it is im-
 possible to relate the deposition  to a postsettlement event or series of
 events. Postsettlement (125 years) deposition, however, corresponded
 to soil losses of  10 tons per acre per year on an Adair County, Iowa,
 site compared with 1.0 ton per year in the preceding 6,800 years.
      Some general comments  can be made about the  applicability of
 geological data to the pollution problem. Most of the detailed studies
 of erosion seem  to be in  the arid and semiarid  areas.  Measurements
 are made on the range or pasture land or in badlands areas where
 there is little vegetation.  Many  studies are made on  spectacular ex-
 amples where land features stand out rather than on more subdued
 arable fields. Most estimates  are on the basis of deposition and for
 long periods—hundreds of years. Shorter-term  estimates are seldom
 on an individual storm basis.  Schumm (1964)  emphasizes that the
 need  for data on erosion processes is  pressing, not only as a guide
 for better land management, but also as a basis for explaining land
 forms as functions of current erosion processes and erosion rates.  He
 is particularly interested in semiarid regions of the western United
 States where erosion proceeds at above average rates.
      Schumm (1969) shows the relationship of erosion and deposition
 to landform characteristics. Studies such as these are  very helpful in
 understanding the role of geomorphic processes in field erosion.

     A concentrated effort is being made by the Soil and Water Con-
 servation Research  Division  of  the Agricultural Research  Service,
 U.S. Department of Agriculture, to develop an erosion model. The
 basic model, as now conceived, considers (1) soil detachment by rain-
 fall, (2) transport by rainfall, (3) detachment by runoff, and (4) trans-
 port by  runoff. These are considered  as  separate but interrelated
 phases of soil  erosion by water (Meyer  and Wischmeier, 1968). An
 example of erosion model results is given  in Figure 1.1 for a com-
 plex slope averaging 8%.  Rainfall intensity was 2 inches per hour,
 and infiltration rate was 1 inch per hour. Comparable results can be
 obtained from  a number of slope shapes and rainfall intensity-infil-
 tration relationships.  The general model can be expanded to a more
 detailed one by introducing other components.
     The advantage  of this model  over  the empirical model is that
 the dynamics of each phase will be described by fundamental hy-
 draulic and hydrologic relationships and  by parameters describing the
 soil properties  that  influence  erosion.  Each  phase  is now being
 studied as a segment  or  submodel by workers at various locations.
Analytical studies of raindrop splash are being carried out, and this
effect is related to soil properties as well as to raindrop size, shape,
 and velocity.
    Studies of soil particle detachment  by  raindrops from soil beds
consisting of a  number of soil types, conditions of soil management,
and size  distribution of clods are being  made at Ames, Iowa. Non-

                       CHAPTER 1 / SOURCES AND TRANSPORT PROCESSES / 9
 60.00 r

         « SEDIMENT LOAD
                       100.00   150.00
                         SLOPE LENGTH
           SLOPE  SHAPE

                I	I	L
      0.00     50.00    100.00   150.00  200.00
                         SLOPE  LENGTH
250.00   300.00
     FIG. 1.1.  Erosion model results plotted for  a  complex  slope averag-
     ing 8% steepness. Rainfall intensity was 2.0 in/hr, and  infiltration
     rate was 1.0  in/hr.  Graph ordinates are relative, but the slope  units
     have been assumed to be in  feet and the erosion  units may be con-
     sidered  as pounds per foot of slope width.  The upper graph shows
     the available  detached soil, the  transportation capacity, and the re-
     sulting  sediment load  plotted  against slope length.  The  middle graph
     shows the  slope shape studied with  an expanded vertical scale.  The
     lower graph shows the net erosion  loss for each  increment. (Meyer
     and Wischmeier, 1968.)

cohesive soils (sands) have  shown  the highest  detachment by  rain-
drops.   When large clods (one-fourth to one inch) were concentrated
on the surface, peak rates of detachment were delayed (Moldenhauer
and Koswara, 1968).  Clod stability  to raindrop impact was inversely
related  to clay content.  The higher the content of montmorillonite
clay, the weaker were the clods.  Concentrating the large aggregates
in the surface kept the infiltration rate high longer than if they were
mixed  with  fine material (Moldenhauer and Kemper,  1969).   The


 effect of surface sealing on detachment has been  studied (Molden-
 hauer and Koswara, 1968).
     Work is being done to determine the effect of the suction (ten-
 sion) gradients on stability, and consequently their effect on detach-
 ment  and  erosion.  Plans are  being made to  develop a  computer
 model for infiltration, runoff, clod breakdown,  and soil detachment
 from the beginning of  the first rain after tillage  throughout each
 succeeding rain.
     When pore size at the soil surface has been reduced to the point
 that rainfall intensity exceeds intake rate, runoff begins.  As runoff
 water becomes concentrated in  the  lower areas of the surface mi-
 crorelief, small rills begin to form because of detachment and trans-
 port of  soil by flowing water.  This  process is  being  studied at La-
 fayette,  Indiana.
     Gully Erosion

     In most areas of the humid region, gully erosion is a relatively
 small  percentage  of  gross erosion.  A study  of 113  watersheds
 (Glymph, 1956), ranging from 23 acres to 437 square miles and lo-
 cated in the humid area of the United States, showed that sheet and
 rill erosion accounted for 90% or more of the sediment yield in half
 the watersheds.  In about 20%  of the cases studied, however,  50%
 or more of the sediment was derived from gullies.  In most instances
 of relatively large  sediment production  from gullies, watersheds of
 less than 1 square mile were involved.  In three  instances, stream
 channel erosion  contributed more than  40% of the sediment yield.
 In  a region of loess-covered sands in Mississippi,  gully erosion con-
 tributed about 20%  of the sediment yield for  watersheds ranging
 from about 8 to 120 square miles (Miller et al., 1963).
     The prediction of gully growth rates  has received little atten-
 tion, although such information is often needed  for design and cost-
 benefit analysis for the Public Law 566 program. A study of 61 gul-
 lies in the deep-loess area of southwest Iowa (Beer and Johnson, 1963)
 related change in gully area to such factors as watershed area, pre-
 cipitation, channel geometry, and terraced area. The R2 statistic used
 to measure the relative fit (R- measures the percentage of total devia-
 tion attributed to regression) varied between  0.70 and  0.89 for 5
 linear regression models with 6  or  7  "independent" variables.  The
 R2s for logarithmic models were lower, but  fewer problems with cor-
 relation between "independent" variables were encountered. Using
 the ratio of the predicted growth rate (equation  derived from  same
 field measurements) to the growth rate measured in the field  as a
 standard, Beer was able to predict growth  rate within 50%  in half
 the cases.  A study (Thompson,  1964) of 210 gully heads  located in
 6 states related gully  advancement to area, a soil factor,  rainfall,
gully depth, and channel slope.  R2 value for the  equation of best fit
was 0.77. An Israeli study (Seginer, 1966) related gully advance to
 area.  Both Thompson's and Seginer's studies showed the gully  head
 growth rate to be proportional to the square root of the contributing
watershed area. The scarcity of reported literature and the approaches

                      CHAPTER 1  / SOURCES AND TRANSPORT PROCESSES / 11

taken to date indicate  that the mechanics of gully growth are poorly
understood. Predictions of gully growth  are usually made by pro-
jecting observed rates  obtained  from the recent past through use of
aerial photograph measurements and by interview.

     Sediment yields are ordinarily reported in tons per acre per year
in agricultural literature  and in tons, or acre-feet, per square mile
per year in  engineering and geological literature. The ratio of sedi-
ment yield to gross erosion is termed sediment delivery ratio, a ratio
commonly used in  design  of small reservoirs.  The  sediment de-
livery ratio is  used to express the fact that the sediment production
per unit area decreases as  the watershed area increases.  There is
strong evidence to support  this as shown in Table 1.2 (Gottschalk,
1964).  Even  though qualitative  reasoning would indicate  this is
true, no cause and effect relationships are available to represent the
decrease in the sediment delivery ratio  with area. The percentage
of area of lesser slopes increases  with drainage area.  Groundwater
in contrast to surface water ordinarily contributes a larger percentage
of flow to a stream, and  local storms  initiate erosion in only a por-
tion of a watershed. From  this  it  would seem that sediment pro-
duction per unit area should decrease with size of watershed if all
other factors remain constant.
     The primary  sources of sediment-yield information are reservoir
sedimentation  surveys  and  suspended  load samplings.  Reservoir
surveys have the  advantages of providing long-term information in
some instances and of including bed-load sediments (sediment mov-
ing but not in suspension).  Disadvantages of the surveys are loss of
sediment through spillage, unavailable individual storm runoff events,
and difficulty in measuring sediment density. A summary of reservoir
sediment deposition surveys is published  periodically (U.S.  Depart-
ment of Agriculture, 1969).  Suspended  load samplings have the ad-
vantage of providing data for specific storm events; time required to
obtain long-term records, difficulty in obtaining accurate data, and
cost are disadvantages.  Federal agencies are the primary source of
the limited sediment-yield data.  Most of the suspended sediment-yield
data are published by the U.S. Geological Survey.
TABLE 1.2.  Sediment production rates for drainage  areas  in the United

                         Number  of
Watershed Size          Measurements          Average Annual  Rate

 (square miles)                                 (acre-feet/square mile)
Under  10 	     650                       3.80
10-100   	     205                       1.60
100-1,000 	     123                       1.01
Over 1,000 	     118                       0.50

Source:  Gottschalk (1964).


      Sediment Yields from Watersheds

      Because of the complexity of the sedimentation process,  only
 statistical attempts have  been used  to relate yield to selected ob-
 served measurable system inputs.  Several regression equations have
 been developed, primarily for watersheds less than 50 square miles.
 Some estimate  of the gross erosion (on-site sheet plus gully erosion)
 is  required  for  most of these equations.  Other  factors related  to
 drainage density and channel geometry are  added.
      Glymph (1954) discusses several of  the equations.  Equations
 were developed for  South Dakota  stock ponds,  California forested
 watersheds, and western  Iowa  and  Texas  watersheds.  The equa-
 tions typically present sediment yield as some function of climate,
 area, watershed geometry, watershed management (if variable), and
 relative capacity of the reservoir if predictions are based on reservoir
 sedimentation surveys.  An  example  of such an  equation (Glymph
 et al., 1951) was developed for 36 western Iowa and eastern Nebraska
 watersheds ranging in area from 0.036 to 2,800  square  miles.

     Log  S = 1-0078 Log E + 0.6460 Log 10 N - 0.1354 Log  100 W
              - 1.4130
          S = Sediment yield, tons per square mile per year
         E = Gross erosion, tons per square mile per year
         N = Number of rainfall  events (average annual number
             equal to or exceeding one inch per day during the grow-
             ing season)
         W = Net drainage area, square miles.

     About  90% of the sediment-yield data points calculated by the
 above equation  were within ±50%  of the   points determined by
 field measurement.
     A more severe test for such prediction equations is to use data
 from the same or a similar area (Beer et al.,  1966) but independent
 of the equation  development. Four methods, three based on equa-
 tions and one based on gross erosion, delivery ratio,  and trap  effi-
 ciency, were tested by plotting the ratio of predicted yield to measured
 yield for 24 reservoirs.  Figure  1.2  shows the  discrepancy  among
 equations for the various reservoirs.  About 40% of the plotted points
 lie in a band in which the actual deposition was  predicted within ±
     The sediment  delivery ratio approach is  used in  Soil Conserva-
 tion Service watershed designs (Adair and Renfro, 1969). A plot of
 delivery  ratio against watershed  area  is defined for a given land re-
 source area and is limited to that land resource area. Recent studies
for river basin planning (U.S. Corps of Engineers, 1968)  indicate a
 similar approach was used in developing logarithmic plots of annual
 sediment yield as a function of area for a given land resource area.
Lines for all land resource areas are drawn parallel and indicate an
exponent (slope) of about  —0.11.  The annual sediment yields  rep-
resenting field data range from  about 30%   to about  300%  of the
yields indicated by plotted lines.  A similar  earlier logarithmic  plot

                      CHAPTER 1 / SOURCES AND TRANSPORT PROCESSES / 13
                                                      • —ILLINOIS
                                 fl               0     •—TP-97
                                            °    »     A— MODIFIED
                                                        GROSS EROSION

i 0^

                                •   •    o
  •            •* A
s   .         e           .               *
        I  2 3 4 5 6 7 8  9  10 II 12 13  14 15 16 17 18 19 20 2122 23 24
                       RESERVOIR  NUMBER
     FIG.  1.2.  Comparison of measured reservoir sedimentation with  that
     predicted by four methods.

(Glymph, 1951) of data from  51 suspended load measurement  sta-
tions and  reservoirs  from different  parts of the country  showed  a
slope of the data envelope lines of about — 0.12.  The sediment-yield
plots from the upper Mississippi basin study ranged over three loga-
rithmic cycles for the entire  basin drainage area;  Glymph's data
ranged over somewhat less than two cycles.
     Thus, measurements indicate that extreme variations  in sedi-
ment yield may occur in  a region made up of different land resource
areas defined by land  use, topography, climate, and soil types. Iowa
provides a good example of the effects of topography and soil types
in a region in which agricultural land use is heavy and rainfall char-
acteristics  are similar. The relatively flat and recently glaciated area
of north-central Iowa, which is characterized by surface depressions,
has  sediment yields of about 50 tons per  square  mile per year for a
100-square-mile watershed.  The rolling loess hills of western Iowa
produce sediment at a rate of  about 6,000 tons  per square mile per
year for  a 100-square-mile watershed (U.S. Corps of Engineers,
     Reservoir Sedimentation

     Three aspects of reservoir sedimentation related to delivered
sediment  are  trap efficiency, specific weight of deposited material,
and  distribution of sediment. The trap efficiency of a reservoir is a


 measure of the efficiency of the structure to retain the incoming sedi-
 ment, expressed in percent.  The trap efficiency depends primarily
 on the particle  fall velocity and rate of flow of water through  the
 reservoir. Trap efficiencies of reservoirs usually decrease with time as
 sediment accumulates.  Trap efficiency studies (Brune, 1953) indi-
 cate most large reservoirs have trap efficiencies greater than  80% .
 Brune presented envelope curves of trap efficiency as a function of
 the ratio of reservoir capacity (acre-feet) to annual inflow (acre-feet).
 Few good data are available on trap efficiency, especially for  small
     The specific weight of sediment is needed to obtain a meaningful
 measure of deposited sediment. The specific weight is expressed in
 terms of dry weight per unit volume in place. Recent studies provide
 a measure of in-place specific weight (Heinemann and Dvorak,  1963;
 Lara and Pemberton, 1963). The range of specific weights for  domi-
 nant grain sizes is as follows:

 Dominant Grain  Size         Permanently Submerged         Aerated
                                       (pounds per cubic  foot)
        Clay                         40-60                 60-80
        Silt                          55-75                 75-85
        Sand                         85-100                85-100

 Specific weights within reservoirs can be estimated if the sand, silt,
 and clay  percentages  and reservoir drawdown characteristics  are
 known. Lara  and Pemberton's data indicate standard errors of pre-
 diction of 11 to  14  pounds per cubic foot for 1,316 samples obtained
 from many reservoirs under different types of operation. Some river
 bed  sediments were included.  The  standard  error indicates that
 68%  of the measured  specific weights were within 11 to 14 pounds
 per cubic foot of the independently predicted  specific weight.  The
 correlation coefficients (R) ranged from 0.57 to  0.84.
     The distribution of deposited sediment is affected by particle size
 and velocity of  flow through the reservoir. The sediment may  be
 deposited  in the form of a delta at the head of a reservoir or deposited
 as a blanket over the bottom of the reservoir. The delta deposits con-
 tain primarily the coarser material  in  transport;  the bottom deposits
 are primarily clay.  Graphs derived  for different reservoir shapes
 have been developed that indicate the proportion of the sediment  lo-
 cated below indicated  percentages  of reservoir depth  (Borland and
 Miller, 1960).  In a study of 23 small reservoirs (Heinemann, 1961),
 a regression equation was developed  that  expressed the percentage
 of original reservoir depth filled with sediment in terms of percentage
 of original storage depleted,  reservoir geometry, storage capacity,
 and  the capacity-watershed ratio.  The coefficient of determination,
R2, for the equation was 0.91.  Graphs of sediment distribution were
 also presented.

     Sediment transported by a  stream may be divided between bed
load and suspended load, depending on mode of transport. Bed load


moves on or very close to the bed, but suspended load is maintained
in the flow by turbulence.  Another term sometimes used is "wash"
load or that portion in transport made up of fine particles not found
in quantity in the bed. The term bed-material load describes that
portion in transport of which the bed is largely composed.  The bed-
material  load may be in suspended  transport.  Although  it is arbi-
trary and depends on velocity, water temperature, and sediment size
available, a division in size may be made at 62-M.
     Suspended Transport

     The suspended load in transport through a unit width of stream
cross section is  determined by the product of concentration times
velocity integrated over the depth of flow.  Several measurements of
vertical distribution of sediment are usually taken at a cross section,
except in very small streams.  Most samples are taken with a depth-
integrating  sampler which  intercepts  a  representative sample  in
the profile while the sampler is lowered and raised. The point-integrat-
ing sampler intercepts a water-sediment sample at a point in the pro-
file and  enables construction  of sediment concentration curves.
     In most streams of  the humid area of mid-America, most sedi-
ment in transport is suspended.   Measurements in the Mississippi
River at  St.  Louis over 10 years indicated that  95%  of the total sedi-
ment discharge was as suspended  load; 85%  of the suspended load
was silt and clay (Jordan, 1965).  In Iowa probably more than 85%
of transported sediment  is in  suspension; 90%  or more of  the sus-
pended particles  are in  the silt and  clay range, as indicated from
average particle  size distributions  cf 7 rivers (Hershey, 1955).  Al-
though there are exceptions, most reservoir deposits contain less than
10%  sand (Gottachalk,  1964). For  303 samples collected from  32
reservoirs in Illinois the sand content was usually 2 to 5% of the
sample (Stall, 1966).
     The capacity of a stream to transport fine particles is restricted
by  the available  supply; the  supply is  usually much  less  than the
stream can  convey.  In instances where the bed load is appreciable
the supply of particles is usually greater than the stream can trans-
port.  Thus  the  amount transported  as bed load depends  on flow
     The amount of sediment in  suspension  is extremely  variable
and depends on  local hydrologic  conditions.   In  general,  the sus-
pended load increases  faster than  the discharge  and  can be ex-
pressed by L = aQb, where L — sediment  load in tons per day, Q  is
stream discharge in cfs,  and  a and b are constants.  The constant b
typically lies between  2 and  3 (Leopold  and Maddock, 1953). Al-
though the  equation roughly expresses the relationship,  a scatter
over two logarithmic cycles is not uncommon.  The  concentration
of suspended sediment is related to climate and physiographic area.
For example, the maximum concentrations of sediment in  the por-
tion of the Des Moines River that  drains the most recently glaciated
area is seldom over 5,000 ppm,  but records  for  the Soldier River
located in western Iowa  deep loess show several concentrations over


 200,000  ppm  (U.S. Corps  of Engineers,  1951).  An instantaneous
 concentration  of 276,000 ppm was sampled in Waubousie Creek of
 that region (Hershey,  1955).  As indicated previously, the concentra-
 tion of suspended sediment in a downstream direction generally de-
     The wash load of a stream travels at about  the velocity of the
 water.  Thus the travel time of clay and silt to a critical downstream
 point would be about  the same as that of dissolved solids.  Bed-load
 material would move more  slowly because of the nature of transport.
 The peak concentrations associated with surface  runoff occur near,
 and in most cases before,  the peak discharge in very small water-
 sheds  (Dragoun  and Miller,  1966).  The  changes in concentration
 with time occur rapidly. In  larger watersheds the peak concentra-
 tions tend to coincide with the peak flows, although local inflow from
 small watersheds may significantly alter concentrations.
     The distribution of sediment in a stream varies laterally across
 the stream  and through the vertical flow profile.  Examples indicate
 that a  variation from  the  average stream concentration  of ± 20%
 is common (Task Committee on Sedimentation,  1969).  Very large
 variations may occur at stream sections  below a tributary stream
 with different sediment transport characteristics.  Sediment  par-
 ticles in the coarse silt through clay range tend to be uniformly dis-
 tributed in the vertical.  But the  sand concentration gradient  de-
 creases from the stream bed upward.   A  mathematical expression
 (based on theory of  turbulence) is  available that defines the concen-
 tration gradient for a given particle size, if the concentration of the
 given size at a given elevation is known (Rouse, 1938).
     Bed  Load

     Considerable effort has been expended in developing bed-load
formulas, but  there is  not  agreement in  the literature  regarding
which approach is best (Shulits and Hill, 1968).
     Several comparisons have been made (Vanoni et al., 1961; Jor-
don, 1965).  Variation  between predicted (by formula) and measured
bed loads may be  greater than  100%.  The most  commonly used
formulas are the Einstein, Schoklitsch, and Meyer-Peter and Muller
formulas (Shulits and  Hill, 1968).  Some of the formulas are used
routinely in planning  and operations (Adair and Renfro,  1969).
     If the fall velocity (particle  size), average stream velocity, and
nature of the  channel  are known, an estimate of the bed load may
be made.  Curves relating bed-material discharge per foot of width
and mean velocity have been developed for sand-bed streams (Colby,
1961). The bed-load discharge may also be estimated in terms  of
percentage of suspended load, where data on suspended load are
available (Lane and Borland,  1951).  Suspended load concentration,
type of material forming the channel, and  texture of  suspended
material are needed for the estimate.  In some  cases where the con-
centration is  over  1,000 ppm,  channel  material  is sand or con-
solidated clay,  and the suspended material is  less  than 25%  sand,
predicted bed load varies from 2 to 15%.

                      CHAPTER 1 / SOURCES AND TRANSPORT PROCESSES / 17


     Although considerable progress has been  made  in  the  last 30
 years, the science of erosion and sediment transport needs  to  ad-
 vance considerably if it is to  be sufficiently flexible  for use  in  de-
 tailed planning.  Most of the  approaches to design and  planning
 are  empirically based and are  subject  to the  restraints of the  ob-
 servations from which they are developed.  Considerable study  of
 on-site erosion and river transport, especially of sand-bed streams in
 the  West, is evident.  The relationship  between on-site erosion and
 the  subsequent response in streams is poorly defined quantitatively.
     A few points stand out in relation to pollution.  Most material
 in transport is in suspension, and is in the silt and clay size  range.
 Most of  the fines in transport in streams are evidently derived from
 surface erosion.  In regions of erosive soils and  well-defined drainage
 systems,  10  to 30 tons per acre per year are delivered to streams if
 vegetation cover is poor.   Concentrations  of suspended sediment  of
 25,000 to 150,000 ppm  are encountered for short  times.  On  the
 other  hand, in flat  country with poor drainage  development,  the
 sediment loads and  concentrations are relatively  low even  though
 the  land is cropped intensively.  While bed loads are  low, the effect
 of man on bed transport may be small.
     The works of man are primarily related to change cf cover and
 alteration of the hydraulic system through which is  transported  the
 water and  sediment.  The options  of altering cover and  channels
 remain open.
     According to Vanoni (1963):

    The theoretical  treatment of  the sedimentation problem is very diffi-
     cult, and will develop slowly.  It will be  based on the understanding
     gained from experiments  rather than by some break-through by a
     stroke of theoretical genius. However, in  order for the experiments to
     contribute understanding, they must be designed carefully to answer
     certain questions or to prove or disprove  hypotheses based on reason-
    ing and  results  of other investigations.  Considering the primitive
    state of knowledge of sedimentation, contributions can be made  in
    many ways.

Adair, J. W., and Renfro, G. W. 1969.  Sedimentation considerations
     in watershed design.  Paper No. 69-209 presented at the meet-
     ing of the Am. Soc. Agr. Engrs., June 1969, Lafayette, Ind.
Beer, C. E., and Johnson, H. P.  1963.  Factors in gully growth in the
     deep loess  area of western Iowa.  Trans. Am. Soc. Agr. Engrs.
Beer, C. E., Farnham, C. W., and Heinemann, H. G. 1966. Evaluating
     sedimentation prediction  techniques in western  Iowa.  Trans.
     Am. Soc. Agr. Engrs. 9:828-33.
Borland, W. M., and Miller, C.  R.  1960.  Distribution  of sediment in
     large reservoirs.  Trans. Am. Soc.  Civil Engrs. 125 (1): 166-80.
Browning, G.  M., Parish, C. L., and  Glass, J. A.  1947.  A method for
     determining the use and limitation of rotation and conservation


     practices in control of soil erosion in Iowa. Am. Soc. Agron. J.
 Brune, G. M.  1953.   Trap efficiency  of reservoirs.  Trans.  Am.
     Geophys. Union 34:407-18.
 Colby, B. R.  1961. Effect of depth of flow on discharge of bed ma-
     'terial. U.S. Geol.  Survey Water Supply Paper 1498-D.
 	.  1963.  Fluvial sediments—a summary of source, transporta-
     tion, deposition, and measurement of sediment discharge.  U.S.
     Geol. Survey Bull.  1181-A.
 Dragoun, F. J., and Miller, C. R.  1966. Sediment characteristics of
     two small agricultural watersheds. Trans. Am. Soc. Agr. Engrs.
 Einstein, H. A.  1964.  River sedimentation.  In  Handbook of applied
     hydrology, ed. V. T. Chow,  pp.  17-35 to 17-67.  New York:
 Glymph, L. M.  1951.  Relation of sedimentation  to accelerated ero-
     sion in the Missouri River Basin.  USDA, Soil Conserv. Serv.,
 	.  1954.  Studies  of sediment yields from watersheds.  Intern.
     Union Geodesy Geophysics, Intern. Assoc. Sci. Hydrol. Publ. 36,
     pp. 178-91.
        1956.  Importance of sheet erosion  as a source of  sediment.
     Trans. Am. Geophys.  Union 38:903-7.
 Glymnh, L. M., Heinemann, H. G., and Kohler, V. 0.  1951.  Unpub-
     lished study from files of U.S. Soil  Conserv.  Serv.,  Lincoln,
 Gottschalk, L. C.  1964.  Reservoir  sedimentation. In Handbook  of
     applied hydrology, ed.  V. T.  Chow, pp.  17-1  to 17-34.  New
     York: McGraw-Hill.
 Hadley, R. F., and Lusby, G. C.  1967. Runoff and hillslope erosion
     resulting  from a high-intensity thunderstorm near Mack,  west-
     ern Colorado.  Water Resources Res. 3:139-43.
 Hadley, R. F., and Schumm,  S. A.  1961.  Hydrology of the upper
     Cheyenne River basin.  U.S. Geol. Survey Water Supply Paper
     1531-B.: 137-98.
 Heinemann, H. G.  1961.  Sediment distribution  in small floodwater
     retarding reservoirs in  the Missouri  basin  loess hills.  USDA,
     ARS  41-44.
 Heinemann, H.  G., and Dvorak, V.  I. 1963.  Improved  volumetric
     survey for small reservoirs.  In Proc. Federal Inter-Agency Sedi-
     mentation Conf. USDA Misc. Publ. 970.
 Hershey, H. G.  1955. Quality of surface ivaters in loiva.  Iowa Geol.
     Survey Water SuDply  Paper  5.
Jordon, P. R.   1965 Fluvial  sediment  of the Mississippi River at St.
     Louis, Missouri. U.S. Geol.  Survey Water  Supply Paper 1802.
Lane, E. W., and Borland,  W. M. 1951. Estimating bed load. Trans.
     Am. Geophys. Union 32:121-23.
Lara, J. M., and Pemberton, E.  L.   1963.  Initial unit weight of de-
     posited sediments. In Proc.  Federal Inter-Agency Sedimentation
     Conf. USDA Misc.  Publ. 970.
Leopold, L. B., and Maddock,  T. 1953.  The hydraulic geometry  of
     stream channels and some physiographic  implications.   U.S.
     Geol. Survey Prof.  Paper 252.
Leopold, B., Emmett, W. W., and Myrick, R. W.  1966. Channel and


     hillslope processes in a semiarid area, New Mexico.  U.S.  Geol.
     Survey Prof. Paper 352G, pp.  193-253.
Meyer, L. D., and Wischmeier, W. H.  1968. Mathematical simula-
     tion of the process of soil erosion by water. Paper 68—732 pre-
     sented at the 1968 winter meeting of  the Am. Soc. Agr. Engrs.,
     Dec. 10-13,  1968, Chicago.
Meyer, L. D., Mech, S. J., Mutchler, C. K., Hermsmeier, L. F., Palmer,
     R. S., Swanson, N. P., Brubenzer, G. D., and Moldenhauer, W. C.
     1965. Symp. on simulation of rainfall for soil erosion research.
     Trans. Am. Soc. Agr. Engrs. 8:63-75.
Miller, C. R., Woodburn, R., and Turner, H. R.  1963. Upland gully
     sediment  production.  Intern. Assoc. Scientific Hydrol.,  Com-
     mission of Land Erosion Publ. 59, pp. 83-104.
Moldenhauer, W. C., and Kemper, W. D. 1969.  Interdependence of
     water drop energy and clod size on infiltration and clod stability.
     Soil Sci. Soc. Am'Proc. 33:297-301.
Moldenhauer, W.  C.,  and Koswara, J.   1968.  Effect  of initial clod
     size on characteristics of splash and wash erosion. Soil Sci. Soc.
     Am. Proc. 32:875-79.
Moore, C. M., Wood, W. J., and Renfro, G. W.  1960. Trap efficiency
     of reservoirs, debris basins, and debris dams.  Am. Soc. Civil
     Engrs. Proc. J. Hydraulics Div. 86, HY2:69-87.
Musgrave, G. W.  1947. The quantitative evaluation of factors in
     water erosion, a first approximation.  /.  Soil Water  Conserv.
Onstad, C. A., Larson,  C. L., Hermsmeier,  L. F., Young, R. A.  1967.
     A method of computing  soil movement  throughout  a  field.
     Trans. Am. Soc. Agr. Engrs. 10:742-45.
Piest, R.  F., and Spomer, R. G.  1968.  Sheet and gully erosion in the
     Missouri Valley loessial region.  Trans. Am. Soc. Agr. Engrs.
Raudkivi, A. J.  1967.  Loose boundary hydraulics. 1st ed. New York:
     Pergamon Press.
Rouse, H.  1938.  Fluid mechanics for hydraulic  engineers.  New
     York: McGraw-Hill.
Ruhe, R. V., and  Daniels, R. B. 1965.  Landscape erosion—geologic
     and historic.  /. Soil Water Conserv. 20:52-57.
Schumm, S. A.  1964.  Seasonal variations of erosion rates and proc-
     esses on hillslones in western Colorado.  Ann. Geomorphol.  Sup-
     plement 5:215-38.
	.  1969.  A  geomorphic approach  to  erosion  control  in  semi-
     arid regions.  Trans. Am. Soc.  Agr.  Engrs.  12:60—68.
Seginer.  I.  1966.  Gully  development  and sediment yield.  Israel
     Ministry of Agr.  Soil Conserv. Div. Res.  Rept. 13.
Shuh>s.  S.. and  Hill,  R. D.   1968.  Bedload  formulas.  University
     Park, Pa.:  Dept. Civil Eng. Hydraulics Lab. Bull.
Smith, D. D. 1941. Interpretation of soil conservation data for field
     use.  Agr. Eng. 22:173-75.
Spraberry, J. A., and Bowie, A. J.  1969. Predicting sediment  yields
     'from complex watersheds.  Trans. Am. Soc. Agr. Engrs. 12:199-
Stall, J. B.  1966.  Man's role in affecting the sedimentation of streams
     and reservoirs. 111. State Water Survey Reprint  Series 68.
Task Committee  on Sedimentation.  1969.  Sediment measurement:


     fluvial  sediment.  Proc. Am. Soc. Civil Engrs. 95 (HY5): 1477-
Thompson, J. R.  1964.  Quantitative effect of watershed variables on
     rate of  gully-head advancement.  Trans.  Am. Soc. Agr. Engrs.
U.S. Corps of Engineers.  About  1951. Suspended sediment in the
     Missouri River, daily record  for water years 1937—1948. Corps
     of Engrs., Missouri River Div., Omaha.
	.  1968. Fluvial sediment.  In Upper Mississippi River compre-
     hensive basin study, Draft 3, appendix G.  North Central  Div.,
U.S. Department of Agriculture.  1968.  A national program of  re-
     search  for environmental quality. Joint Task Force Rept. of the
     USDA and the  state universities and land-grant colleges.
	.   1969. Summary of  reservoir sediment  deposition surveys
     made in the United States through 1965.  Misc.  Publ.  1143.
Vanoni, V. A. 1963. Review of research activities in sedimentation.
     In Proc. Federal Inter-Agency Sedimentation Conf. USDA Misc.
     Publ. 970.
Vanoni, V. A., Brooks, N. H., and Kennedy, J. F.  1961.  Lecture notes
     on sediment transportation  and  channel stability.  Pasadena,
     Calif.:  W. H. Keck Lab. of Hydraulics  and Water Resources,
     Calif. Inst. of Technol. Rept. KH-R-1.
Wischmeier,  W. H.  1959.  A  rainfall  erosion index for a  universal
     soil-loss equation. Soil Sci.  Soc.  Am. Proc. 23:246-49.
	.  1960.  Cropping-management factor evaluations for a uni-
     versal soil-loss equation. Soil Sci. Soc. Am. Proc. 24:322—26.
       1962. Rainfall erosion potential. Agr. Eng. 43:212-15.
Wischmeier,  W. H., and Mannering, J.  V.  1969.  Relation of soil
     properties to its credibility.  Soil Sci. Soc.  Am. Proc.  33:131-37.
Wischmeier, W. H., and Smith, D. D.  1958. Rainfall energy and its
     relationship to soil loss. Trans. Am. Geophys. Union  39:285-91.
	.  1965.  Predicting rainfall-erosion  losses from the  cropland
     east of the Rocky Mountains.  USDA Handbook 282.
Young, R. A., and Mutchler, C.  K.  1969. Effect of slope shape on
     erosion and  runoff. Trans. Am. Soc. Agr. Engrs.  12:231-33,

     IN  WATER

      §  HE sediment that is carried off sloping lands and transported
into surface water supplies has been called the greatest single pollut-
ant of our natural waters.  In a certain sense its physical effects are
much more obvious than its chemical effects. The clogging of navi-
gation channels and the silting of lakes and reservoirs are expressions
of the physical existence of the sediment.  It is also a costly existence,
for millions of dollars are spent annually for dredging operations to
remove this material from  its place  of deposition.  Relatively little
is known about the chemical effects of sediment on the water supply
in which it resides  or its influence on the chemical status of flowing
water during transport.
     There has been  growing concern for  the  quality of  surface
waters, particularly in  terms of  the  nutrient levels that  permit
nuisance growth of aquatic plant life. The two elements most closely
associated with these noxious growths are nitrogen and phosphorus.
These elements are also  closely associated with  agriculture, for they
occur in all plant life. Since these are the chemicals most apt to be
in insufficient  supply for crop growth, they are the nutrients most
frequently supplied as fertilizers.  Fertilizers are applied to the sur-
face of soils and thus are quite vulnerable to removal by erosion.  It is
this eroded topsoil which makes up the bulk of the sediment being fed
into  surface water supplies, and it is this material that we are con-
cerned with in this  discussion.

     The mineralogical composition of sediment, both suspended and
bottom material, is as complex as that of the soil from which it was

     R. F. HOLT is Soil Scientist, USDA, and Professor, University of Min-
     nesota. R. H. DOWDY is Soil Scientist, USDA, and Assistant Profes-
     sor, University of Minnesota.  D. R.  TIMMONS is Soil Scientist, USDA.
     Contribution from the Corn Belt Branch, Soil and Water Conserva-
     tion Research Division, ARS, USDA, Morris, Minn., in  cooperation
     with the Minnesota Agricultural Experiment Station.



 derived. Little is known about the mineralogical composition of nat-
 ural water sediments.  However,  from soil erosion literature it  has
 been established that erosion is selective (Massey and Jackson, 1952;
 Massey et al., 1953).  While a discussion of soil erosion is beyond the
 intent of this chapter, some observations are  appropriate.  Manner-
 ing and Wischmeier1 observed that the sediment from five soils con-
 taining 60 to 75% silt was higher in silt than the original soil. It was
 also observed that clay was preferentially removed from two soils of
 high  sand content.
     Working with a Connecticut watershed, Frink (1969) reported a
 large decrease in sand content of the lake sediment compared to that
 of the watershed soils, 18 and 60 to 65% , respectively.  On the other
 hand, the clay content of the sediment was fivefold greater than that
 of  the upland watershed soils.  Similarly, the  silt  content increased
 from 20 to 25% for the watershed soils to 34%  for  the lake sediment.
 Stall  (1964) reported  the following representative comparison be-
 tween Illinois lake sediments and the watershed soils from which they
 were derived: (1) the sand content of the two  remained the same at
 approximately 5% and (2) the clay content of the sediment was 39%
 compared to a 15% clay content for the watershed  soils. These types
 of observations  strongly support the hypothesis of  the selective, size-
 sorting nature of erosion.  It is the product of this erosion that be-
 comes the colloidal stream load and sediments  of natural waters.
     The clay mineralogy of sediment reflects  the mineralogy of the
 soils from  which it was derived.  Frink (1969) identified vermiculite
 (40%), "illite" (hydromica, 35%), and kaolinite (25%) in the lake
 sediment he  studied.   Vermiculite  concentration  decreased  in  the
 sediment compared to  that of the watershed  soils, while illite and
 kaolinite concentrations increased. In the southeastern United States
 where kaolinite is the predominant clay mineral in  soils, it is likewise
 observable in the  sediments of that region (Pomeroy et al., 1965).
 Sediments derived from western Minnesota soils high in montmoril-
 lonite contain montmorillonite as  the major clay mineral (Burwell et
 al.2).  Working with a Tennessee  lake bed  sediment, Lomenick and
Tamura (1965)  observed that "illite" (hydromica)  was  the predomi-
nant clay mineral present in the sediment as well as in the shale for-
mations of the surrounding area. Information as to the presence of X-
ray amorphous clay material (i.e., crystalline relics, amorphous Fe,
and Al hydrous oxides) is indeed limited, but this material must exist
in young sediments and suspended water load  if these clay size par-
ticles  were present in  the  source  materials. Work by  Frink  (1969)
showed the presence of free Fe oxides in a  neutral, freshwater lake.
The free Fe oxide content was highly correlated with the clay content
of the  sediment which suggested  that this material also obeys size
    The chemistry of sediments is in reality the surface chemistry
and properties of  the  colloidal  (inorganic and organic)  fraction of
those sediments. To understand the behavior of the  colloids, one must
     1.  Personal communication.
     2.  Unpublished data.

                                CHAPTER 2 / CHEMISTRY OF SEDIMENT / 23

first look at the structure and characteristics of the colloidal material.
For a complete discussion of clay minerals, the book by Grim (1968)
is recommended.
     Clay minerals  are defined  as  two-dimensional arrays  of Si/O
tetrahedra and Al/ or Mg/O/OH octahedra.  In soil science we often
use the term "clay" to include all inorganic particles 2^ in equivalent
spherical diameter.  Hence, clay could include 2/x quartz as well  as
hydrous Fe and Al oxides. The structure of clay minerals is the super-
imposition of tetrahedral and octahedral sheets in many  different
arrays.  When one tetrahedral and  one octahedral sheet are super-
imposed to form a layer, the resultant is referred to as 1:1 type clay
mineral or the kaolin group. The 2:1 type clay minerals are charac-
terized  by two tetrahedral sheets sharing a central octahedral sheet.
For both the 2:1 and 1:1  type clays, the crystals  are formed by the
stacking of unit layers with a constant periodicity. The third type of
clay minerals  is the 2:1:1 minerals  or the chlorite group.  It is
formed by interlayering 2:1 type minerals with a brucite sheet (either
iron, aluminum,  or magnesium hydroxide  or  any  combination);
hence, 2:1:1 type minerals.
     From a physicochemical  perspective, the two most important
properties of clay colloids are electrical charge  and surface area. To
discuss these characteristics in any  detail is beyond the scope of this
chapter; however, the magnitudes of these properties  for some clay
materials are shown in Table  2.1.  As a general rule, clays carry a
net negative charge. This negative charge arises from two  sources:
(1) isomorphous  substitution in the crystal structure, and (2) broken
bonds  on crystal edges.  Isomorphcus  substitution refers to the re-
placement of tetravalent  Si by trivalent Al in  the tetrahedral sheet
of a clay crystal and/or replacement of trivalent Al by divalent cations
such as Mg in the octahedral  sheet. This type of substitution gives
rise  to  a permanent negative charge on the clay crystal which is
balanced  by a surface layer of cations too large to penetrate to the
crystal  structure which  can be  replaced by  another cation—hence
the term  exchangeable cation.  The  site  and  magnitude  of  this sub-
stitution are used to differentiate clay minerals.  Vermiculite is substi-
tuted in the tetrahedral sheet, while montmorillonite is  substituted
octahedrally.  For this reason and the fact that vermiculite possesses
TABLE 2.1.  Cation exchange capacity and  surface  area of several  clay

                                Cation Exchange           Surface
     Clay                           Capacity                Area

                                   (me1100 g)              (M'/g)
Vermiculite  	    100-150              600-800
Montmorillonite  	      80-120              600-800
Hydromica  	      10-40                 65-100
Chlorite  	      10-40                 25-40
Kaolinite  	       3-15                  7-30
Hydrous oxides 	       2-6                100-800


 a higher surface charge, montmorillonite will expand more readily
 and to a greater extent in aqueous solutions than does vermiculite. In
 contrast to these expanding 2:1  type clay minerals, hydromica has a
 much higher surface charge which is satisfied with K ions. The elec-
 trostatic interactions in hydromica are so great that expansion upon
 hydration  is not observed  generally.  Hence,  the  K ions are non-
 exchangeable, with the result that hydromica has a lower exchange
 capacity than vermiculite or montmorillonite.
     Kaolinite has very  little isomorphous substitution (permament
 charge). Its ion exchange capacity is derived from a second source of
 charge, which is  broken bonds at the edges of crystals.  The charge
 arising from broken bonds is a function of particle size and pH.  As
 particle size decreases, so does the ion exchange capacity. It may be
 either positive  or negative, depending upon the pH  of  the system.
 However,  for pH values in  the neutral and  alkaline range,'the edge
 charge  will be neutral  or negative (Schofield and Sampson,  1954).
 Other clay  minerals have  similar pH-dependent charge, but such
 charge  becomes  of less importance  as the permanent negative
 charge  increases  in magnitude.   Amorphous,  clay  size,  hydrous
 oxides of Fe and Al possess positive charges due to the  protonation
 of  a formerly  shared  OH  group—for example  [(— OHo)^]  (Rich,
     Organic colloids are a significant  constituent of natural sedi-
 ments.  Losses of organic matter  in eroded soil can be  as high as
 1,100 pounds per acre  (Barrows and Kilmer, 1963).   Since organic
 matter is concentrated in the  soil surface and has a low density, it
 is among the first components to be removed.  Thus, the  eroded  soil
 contains more  organic  matter than the surface soil from which it
 came, and enrichment  ratios  of about  1 to 5 have  been reported.
 Unfortunately, little  definitive information is known about the phys-
 icochemical character of this  material, perhaps  due to  the lack of
 understanding of the chemical character of  soil organic matter from
 which it is derived.  The organic colloids may  be cationic and/or
 anionic in  nature or exist as neutral entities with functional groups
 such as hydroxyl, carboxyl, and amino, which may ionize in the same
 manner as  other organic compounds.  If organic colloids of sedi-
 ments bear any relationship to  those in the  soils  from  which they
 were derived, a cation exchange capacity from 250 to 400 me/100 g
 soil  can be expected.  What role  selected  erosion may  play in  the
 qualitative nature of the organic sediment is unknown.  Likewise,
 the physicochemical  character of organic colloids produced in natural
 waters is unknown.
     Chemical reactions of  sediments with  the properties discussed
 above can be divided into two groups: (1) interactions with charged
 ions, and  (2) interactions with  neutral compounds.  Adsorption of
 charged ions by colloids is  referred to  as ion exchange  and  is  the
 straightforward process  of  maintenance of electrical neutrality by
 the interaction of oppositely charged species.  Both inorganic  and
 organic cations are  adsorbed on negatively charged clay colloid  sur-
faces.  The ease or difficulty with which an inorganic cation can re-
place an adsorbed cation is dependent upon  numerous  factors which

                               CHAPTER 2  / CHEMISTRY OF SEDIMENT / 25

include (1) cation concentration, (2) complementary adsorbed cation,
(3) solution anion, (4) nature of the clay mineral, (5) temperature of
suspension, and (6) nature of replacing the cation.  These factors are
discussed by Grim (1968). As a general rule, the higher the valence,
the greater the replacing power of an ion.  Also, the replacing power
for ions of the same valence tends to increase with increasing atomic
radius and decreasing hydration.
     Cation exchange can play a very important role in the chemistry
of natural waters.  In some instances, cesium-137 is a major radionu-
clide contaminant of natural waters (Lomenick and Tamura,  1965).
Sawhney (1964) and Colman et al. (1963)  have shown that vermicu-
lite and  hydromica fix Cs  against  Ca extraction.  Lomenick and
Tamura (1965) observed that hydromica in lake sediments fixed large
quantities of Cs-137 from contaminated wraters.  Hence, .sediments
can serve as chemical scavengers of contaminated natural waters.
Diagenesis of illite from vermiculite by the exchange and fixation of
K from lake waters as suggested by Frink (1969) is another important
exchange reaction in nature.
     Adsorption of  anions such as phosphates  is of great interest to
those concerned with the fate  of nutrients in  natural  waters. Rich
(1968) stated that amorphous Al and Fe hydrous oxides and hydrox-
ides are the most reactive soil colloids  with respect to anion adsorp-
tion.  This type of reaction is  an electrostatic interaction.  Anions
may also  enter into exchange with hydroxyl groups exposed at the
broken edges of crystals or on the surfaces of amorphous hydrous
oxides.  By studying the adsorption of sulfate in soil suspensions,
Chao et al. (1965)  concluded that sulfate ions were exchanging for
"structural" hydroxyl groups.  It is suggested  that  reactions  of the
same type can occur in natural colloidal suspensions.
     The  second type of adsorption  on sediments  is the interaction
of colloids with neutral polar molecules. In natural waters it must be
remembered that water is the  solvent  and that  one is studying  an
aqueous system.  Hence,  in any equilibrium adsorption reaction, the
adsorbate is competing with an  extensively hydrated adsorbent.  This
is  why Hoffmann and Brindley (1960) observed that adsorption of
alcohols from  aqueous suspensions of montmorillonite  did not occur
until  the  compound contained  five or six  carbon atoms.  This same
phenomenon occurs in the  adsorption of sugars onto  clays.   Clapp
et  al. (1968) reported the adsorption of polysaccharides on montmoril-
lonite from aqueous suspension, while it is not possible  to show an
adsorption of mono- and  disaccaride under  the  same  conditions
(Dowdy3). While an exact division between adsorption and no adsorp-
tion of sugars  on a  size basis cannot be made,  it  is possible that mo-
lecular weights in the tens of thousands are required before adsorp-
tion occurs from aqueous suspension.  However, once  adsorption of
uncharged polymers occurs, it is very difficult  to remove them from
the clay surface (Greenland, 1963). This very  strong interaction can
be explained by the development of  many  weak polymer-surface
bonds (van der Waals. H-bonding) and the statistical  improbability
     3.  Unpublished data.


 of simultaneous rupture  of  all bonds at a given time (Greenland,
 1965).  In light of the above discussion,  the published information
 about adsorption  of neutral polar molecules  must be studied very
 critically, if extrapolations are to be made  to  sediments in natural
     Lotse et  al. (1968) determined  that adsorption  of  lindane  on
 lake sediments from aqueous suspensions was correlated significantly
 with both clay and organic matter content of the sediment. In some
 cases the interactions  of clay with organic compounds can enhance
 adsorption reactions.  Lee and Hoadley (1967) showed that the sorp-
 tion of organic materials from natural  water onto clays  activated
 new adsorption sites.  They noted increased adsorption of parathion
 on the clay-organic complex versus adsorption on clay alone.  Hence,
 in some natural situations it may be possible  to observe "chemical
 scavenging" of uncharged pollutants in natural waters.
     Another physicochemical phenomenon worthy of note in nat-
 ural waters is the production of sediments  by precipitation. Theabold
 et al. (1963) observed  the precipitation of hydrous oxides of  Fe, Al,
 and Mn when waters  containing these elements came into contact
 with a  body  of water  of  sufficiently high pH.  Once formed, these
 hydrous oxides enter into other chemical reactions such as the ad-
 sorption of phosphates. In aquatic environments supporting photo-
 synthesis, it is  possible to have sufficient CCX evolved to increase the
 pH of the water to exceed the solubility product of CaCO3—hence,
 precipitation  of CaCO3. Upon cessation of  photosynthesis, Lee and
 Hoadley (1967) suggest that the pH will return to its original equi-
 librium level, followed by solution of the  precipitated CaCO3.  How-
 ever, Chave (1965) observed  that in some situations CaCO3 did not
 redissolve and postulated  that it  had been coated with resistant or-
 ganic material. Lee and Hoadley (1967) also stated that Sr2+, Pb2+, and
 Zn2+ can be co-precipitated with CaCO3 if present in the given system.

     A waterlogged soil becomes differentiated into a surface-oxidized
layer and an underlying reduced layer.  The thickness of this oxidized
zone has been reported to vary from  1 or 2 millimeters to  several
centimeters with an  average of about 20 mm (Mortimer, 1942; Gor-
ham, 1958; Holden,  1961; Patrick and Mahapatra, 1968).  There is
general agreement among researchers that these zones exist,  but
different mechanisms have been proposed as to how the two layers
are formed and maintained.
     Mortimer (1942) described the existence of an oxidized zone and
believed this oxidized layer was maintained by the diffusion of oxygen
from water into the sediment. He suggested the distance of this diffu-
sion into the sediment during winter depended mainly upon the re-
ducing power of the sediment. In  waterlogged soils, Patrick and
Mahapatra (1968)  stated the thickness  of the oxidized layer is deter-
mined by the net effect of the oxygen consumption rate in  the soil and
the oxygen supply rate through the overlying water.  A soil with an
abundant source  of  organic  matter  (energy)  will  utilize  oxygen

                               CHAPTER 2 / CHEMISTRY OF SEDIMENT / 27

faster than it can be supplied through the water,  and this high con-
sumption rate results in a thin oxidized zone.  When the oxygen con-
sumption rate is low, the oxidized layer becomes thicker.
     Gorham (1958) suggested the thickness of the oxidized zone of
lake sediment may depend on two factors: (1) the  turbulent displace-
ment of the uppermost sediments into the overlying aerated water,
and  (2) the reducing  power of the sediments.  When sedimented
plankton decomposes, the winter oxidized zone may  disappear from
the surface downward because of the greater oxygen consumption.
Gorham placed more emphasis on the turbulent mixing of the sedi-
ments with aerated water than on the reducing power of the sediment.
Regardless of the exact mechanism involved in forming the oxidized
and reduced zones, these two layers are extremely important in con-
versions and equilibrium phenomena involving chemical nutrients.
     When soils are submerged, the oxidation-reduction (redox) po-
tential decreases.  Patrick and Mahapatra (1968) reported a range
of —300 to +700 millivolts redox potential in waterlogged soils.  Since
aerated soil measures about 400 to 700 millivolts potential, it appears
the oxidized layer of sediment would be within this same range, and
the reduced layer redox potential should range from —300 to +400
     The reduction of oxidized inorganic components is generally a
sequence of the various redox systems (Fig. 2.1).  Oxygen was found
to disappear at +320 to +340 millivolts (Turner and Patrick, 1968),
nitrate became unstable at +225 millivolts  (Patrick,  1960),  ferric
iron was reduced at +120 millivolts (Patrick,  1964), and sulfate  re-
duction started at —150 millivolts (Connell  and  Patrick,  1968).
Usually the reduction of one component is not completed before  re-
duction of the next most easily reduced component begins.

     The chemistry that is important in the influence of sediments
on the quality of water involves  the nitrogen and phosphorus rela-
tions between the sediment and the water. Nitrogen relationships are
difficult to study because many conversions to different forms occur
for different biological and chemical conditions.
     When sediment  is transported to surface  waters,  it contains
1 1
i i

i i i i
+ 200    +400   +600
                                          FIG. 2.1. The approximate
                                          oxidization-reduction    po-
                                          tentials  at which oxidized
                                          forms of several  inorganic
                                          redox systems become  un-
                                          stable.  (After Patrick and
                                          Mahapatra, 1968.)
          (CORRECTED TO pH 7)

     —*-HN02 —
- HN03
                                                SOIL  LAYER
                                      SOIL  LAYER
      FIG. 2.2.  A schematic diagram of the processes by which ammonium
      fertilizer can be lost from  a  waterlogged soil.  (After Mitsui,  1954.)

nitrogen in the forms of organic-, NH4-, NO2-, and NO3-N.  Before
being deposited the sediment will probably lose soluble organic-, NO2-,
and NO3-N, whereas  the insoluble organic  N and NH4-N will essen-
tially remain with the sediment.  Flooded soils would react similarly
except losses of the soluble components would probably be slower be-
cause more diffusion  and less  turbulence would be involved.
     The scheme for nitrogen  reactions in submerged  soils has been
depicted by Mitsui (1954) and  is shown in Figure 2.2.  Under anaero-
bic conditions, nitrogen mineralization cannot proceed past the NH4-N
stage because insufficient oxygen is available to convert NH4-N to
NO3-N.  Since  the organisms  involved in anaerobic organic matter
decomposition  are less efficient  than  their aerobic  counterpart,  the
conversion of organic matter to NH4-N is slower in waterlogged soils
(Tenny and  Waksman,  1930).  Although the conversion rate was
slower in waterlogged  soils, Waring  and Bremner (1964a, 1964b)
found that more nitrogen  was mineralized for  several soils under
waterlogged conditions than under aerobic conditions.
     According to Patrick and Mahapatra  (1968), denitrification is
one  of  the  major mechanisms  by which  nitrogen is lost  from  a
flooded soil. In the oxidized zone, NH4-N from organic  matter decom-
position or already present  on the sediment base exchange is con-
verted to NO3-N.  This NO3-N either diffuses or is leached into the re-
duced zone where it is converted by certain facultative anaerobic or-
ganisms to No  or NoO and  lost  to the  atmosphere. Broadbent and
Stojanovic (1952) found that only 0 to 6%  of the NO,-N denitrified
in a waterlogged soil  was  reduced to NH4-N. Only  in  specialized
cases is NH4-N volatilization  an  important  mechanism of nitrogen
loss from a waterlogged soil.
     The rate of NO3-N reduction after submergence of a soil can be
quite rapid.  With no additional energy source,  Patrick (1960) re-
ported a NO3-N reduction rate of 15  ppm  per day in reduced soil,
whereas Bremner and Shaw (1958a, 1958b)  recorded a loss of 1,000

                               CHAPTER 2 / CHEMISTRY OF SEDIMENT / 29

ppm NO3-N in 4 days from a submerged soil which had an energy
source added.
     Sediments are apparently poor conservers of nitrogen supplies.
Attempts to overcome poor utilization of fertilizer nitrogen by  rice
under flooded conditions have resulted in the development of a sys-
tem  in which denitrification of  added nitrogen is minimized.  This
involves deep placement of ammonia nitrogen in the  reduced soil
layer where it is protected from  nitrification and subsequent denitri-
     The increase  of fresh organic material to the surface sediments
in lakes and reservoirs by the periodic deposition of aquatic  vegeta-
tion  should create conditions favorable for the rapid loss of NO3-N,
with perhaps slight increases in NH4-N.

     When phosphorus  is added to an aerated soil, it is converted
rapidly  to water-insoluble forms and becomes extremely immobile.
If the soil is submerged continuously or becomes  sediment in a lake
or stream, there  may be a marked increase  in  the  availability of
native and applied phosphorus compared to well-aerated conditions.
The  mechanism of this phosphate release, as given by Patrick  and
Mahapatra (1968), consists of (1) reduction of insoluble ferric phos-
phate to the more soluble ferrous phosphate, (2) release of occluded
phosphate by reduction of the hydrated ferric  oxide coating, (3) dis-
placement  of phosphate from  ferric and aluminum  phosphates by
organic  ions, (4) hydrolysis of ferric and aluminum phosphates,  and
(5) phosphate exchange between clay and organic anions.  However,
the phosphate that becomes  soluble from  reduction of ferric phos-
phate can be refixed if sufficient alumnium is available, and can  also
be refixed as ferric phosphate if Fe2t is oxidized to  Fe3+ in the oxidized
zone. Thus, submergence of  soil does not necessarily increase phos-
phate solubility and availability.
     Under waterlogged conditions, organic matter affects the mech-
anisms  of  reduction  and chelation.  Shapiro (1958) reported both
processes increase soil  phosphate  solubility and  availability, so the
addition of organic matter to the surface of sediments should create
conditions favoring increased availability of phosphorus.  However,
others (Bartholomew, 1931; Paul and  DeLong, 1949) have reported
that a transformation of inorganic phosphorus to the organic form
in flooded  soils reduced the  availability of the phosphorus.  It  is  a
complex system and  contradictory findings are not unusual.
     The equilibrium reactions involving phosphorus in  sediments,
water, and aquatic plants are influenced by many biological,  chem-
ical, and physical factors, making this dynamic system very difficult
to study in situ.  Studies using radiophosphorus placed either in the
bottom  sediments or in the surface water have provided needed  data
about the behavior of phosphorus.
     Rigler (1956) found that only 3%  of the radiophosphorus added
to the surface of a small, acid-bog lake was  lost to  the  sediments.


 He  concluded there was a turnover of "mobile"  phosphorus of the
 epilimnion with  phosphorus of the littoral organisms in 3.5 days.
 The turnover time  (in summer) for soluble inorganic phosphorus
 was about 5 minutes. Hayes et al. (1952) reported that lake  sedi-
 ment increased in radiophosphorus for 10 days  after  deposition in
 the lake surface, but suggest that in  addition to the sediments, aquatic
 organisms are very active in the exchange process.
     Using sediment core samples in the  laboratory, Holden (1961)
 concluded that bottom sediment can slowly take up large amounts
 of phosphate and that about 85% of the phosphate removed by the
 sediment occurred in the aerobic zone which extends to about 20  mm.
 In unfertilized lakes, the phosphorus content of the sediment surface
 was very high relative to the equilibrium  concentration in the over-
 lying water.
     Harter (1968) also found that lake sediment  can absorb a large
 amount of phosphorus from the water.  When more  than  0.1 mg
 phosphorus was  added it was adsorbed in a loosely  bonded form,
 and he suggests  that large influxes of phosphorus into a lake  may
 be held temporarily and subsequently released to aquatic plants.
     Phosphorus  equilibria  between lake bottom sediments and cal-
 cium phosphate  solutions were studied on samples collected from
 two eutrophic lakes in western Minnesota (Latterell et al.4). When
 solutions containing up to 42 ppm phosphate were equilibrated  with
 sediment,  the resulting solutions contained about 0.03  ppm phos-
 phate, so the sediment adsorbed large amounts of orthophosphate.
     The release  of  phosphorus from lake sediments  to lake water
 was investigated  for several lakes at Madison, Wisconsin, by Sawyer
 et al. (1944). They reported that continuous leaching of 1-liter  sedi-
 ment samples for 220 days released 12  and 5 mg of phosphorus
 from Lake Monona and Lake Waubesa sediments, respectively,  as
 compared  to  180 and 90 mg from undigested sewage sludge  and
 storm  sewer sludge for these two lakes, respectively.  The  amount
 of phosphorus removed by this continuous leaching, however,  does
 not indicate the equilibrium concentration.
     Diffusion of  phosphorus from sediment into the overlying water
 is negligible in undisturbed systems.  Hasler (1957) found that the
 percentage as well  as the amount of phosphorus released  to  the
 superimposed water was very small when it was placed  at depths
 greater than 1 cm.
     In a similar  study, Zicher et al. (1956) reported that phosphorus
 placed at y2 inch below the sediment surface  showed only a  very
 slight tendency to diffuse into the above water and did not diffuse
into the water at all when placed at 1-inch depth. Water samples
 taken near the sediment surface contained a  higher percentage  of
 soluble  phosphorus than water samples taken  at greater distances
from the sediment surface.
     The establishment of an upper  oxidized layer  and lower reduced
layer may be  expected in sediments left undisturbed for appreciable
periods  of time, but sediments  covered by  shallow waters and  sub-
     4.  Unpublished data.

                               CHAPTER 2 / CHEMISTRY OF SEDIMENT / 31

jected to wave action may not  exhibit the same characteristics.
Stephenson (1949) found  that agitation of sea water  with bottom
mud may either increase or decrease the concentration of phosphate
in solution.  The  changes in phosphate levels  are ascribed  to  (1)
destruction of organisms with release  of protoplasm,  (2) breakdown
of this protoplasm by bacteria with release of phosphate, and  (3)
absorption of phosphate by bacteria.
     Certain benthic ciliates are capable of splitting inorganic phos-
phorus from dilute solutions of organic phosphates that occur  in
lake sediments (Hooper and Elliott, 1953).  This process may provide
a source of energy supplementary to  that obtained from ingestion
of participate organic matter and bacteria, but its  importance  in
nutrition is unclear.
     Sediment plays an important role in  the assimilation of phos-
phate  during transport in  waters. Keup (1968) has quoted Gessner
(1960) on  studies of the turbid Amazon River that  indicate when
soluble phosphorus concentrations exceed 0.01 ppm, it is sorbed on
finely  divided inorganic suspended material.
     Obviously,  bottom  sediments  can  remove  relatively  large
amounts of dissolved  phosphate  from waters but the partition  of
this removal is not well understood. It is probable that these sedi-
ments act  as  a  control,  removing phosphate  from the water when
the concentration  is above the equilibrium value and releasing phos-
phate to the water when the concentration falls below the equilibrium
point.  The contribution that these sediments make to support algal
growth is not known,  but  without thermal or mechanical  mixing it
is  doubtful that sufficient phosphate  could diffuse  at a rate  fast
enough to support algae more than a few inches from the  sediment.

     Sediment can be considered a major pollutant of surface waters.
However, its contribution  to the dissolved chemicals in lakes and
streams  is largely unknown.  The  composition of sediment closely
resembles the soil from which it is derived but is generally higher
in silt, clay, and organic matter.
     Chemical reactions involving sediment are essentially the sur-
face chemistry of their  colloidal fractions which  is a  function  of
their surface area and electrical charge. As a  result, reactions with
sediment can be divided into interactions with charged ions and with
neutral compounds.   Cation exchange, an example of  the  former,
can  play an important role in the uptake and release  of elements
from sediments.  Adsorption of anions such  as phosphates is also
of great interest with respect to  the fate of nutrients in  natural
waters.  Similarly,  the adsorption  of neutral polar molecules  in-
fluences the chemical composition of the surface water  supplies and
may have distinctly beneficial effects.
     The chemistry of sediments in situ can be surmised from studies
of submerged soils.  An oxidized zone exists at the soil-water interface
and a reduced zone is established beneath the oxidized zone. Nitrogen


 transformations that occur in  these two zones may be postulated to
 explain the inefficiency of nitrogen utilization in submerged culture.
 Also, an  effective mechanism  for controlling the percolation of ni-
 trates into groundwater supplies becomes operative when these zones
 are established.  The existence of anaerobic conditions in the  sedi-
 ment may increase the availability of phosphorus above  that antic-
 ipated under aerobic conditions.
     Sediments carry relatively large amounts of total nitrogen and
 phosphorus into surface waters, but  in  both  cases only a small
 proportion of this total is readily available to the biosystem.  Sedi-
 ments apparently have a high capacity to remove phosphate from
 solution,  but without turbulence  the  release of phosphate from bot-
 tom sediments will not support algal growth at appreciable distances
 from the sediment.  However, if the  concentration of phosphorus
 in  the surrounding  solution drops low enough,  the sediments will
 release phosphorus.  Nitrogen may be added to or removed from
 the biosystem by nitrification  or denitrification in the bottom  sedi-
 ments.  Thus, it appears that sediments have a leveling influence
 on nitrogen and phosphorus concentrations in surface waters.
     Available inorganic nutrients, particularly phosphorus, are rap-
 idly taken up by the biosystem in natural  waters.  They eventually
 become a part of the organic fraction of the sediment  and their
 release back to the waters is not well resolved.

Barrows, H. L., and Kilmer, V. J.  1963.  Plant nutrient losses from
    1 soils by water erosion.  Advan. Agron. 15:303—16.
Bartholomew, R. P.  1931.  Changes in the availability of phosphorus
     in irrigated rice soils.  Soil Sci. 31:209-18.
Bremner, J.  M., and Shaw, K.  1958a.  Denitrification in  soil.  I.
     Methods of investigation. /. Agr. Sci. 51:22-39.
	.   1958b. Denitrification in soil. II. Factors affecting denitrifica-
     tion. J. Agr. Sci. 51:40-52.
Broadbent, F. E., and Stojanovic, B. F.  1952.  The effect  of partial
     pressure of oxygen on some soil nitrogen transformations. Soil
     Sci. Soc. Am. Proc. 16:359-63.
Chao,  T. T., Harward, M. E., and Fang, S. C.  1965. Exchange reac-
     tions  between  hydroxyl and  sulfate  ions in  soil.  Soil Sci.
Chave, K. E.  1965.  Carbonates: association with organic  matter in
     surface seawater. Science 148:1723-24.
Clapp, C. E., Olness, A. E., and Hoffmann, D. J.  1968. Adsorption
     studies of  a  dextran on montmorillonite.  Trans.  9th Intern.
     Congr. Soil Sci. 1:627-34.
Coleman, N. T., Craig, D., and Lewis, R. J. 1963. Ion-exchange reac-
     tions of cesium.  Soil Sci. Soc. Am. Proc. 27:287-89.
Connell, W. E., and Patrick, W. H., Jr.  1968.  Sulfate reduction in
     soil:  effects of redox potential and pH. Science 159:86-87.
Frink,  C.  R.  1969.  Chemical and mineralogical characteristics  of
     eutrophic lake sediments.  Soil Sci.  Soc.  Am.  Proc. 33:369-72.

                               CHAPTER 2 / CHEMISTRY OF SEDIMENT / 33

Gessner, F.  1960. Investigations of the phosphate economy of the
    Amazon.  Intern. Rev. Hydrobiol. 45:339-45.
Gorham, E.  1958.  Observations on the formation and breakdown
    of the oxidized microzone at the mud surface in lakes.  Limnol.
    Oceanog.  3:291-98.
Greenland, D.  J.  1963. Adsorption  of polyvinyl alcohols by mont-
    moriUonite. /. Colloid Sci. 18:647-64.
	.  1965.  Interaction  between clays and organic compounds in
    soils. I. Mechanisms of interaction between clays and defined
    organic compounds.  Soils Fertilizers 28:415-25.
Grim, R. E.  1968. Clay mineralogy. New York: McGraw-Hill.
Harter,  R. D.  1968.  Adsorption of  phosphorus by lake sediments.
    Soil Sci. Soc. Am. Proc. 32:514-18.
Hasler,  A. D.  1957.  Natural and artificially (air-plowing)  induced
    movement of radioactive phosphorus from the muds of lakes.
    Intern.  Conf. Radioisotopes in  Scientific  Res., UNESCO/NS/
    RIC/188 (Paris) 4:1-16.
Hayes, F. R., McCarter, J.  A., Cameron, M. L., and Livingstone, D. A.
    1952.   On the kinetics of  phosphorus exchange  in lakes. /.
    Ecol. 40:202-16.
Hoffmann, R. W., and Brindley, G. W. 1960. Adsorption of non-ionic
    aliphatic molecules from aqueous solutions on montmorillonite.
    Geochim.  Cosmochim. Acta 20:15—29.
Holden, A. V.  1961. The  removal of dissolved phosphate from  lake
    waters  by  bottom deposits.  Verhandl. Intern.  Ver.  Limnol.
Hooper, F. F., and Elliott, A. M. 1953. Release of inorganic phos-
    phorus  from extracts of lake mud by protozoa.  Trans.  Am.
    Microscop. Soc.  72:276-81.
Keup, L. E.  1968. Phosphorus in flowing waters.  Water Res. (Great
   1 Britain) 2:373-86.
Lee, G.  F., and Hoadley, A. W.  1967. Biological activity in relation
    to  the  chemical equilibrium composition  of natural waters.
    Ad-van.  Chem. Ser. 67:319-39.
Lomenick, T. F., and Tamura, T.  1965.  Naturally occurring fixation
    of cesium-137 on sediments of locus trine  origin.   Soil  Sci.  Soc.
    Am. Proc. 29:383-87.
Lotse, E. G., Graetz, D. A., Chesters, G., Lee, G. B., and Newland,
    L. W.  1968.  Lindane adsorption by lake  sediments.   Environ.
    Sci. Technol. 2:353-57.
Massey, H.  F., and Jackson,  M. L.  1952.  Selective erosion of soil
    fertility constituents.  Soil Sci. Soc. Am. Proc. 16:353-56.
Massey, H. F., Jackson, M. L., and Hays, O. E.  1953.  Fertility ero-
    sion on two Wisconsin soils. Agron. J. 45:543—47.
Mitsui, S. 1954.  Inorganic nutrition, fertilization, and soil ameliora-
    tion for loivland rice. Tokoyo:  Yokendo.
Mortimer, C. H.  1942. The exchange  of dissolved substances be-
    tween mud and water in lakes.  J. Ecol. 30:147-201.
Patrick, W.  H.,  Jr.   1960. Nitrate reduction rates in  a submerged
    soil as  affected  by redox potential.  Trans. 7th Intern. Congr.
    Soil Sci. 2:494-500.
	.  1964. Extractable  iron and phosphorus in a submerged soil
    at controlled redox potentials. Proc. 8th Intern. Congr. Soil Sci.
    Bucharest, Roumania 4:605-10.


 Patrick, W. H., Jr., and Mahapatra, I. C.  1968.  Nitrogen and phos-
     phorus in  waterlogged soils. Advan. Agron. 20:323-59.
 Paul, H., and DeLong, W. H.  1949.  Phosphorus studies.  I. Effects
     of flooding on soil phosphorus.  Sci. Agr. 29:137-47.
 Pomroy, L. R.,  Smith, E.  E., and Grant, Carol  M.  1965. The ex-
     change  of phosphate between  estuarine water  and  sediments.
     Limnol. Oceanog.  10:167-72.
 Rich, C. I.  1968. Applications of soil mineralogy in soil chemistry
     and  fertility investigations in  mineralogy in soil science and
     engineering.  Soil Sci. Soc. Am.  Spec. Publ. 3, pp. 61—90
 Rigler,  ±*'. H.  1956.  A tracer study of the phosphorus cycle in lake
     water. Ecology 37:550-62.
 Sawhney, B. C.  1964. Sorption and fixation of micro-quantities of
     cesium by  clay minerals:  effect of saturating cations.  Soil Sci.
     Soc.  Am. Proc. 28:183-86.
 Sawyer, C. N., Lackey, J.  B., and Lenz,  A.  T.  1944.  Investigation
     of the odor nuisance occurring in the Madison lakes, particularly
     Lakes  Monona,  Waubesa,  and  Kegonsa from  July  1943  to
     July 1944.  Report to the Governor's Committee, State of Wis-
 Schofield, R. K., and Sampson, H. R.  1954.  Flocculation of kaolinite
     due to the attraction of oppositely charged crystal faces.  Dzs-
     cussions Faraday Soc. 18:135.
 Shapiro, R. E.  1958.  Effect of organic matter and flooding on avail-
     ability of soil and synthetic phosphates.  Soil Sci. 85:267—72.
 Stall, J. B.  1964.  Sediment movement and deposition patterns  in
     Illinois impounding  reservoirs.   /. Am. Water  Works  Assoc.
 Stephenson, W.  1949.  Certain effects of agitation upon  the release
   ,  of  phosphate from mud. /. Marine Biol. Assoc.  28:371-80.
 Tenny,  F.  G., and Waksman,  S. A.  1930.   Composition  of natural
     organic materials and their decomposition in the soil.  V.  De-
     composition of various chemical constituents in plant materials,
     iinrior -i-~erobjc conditions.  Soil Sc.i. 30:143-6°.
 Theabold, P.  K., Jr., Lakin, H. W.,  and Hawkins, D.  B.  1963.  The
     precipitation of aluminum, iron and  manganese  at the junction
     of Deer Creek with the Snake River in Summit County, Colorado.
     Geochim. Cosmochim. Acta 27:121—32.
 Turner, F. T., and Patrick, W.  H., Jr.  1968.  Chemical changes  in
     waterlogged soils as  a result  of  oxygen depletion.   Proc. 9th
     Intern. Congr. Soil Sci. Australia 4:53—65.
 Waring, S. A.,  and Bremner, J. M.  1964a.  Ammonium  production
     in  soil under waterlogged  conditions as an  index of nitrogen
     availability.  Nature 201:951-52.
	.  1964b. Effect  of soil mesh-size on the estimation of mineral-
     izable nitrogen in soils.  Nature 202:1141.
Zicher,  E.  L., Berger,  K. C.,  and Hosier.  A. D.  1956. Phosnhorus
     release from bog lake  muds. Limnol. Oceanog. 1:296-303.


     PEDIMENTS are primarily soil particles washed into streams by
water. They are products of land erosion  and are largely derived
from sheet and rill erosion from upland areas,  and by cyclic erosion
activity in gullies and drainageways.  It is estimated (Wadleigh, 1968)
that at least half of the 4 billion tons of sediment washed annually
into tributary streams in the United  States  is coming from agricul-
tural lands.
     Erosion can be natural or can be accelerated by man's activities.
Natural or geologic erosion  pertains to that  occurring under natural
environmental conditions. Man-made or accelerated erosion is that
induced by man through reduction of natural  vegetative cover and
improper land use, and occurs at a rate greater than normal for the
site under natural cover.
     Although sediment yield and soil erosion  are not synonymous,
they  are  closely related—and  occasionally used  interchangeably.
Sediment yield can be denned as the quantity of soil material trans-
ported into a stream.  Soil  erosion refers to detachment and move-
ment of soil particles on site, but does not imply movement into
stream channels. Thus, soil erosion is a primary  requisite for sedi-
ment production. The most logical and direct approach to  solving
our agriculturally related sediment problem is the stabilization of the
sediment source by controlling soil erosion through the use of proper
land  and  water management practices or  structures.  In short, to
minimize sediment yield, soil erosion  must be minimized.
     Soil erosion occurs in two basic steps  (Smith and Wischmeier,
1962): (1) detachment  of soil particles  from  adjacent particles by
raindrop impact and splash, and (2)  transport  of  detached particles
by flowing water.  Only when conditions for these steps exist does
soil erosion become a serious problem as a direct source of sediment.
Soil erosion by water is  a physical process requiring energy, and its
control involves the dissipation of  energy—that of falling raindrop
impact and splash, and that due to elevation differences which affect
the flow velocity of water.

     MINORU AMEMIYA is Associate Professor and Extension Agronomist,
     Department of Agronomy, Iowa State University.


      The present state of knowledge concerning the mechanics  and
  hydrology  of soil detachment and transport have already been ade-
  quately reviewed (see  Chapter 1). The properties of sediments from
  agricultural lands have been described and interpreted (see Chapter
  2).  It is the purpose  of this chapter  to briefly review management
  practices for controlling soil erosion to minimize consequent sediment
  production from agricultural lands. Emphasis will  be on sediments
  derived from sheet-rill or microchannel erosion. This does not imply
  that sediments resulting from gully or macrochannel erosion are not
  serious contributors to total sediment  yield.  However,  it has been
  shown that the best method of controlling gully erosion is to minimize
  runoff and sheet erosion above a gully or potential gully  site (Jacob-
  son, 1965).

      The Universal Soil Loss Equation (Smith and Wischmeier,  1962)
 provides a framework for discussing erosion control  measures.  In
 this  equation, soil erosion is described as a function of rainfall, soil
 properties, slope length and steepness, cropping sequence, and sup-
 porting practices.
      At present, little can be done to readily change the amount, distri-
 bution, and intensity of rainfall per se, but measures can be adopted
 to modify its erosiveness—that is, to decrease raindrop impact and
 splash energy or to  decrease  the  amount and velocity  of overland
 flow, or both—to minimize sediment production.
      Soil properties affect both detachment and transport processes.
 Detachment is related  to soil stability,  size, shape, composition, and
 strength of soil  aggregates and clods.  Transport is  influenced by
 permeability of soil to water which  determines infiltration capabili-
 ties and drainage characteristics, aggregate stability which influences
 crusting tendencies, porosity which affects storage and movement  of
 water,  and soil macro-structure or surface roughness which creates a
 potential for temporary detention of water.
     The slope factor determines the transport portion  of the erosion
 process since  flow velocity is a  function of hydraulic gradient which
 is influenced  by slope length  and steepness.  The  remaining two
 factors, cropping sequence and  supporting practices, serve to modify
 either the soil factor or the slope factor or both, as they  affect the ero-
 sion sequence.
    Water runoff  and accompanying soil erosion  resulting  from
 rainstorms  are inversely related to the  water infiltration capacity of
 soil, plus any surface storage capacity.  Hence, one way to prevent
 erosion would be to maintain high water  intake rates and surface
 ponding capacities at levels sufficient to prevent runoff  from all  rain-
 storms  (Meyer and Mannering, 1968).  This is seldom possible, but
 any increase  in  infiltration capacity and surface and subsurface
 storage capacity  can  greatly reduce erosion as well as benefit  crop
water supply.  In most  cases water intake  and storage capacities are
not sufficient to prevent runoff. Soil erosion then becomes a func-

                                 CHAPTER 3 / MINIMIZING SEDIMENT / 37

TABLE 3.1.  Effect of rotes of applied wheat straw mulch on runoff, infiltra-
           tion, and soil loss from Wea silt loam with 5% slops.
(tons/ a)
Soil Loss
Source: Adapted from Mannering and Meyer (1963).
* Water applied at constant intensity of 2.5 inches per hour.

tion of runoff velocity and the resistance of the soil to the forces of
flowing water.
     Laboratory studies have  shown  that the amount of energy re-
quired to initiate runoff was a function of clod size (Moldenhauer and
Kemper, 1969).  Rough, cloddy surfaces enhanced water intake and
contributed to  surface  detention of water,  even after water intake
was reduced by pore sealing.  It was apparent that  large clods created
many steep micro-slopes.  Dispersed particles from soil peaks eroded
into depressions, leaving exposed areas still receptive to water.
     A vegetative cover or surface mulch is one of the most effective
means of  controlling runoff  and  erosion (Duley  and Miller, 1923;
Borst and Woodburn, 1942; Baver, 1956; McCalla and Army, 1961;
Smith  and Wischmeier,  1962).   Wheat straw  mulch applied on
freshly plowed land at a rate exceeding one ton per acre almost com-
pletely eliminated runoff from, and controlled erosion on, a 5% slope,
as shown in Table 3.1 (Mannering and Meyer, 1963).  Mulch on the
surface protected it  from raindrop impact energy, reducing detach-
ment of soil particles and surface sealing.  In so  doing, high water
intake rates were maintained. The effectiveness of mulch in main-
taining high intake  rates was correlated with the proportion of the
surface covered.  In addition, the mulch created barriers and  ob-
structions that apparently reduced flow velocity and carrying capacity
of runoff.  This was evident especially at the '/r  and  i/o-ton mulch
applications where total runoff was 87 and 56% , respectively, of the
zero mulch treatment.  In  contrast, soil loss was 27 and 11% , respec-
tively, of the zero rate.
     In  another  study (Meyer  and Mannering, 1968), runoff  velocity
was measured as a function of mulch rate.  Five inches of simulated
rain were applied at a  constant intensity of 2.5 inches per  hour to
soil treated with straw mulch  at various rates.  Data shown in Table
3.2 indicate that small amounts of surface mulch caused considerable
reduction  in flow velocity.  Moreover, large reductions in erosion rates
were associated with relatively small reductions in  flow velocity. This
was not unexpected because the  quantity of material  moved is con-
sidered proportional  to about the fourth power of velocity.
    In a laboratory study, Kramer and  Meyer (1968) studied the effects


          TABLE 3.2.   Effect of applied wheat straw mulch on run-
                     off velocity, and soil loss from Wea silt loam
                     with 5% slope.
Soil Loss
         Source:  Adapted from Meyer and Mannering (1968).
         * After application of about 5  inches of rainfall when
         runoff rates were essentially constant.

 of mulch rate, slope steepness, and slope length on soil loss and run-
 off velocity.  Using a glass  bead bed to simulate a soil slope, they
 showed that less than a ton of mulch on the surface reduced erosion
 on slopes greater than 70 feet long at 4%  slope. Mulch rates of less
 than 1 ton reduced erosion from moderate to steep slopes (4 to 6%).
 However, on slopes of 8 and 10% , J/s- and  i/^-ton mulch rates did not
 greatly decrease erosion compared to no mulch. Erosion more than
 doubled as slopes increased from  8 to 10%.  Again, mulch rates
 ]/i ton  or greater reduced runoff velocity considerably.  It was noted
 that for some conditions low mulch rates increased erosion as com-
 pared to no mulch.  This was attributed to increased flow velocity
 and turbulence around mulch pieces, causing particle movement.
     In  some area soil wettability is  considered  a factor in  soil
 credibility.  Water repellency,  often developed as a result of fires on
 some soils, can cause much sediment production by curtailing infil-
 tration and encouraging runoff. Reduction in erosion is effected by
 modifying the wetting characteristics of hydrophobic  soil.  By  me-
 chanical or chemical means, soil wettability can be increased so that
 infiltration  rate is increased (Osborn and  Pelishek,  1964; De Bano,
     Another  means  of preventing runoff and increasing total infil-
 tration is through surface storage. Rough soil surfaces can retain
 several more  inches  of rainfall than  smooth surfaces,  due  to water
 being trapped in  the depressions of the rough  topography (Larson,
 1964).  Available subsurface storage  capacity has also been recog-
 nized (Holtan, 1965) to be important in the infiltration process. Thus,
 for soils to have high infiltration capabilities, they must have a high
 inherent permeability to water, show resistance to crusting, and have
 a high surface and subsurface storage  capacity.


     Practices or structures for erosion control are designed to do  one
 or more of  the following: (1) dissipate raindrop impact forces, (2)
 reduce quantity of runoff, (3) reduce runoff velocity, and (4) manipu-
late soils to enhance the resistance to erosion (Meyer and Mannering,

                                  CHAPTER 3 / MINIMIZING SEDIMENT / 39


     The relationship between tiDage methods and soil  erosion has
been reported by many investigators. Principles involved have been
well  documented (Larson, 1964;  Mannering  and Burwell,  1968).
Some tillage  methods deter soil erosion  by creating rough surfaces
which provide surface storage, reduce runoff,  and delay or prevent
surface crusting. Other tillage methods provide increased subsurface
storage,  and still others provide both. There are tillage methods that
leave all or part of the residue from previous crops on or near the soil
surface,  protecting the surface  from raindrop forces and enhancing
water infiltration.  Excessive tillage  can  be a factor in soil erosion,
however, because tillage is a source  of energy for breaking soil into
erodible  sizes just as are rainfall and  runoff. Tillage-induced soil con-
ditions play  a significant role in soil erosion through effects  on the
infiltration capabilities of  soil  (Burwell et  al,  1966; Burwell et al.
     On  a silt loam soil, 6.7 inches of simulated rainfall, applied at a
constant intensity of 5 inches per hour, infiltrated a surface created
by  moldboard plowing before  runoff began.  When  the  soil  was
plowed, disked, and harrowed, only 2.1 inches of water infiltrated be-
fore initiation of runoff.  Comparable values for unfilled  and rotary
tilled soil were 0.4 and 0.9 inch, respectively.  Cumulative water in-
take was fifteen times greater on rough, plowed soil and three times
greater on plowed, disked, and harrowed soil than on untilled  soil.
These differences were related to plow layer porosity and to surface
roughness (Burwell et al., 1966).
     Another  study conducted on the same soil compared infiltration
of simulated  rainfall of mulch-tilled  and clean-tilled  surfaces (Bur-
well et al., 1968).  The soil was previously cropped to oats.  Mulch
tillage consisted of  a pass with a chisel-type cultivator to a depth of 6
inches.   This tillage  operation  incorporated  about half  of the oat
stubble residue, leaving about 0.6 ton per acre on the surface. Clean
tillage consisted of moldboard  plowing in the fall, with and without
secondary disking and harrowing  the following spring,  and  spring
plowing  alone. Table 3.3  is a summary of this study.  Fall mulch-
tilled surfaces provided nearly eight times greater infiltration capacity
        TABLE 3.3   Influence of tillage treatment on water infiltra-

                Tillage                      Infiltration


Disk, harrow
To initial
During 2"
        Source:  Adapted from  Burwell. Stoneker, and  Nelson


  before runoff started and four times greater infiltration capacity dur-
  ing runoff than did fall-plowed surfaces, disked and harrowed in the
  spring.  Infiltration for fall  mulch-tilled  surfaces  was  more than
  three times greater than for spring-plowed surfaces.  Fall-plowed sur-
  faces were  altered by fall to spring weathering, resulting in little, if
  any,  infiltration  advantage  over  fall-plowed,  spring-disked,  and
  harrowed surfaces. Rainfall  action,  wetting-drying,  and  freezing-
  thawing cycles between fall plowing and spring planting act to dis-
  perse soil material which seals the surfaces by filling in depressions
  and open channels created by plowing.
      These  representative data indicate that the amount  of water
  entering soil can  be controlled significantly by soil physical condi-
  tions  created by tillage operations. Conventional tillage (plow, disk,
 harrow)  usually creates  conditions that restrict  water movement.
 Mulch and other so-called minimum tillage systems can produce soil
 conditions conducive to water intake.  Plowing, followed by disking
 and harrowing, usually leaves the soil clean or void of crop residue.
 Rain falling on these bare or  only partially covered surfaces washes
 fine soil into depressions and open channels, resulting in progressive
 soil sealing.  Rate of sealing  depends  on how cloddy or how rough
 the surface is after tillage. Where clean tillage is practiced,  it should
 create rough, cloddy surfaces that resist dispersion  and subsequent
 surface sealing so  as to delay the first runoff event during the spring.
     In a recent summary (Burwell and Larson, 1969) it was shown
 that prior to initial runoff, tillage-induced roughness accounted for
 most of the variation in infiltration, whereas differences in pore space
 caused only minor variations. In contrast, during a  2-inch runoff
 period, water intake was little affected  by roughness  or porosity—
 indicating that surface seals were already formed when runoff started,
 and overshadowed  roughness or porosity changes induced by tillage.
     Mulch  tillage—a tillage  system that  loosens the  soil without
 soil inversion—leaves all  or most  crop residue on the  soil surface.
 This creates  a condition highly resistant to raindrop and runoff forces.
 A comparison of runoff and soil loss from conventional and mulch
 tillage is  typified in Table 3.4.  In each instance the benefits of this
 type of tillage are apparent.
     Deep tillage or subsoiling  of some  soils can reduce soil losses by
 increasing volume  of  subsurface storage  available for  infiltrated
 water.  If deep tillage shatters  or fractures  a soil pan, this increased
 storage may be much greater than indicated by the increased depth
 of tillage. However, subsoiling generally has not been effective unless
 channels  were kept open  to the soil surface.  If subsequent tillage
 obliterates subsoiler slots in the surface few inches, little difference
in soil loss or infiltration can be expected (Meyer and Mannering,
     Postplanting tillage is used with most  tillage systems.  If a sur-
face seal  has developed,  cultivation  to break it may materially in-
crease water intake. In a 5-year tillage study (Mannering et al., 1966),
cultivation of minimum  tilled treatments reduced average runoff
from 3.5 to 2.1 inches and soil loss from 16.3 to 9.5 tons per acre as
compared to  the same treatments uncultivated. Under  some condi-

                                  CHAPTER 3 / MINIMIZING SEDIMENT / 41

TABLE 3.4.   EfFect of mulch tillage on runoff and soil losses in the Corn Belt.
   Soil,  and Slope
  Miami si, 6%

  Miami si, 9%

  Fayette si, 16%

  Muskingum si,

  Russell si, 5%
                                                  (inches)  (tons/a)
Noncontoured   Conventional
Contoured      Conventional


Noncontoured   Minimum


Source: Adapted from Mannering and Burwell (1968).

tions,  cultivation of rough, cloddy surfaces may increase credibility
by decreasing soil aggregate size, decreasing surface roughness, and
reducing existing crop residue surface cover.

     Contour planting and tillage function to control runoff and soil
loss from storms that are moderate in extent, or until capacity of soil
to hold or to conduct runoff is exceeded.  In field practice, rows are
oriented on the contour, generally with a slight grade toward a water-
way. On slopes  of moderate steepness and length, average annual
soil loss can be reduced by about 50% (Smith and Wischmeier, 1962).
Runoff is ponded and flows slowly around the slope rather than down-
slope.  However, when  smooth tillage is  used, or  when infiltration
rates are low, runoff from high intensity rains may overtop rows, re-
ducing runoff  and erosion effectiveness.  In addition, because con-
touring generally results in point rows and irregular field shapes, its
use as an erosion control practice is  declining.  Large farming equip-
ment and narrow rows are not compatible with point row farming.
     Contour strip-cropping is the  practice of alternating strips of a
close-growing meadow or grass crop with strips of grain or row crops
across a hillside. The erosion control aspect of strip-cropping is the
reduction in length of slope of land in row crop.  In  addition, flow
velocity  of runoff  water is reduced  as it moves through  the  close-
growing  grass strip, causing sediments to drop out. The sod literally
acts as a filter strip. The  reduction in  soil erosion from a strip-cropped
slope is proportional to the fraction of the  slope that is  in grass strips
(Wischmeier and Smith, 1965).
     Terracing is one of the oldest practices used to control erosion.
Terraces are combinations of ridges and channels laid  out  across the


 slope to trap water running downslope, and to conduct the water to
 suitable surface or subsurface outlets at a nonerosive velocity.  The
 primary benefit of terracing is the reduction in slope length. Since
 erosion is  approximately proportional to  the  square root of slope
 length (Smith and Wischmeier, 1962), reducing slope length in half
 can reduce erosion by more than 20% . Bench-type terraces also pro-
 vide for a reduction in slope steepness. Terracing with contour farm-
 ing is generally considered more effective as an erosion control prac-
 tice than  strip-cropping, but it is also more expensive.  With  both
 practices soil loss  is  confined within field boundaries.  In strip-
 cropping the saved soil from one storm event is deposited in the sod
 strip and  can be transported  further  downslope during subsequent
 storms. With terracing, the deposition is in  the terrace channel which
 offers positive sediment retention, unless overtopping occurs.
      Although effective for erosion control,  conventional broad-based
 terrace systems are not compatible with efficient tillage operations  or
 modern farm equipment. In  addition, herbicides are making it in-
 creasingly  difficult  to  maintain grassed waterways.  To  overcome
 these problems, a system of bench terraces with permanently vege-
 tated  backslopes  is  gaining  popularity (Jacobson,  1966).  In  this
 system, all  runoff is collected in low spots in the terrace channel and
 if necessary removed through  underground tile outlets, thus grassed
 waterways.  Parallel terraces  materially straighten field  alignment
 and eliminate objectionable point rows. In time sediment deposited  in
 the channel reduces the slope in the terrace intervals.
     Studies on instrumented  watersheds in  western Iowa  on deep
 loess soil indicate that although terracing did not affect total water
 yield from  a  watershed, the surface flow component of water yield
 was significantly reduced.  Only 14%  of water yield from  terraced
 watersheds  was surface flow, while on unterraced but contour-farmed
 watersheds, surface flow accounted for 64%  of water yield (Saxton
 and Sportier, 1968).  These differences in surface flow were associated
 to  sediment yield  from  these  watersheds  as  shown  in  Table 3.5
 (Piest and Spomer, 1968).

     Slope modification  measures  combined with  soil-conserving
tillage practices can be effective in reducing soil erosion from cropped
land.  However, to become  widely accepted, such practices  must fit
efficient  farming operations and must be economically feasible.  If
presently available practices do  not meet  these requirements, new
practices or systems that will control erosion and sediment produc-
tion without loss of net income to the operator must be developed.
     For example, consider a system where sheet erosion is controlled
through  till-plant tillage, and runoff is controlled by storage fills con-
structed  across  waterways  (Jacobson,  1969).  The fills,  like bench
terraces, would have favorable uphill slopes with a seeded backslope.
Water would be removed from fills  by tile  outlets. It is  anticipated

                                 CHAPTER 3 / MINIMIZING SEDIMENT / 43

TABLi 3.5.  Effect  of land treatment on sediment yield  of watersheds in
           western Iowa.

                                                   Sediment Yield
Watershed  Size        Crop      Land Treatment   1964  1985 1966



Source:  Adapted from Piest and Spomer (1968).

that such a system would almost eliminate soil loss  from cropped
fields on slopes up to 6%. Soil-moved sheet erosion is stored in the
fills and eventually helps reduce slopes.  Again, troublesome hillside
waterways are eliminated. Straight row farming is possible, adding to
farm adaptability. And the cost of such a system should be relatively
low.  Tillage costs will be lower, and building the system of storage
fills often would be less costly  than building waterways.  On lands
with slopes steeper than 6% , farming becomes progressively difficult.
Unless the slope can be reduced to permit more efficient machinery
operation, economics will force the retirement of much of these lands
from row-cropping (Jacobson,  1969).  Erosion  control on such land
will require bench terraces with tile outlets.
     To reiterate, nearly all sediment is  the result of man's removal
or disturbance  of natural soil cover of trees and grass.  Since all land
cannot be returned to its original cover,  wise land use planning and
careful use and treatment of land can reduce soil erosion, the source
of sediment. Although  the mechanics of the erosion process are not
completely understood,  guidelines have been developed, satisfactorily
tested, and  translated into erosion control practices, measures,  and
structures. Existing erosion control  technology has not been univer-
sally  accepted  and used,  primarily because  of direct or  indirect
economic considerations (Swanson and MacCallum,  1969). The chal-
lenge to  agriculturists, conservationists, engineers, and economists is
to continue their efforts to develop an improved erosion control tech-
nology that will  be compatible with modern requirements  and  eco-
nomically feasible. Only when this  challenge is met will there be a
significant reduction in sediments redrived from agricultural lands.

Baver L. D. 1956.  Soil phijsics. 3rd ed. New York: John Wiley.
Borst,'H. L., and Woodburn, R.  1942. Effect of mulches and surface
     conditions  on the water relations  and erosion of Mulkingum
     soil. USDA Tech. Bull.  825.
Burwell, R. E., and Larson, W. E. 1969. Infiltration as influenced by
     tillage-induced random roughness and pore space. Soil ScL Soc.
     Am. Proc. 33:449-52.


 Burwell, R. E., Allmaras, R. R., and Sloneker, L. L. 1966. Structural
      alteration of soil surfaces by tillage and rainfall. J. Soil Water
      Conserv. 21:61-63.
 Burwell, R. E., Sloneker, L. L., and Nelson, W. W.  1968. Tillage in-
      fluences water intake. /. Soil Water Conserv. 23:185-88.
 De Bano, L. F.  1969.  Water repellent soils. Agr. Sci. Rev. 7(2): 11-
 Duley, F. L.  1939. Surf ace factors affecting rate of intake of water by
      soils.  Soil Sci. Soc. Am. Proc. 4:60-64.
 Duley, F. L., and Miller, M. F.  1923.  Erosion and surface runoff un-
      der different soil conditions. Mo. Agr. Exp. Res. Sta. Bull. 63.
 Ellison, W. D. 1947.  Erosion studies, Parts I,  II, and III. Agr. Eng.
      28:145-46, 197-201, 245-48.
 Holtan, H. N.  1965.  A  model for computing watershed retention
      from  soil parameters. /.  Soil Water Conserv. 20:91-94.
 Jacobson, P.  1965.  Gully control in Iowa.  In Proc. Fed. Inter-agency
     Sedimentation Conf.  1963, pp  111-14. USDA Misc. Publ. 970.'
 	. 1966. New developments in land terrace systems. Am.  Soc.
     Agr. Engrs. Trans. 9:576-77.
       1969.   Soil erosion control practices  in perspective.  J.  Soil
     Water Conserv. 24:123-26.
 Kramer, L. A., and Meyer,  L. D.  1968.  Small amounts of surface
     mulch reduce erosion  and runoff  velocity.  Paper 68—206 pre-
     sented at meeting of Am. Soc. Agr. Engrs., 18-21 June  1968,
     Logan, Utah.
 Larson, W. E. 1964.  Soil parameters for evaluating tillage needs and
     operations.  Soil  Sci.  Soc. Am.Proc. 28:119-22.
 McCalia,  T.  M., and  Army, T. J.  1961. Stubble mulch  farming.
     Advan. Agron. 13:125-97.
 Mannering, J. V., and Burwell,  R. E.  1968. Tillage  methods  to re-
     duce runoff and erosion in the Corn Belt.  USDA Information
     Bull. 330.
 Mannering, J. V., and Meyer, L. D.  1963. Effects of various rates of
     surface  mulch on infiltration and erosion.  Soil Sci. Soc. Am.
     Proc.  27:84-86.
 Mannering, J. V., Meyer, L. D., and Johnson,  C. B.  1966. Infiltra-
     tion and erosion as affected by minimum tillage for corn (Zea
     mays L.). Soil Sci. Soc.  Am. Proc. 30:101-4.
 Meyer, L. D., and Mannering, J. V. 1968. Tillage and land modifica-
     tion for water erosion  control.  In  Tillage for greater crop pro-
     duction, pp. 58-62.  St.  Joseph, Mich.: Am. Soc. Agr. Engrs.
 Moldenhauer, W. C.,  and  Kemper, W. D.  1969. Interdependence of
     water drop energy and clod  size on infiltration and clod sta-
     bility. Soil Sci. Soc. Am. Proc. 33:297-301.
 Osborn, J.  F.. *nd Pelishek, R. E.  1964.  Soil xvettability as a factor
     in credibility. Soil Sci.  Soc. Am. Proc. 28:294-95.
Piest, R. F.,  and Spomer, R. G.  1968.   Sheet  and gully  erosion in
     the Missouri Valley loessial legion.  Trans. Am. Soc. Agr. Engrs.
Saxton, K.  E., and Spomer,  R. G.  1968.  Effects  of conservation on
     the  hydrology of loessial watersheds.   Trans.  Am. Soc. Agr.
     Engrs. 11:848, 849, 853.

                                 CHAPTER 3 / MINIMIZING SEDIMENT / 45

Smith, D.  D.,  and  Wischmeier,  W.  H.  1962.  Rainfall  erosion.
     Advan. Agron. 14:109-48.
Swanson, E. R.,  and MacCallum, D.  E.  1969.  Income effects  of
     rainfall erosion.  J.  Soil Water  Conscru. 24:56-59.
Wadleigh, C. H.  1968.  Wastes in relation to agriculture and for-
     estry.  USDA Misc. Publ. 1065.
Wischmeier, W. H., and Smith, D. D.  1965.  Predicting rainfall ero-
     sion losses from cropland east of  the Rocky Mountains. USDA
     Agricultural Handbook 282.



 G.  M. BROWNING, Leader
 H.  G.  HEINEMANN,  Reporter
     Browning: There are four or five things  that we want to do.
We  should consider what is known about soil  as  a pollutant.  A
good bit of that evidence was discussed at  the session yesterday
morning.  We should also consider where we are now and what we
know—and what additional  knowledge we need  to get to where we
want to be in the next  5 to  10 years.  A workshop such as this is
an excellent  way to identify and get  a consensus of what the im-
portant problems  are and to learn  how  we  might do something
about them.

     Verduin: I think the whole  picture is really long term—very
encouraging,  from what I have heard since I  got here.  Amemiya's
paper shows  that  if you have farmland so good that you want to
farm it but it has too much  slope, you can terrace it, and you won't
lose  much from it—any more than you did from grassland, which
we consider  pretty good soil-holding  land.   The whole thing,  of
course, is  tied to our problem of feeding the people  who need food,
but we have  been doing that with less  and less acreage.  It seems
to me that in the 20-year future,  we may well  have all farmland
that does not erode. We have a chance to get our erosion under con-
trol in this country and  set a model for the whole world.

     Browning: Does anyone want  to respond  to  that?

     Laflen:   We have had erosion control practices  since sometime
in the late eighteenth or nineteenth  century.  Our terracing pro-
gram started  out fairly strongly with the USDA in the thirties, and
today we still have between  5 and 8  million acres of cropland that
needs terracing. With the independent farmer, I don't see how we
are going to get the terracing done in the near future.

     Browning: I am concerned because I doubt that we are keep-
     G. M.  BROWNING  is Regional Director, North Central Agricultural
     Experiment Station Directors, Iowa State  University.  H. G. HEINE-
     MANN  is  Director,  North Central Watershed Research  Center,
     SWCRD, ARS, USDA, Columbia, Missouri.

                                  CHAPTER 4 / WORKSHOP SESSION / 47

 ing up with this. On a lot of the land that is in  tilled crops, with 8,
 10, and even 12%  slope, you can terrace and control this sediment.
 But modern-day 6- to  8-row  equipment  doesn't fit  on  irregularly
 shaped slopes very  well. So, if we are going to any more than keep
 up on those areas,  we need to devise new and  acceptable methods
 and procedures that will  help  control erosion.  We are not  in  the
 ball game economically with 2- or 4-row equipment when  compared
 with  the 6- and 8-row equipment of  farmers on flatter ground.  So
 we have some really tough going, admitting that we know what
 needs to be done and could do it. We have made a lot of progress
 in  the past 20  or 30 years, but we still have a whale of a long way
 to go.

     Herpich: I am wondering if the  voracious appetites  Americans
 have for meat  and  the  limited  acreage on  which we can put cattle
 to produce it will force us to retire a lot of this poor land to the pro-
 duction of crops that we  could use to produce  livestock.   It would
 probably help to solve the problem.

     Browning:  We all know, of course, that grass  is a wonderful
 conservation practice and there is a lot of land in crops that  should
 be  growing grass if we are going to control erosion from a  practical

     Verduin:  What you said is being  said in a number of situations.
 The pollution people are saying it. We have the technology but it
 is going to be expensive. You  look at the actual capacity of our so-
 ciety  and at the fact that  the farmer, over the  past 20  years, has
 practically  been held steady with a little  bit of  subsidy—and then
 we say we can't subsidize  him any more.  Why can't we? We  are
 subsidizing everyone else more.

     Herpich:  We live on an economy of waste. You have only to
 travel in some  of  the  European countries to understand what  I
 mean. We have great big, wide turn rows and waste land on water-
 ways.  If we are really concerned with producing food, we can put
 pipelines down the  waterways  and plant something on them.

     Kerr:  It bothers me that we deal so much in the ideologies of
 what  we are trying to  do  and spend  so little time in how  to do  it.
 We are talking about the sediment deposition and  the ramifications
 that it is going to have over a period of time—and we know that this
 is very serious.  But what are we going to do about it? If we take  a
 good  look at history, we learn  that  the Soil  Conservation Service
 has  accomplished  some water  conservation  and  some   pollution
 abatement  as an incidental thing to what  it is really trying to  ac-
 complish.  Its prime objective when it was formed was to  save  the
 soil. This needed to be done, but since then we have evolved through
 a couple of steps with  the Soil  Conservation Service in  its small
watershed program.  I think most people agree that a small water-
 shed program is the best  tool for keeping soil and  water  where it


 is supposed to be. I can't see how we are going to accomplish sedi-
 ment control for pollution purposes any  more effectively  than we
 could with the principles of the small watershed project. So I think
 that the SCS should be  permitted to accept benefits accruable from
 pollution abatement.  It cannot do that at this time.

      Cochran: I like to  hear that, because in Iowa we  do have  the
 tools. If we could get the Iowa Legislature to pass a bill, we could
 do the job that  everyone  has been talking about.  Back in 1965  I
 had what turned out to be the grand opportunity of being a member
 and becoming the chairman of a  committee in Iowa to revise  the
 drainage codes,  which  were 50 years  out of  date.  We  began to
 realize that drainage was related to flood  control, soil conservation,
 water pollution, recreation, and all the things that go hand in hand
 with  controlling  soil erosion  and  water pollution.  And  so  we
 launched a  program and  developed  the  Ccnservacy District  Bill,
 which is an enlargement of the watershed program.  We presented
 this to the Legislature in 1969.  We divided Iowa into six conservacy
 districts; so,  for  example, we are now in the Des  Moines River
 Conservacy District and any water that falls on Iowa soil that even-
 tually gets  to the Des Moines River is  in that conservacy district.
 According to our Conservacy District  Bill, we  start  in the upper
 reaches  of  our  various  tributaries and begin to solve  the  erosion
 problem on the individual pieces of land, then work  downward on
 the tributaries to  our major streams. The farmer is responsible  for
 setting up  soil  conservation practices found in any  conservation
 handbook.  The Conservacy District is responsible for internal  im-
 provements, flood control,  dams, and other structures. We have set
 goals to be met by voluntary action.  If it is  not voluntary, we will
 set up rules and regulations which will have to be abided by.   We
 can control pollution with this method.  In Iowa, we  have the bill,
 and now we have to convince the Legislature.

     Jones:  My question is:  What  do we need to know?  We would
 like to know what is a tolerable level for sediment as it moves off  the
 land. We have the universal soil loss equation.  The Soils people  tell
 us that we can lose so many tons per acre per year from our soils and
 still maintain production.   Yet, if  we look at  some sedimentation
 surveys of some lakes where we have public water supply and work
 this information back, we will find that  the lakes are  accumulating
 something like a quarter of a ton per acre per year from these water-
 sheds, and the people are up in arms at the rate at which they  are
losing their water supply  from sedimentation.  This  quarter of a
 ton per year is a far cry  from the 3 to 4 tons  per acre per year that
our Soils people say is a satisfactory level for controlling soil loss.  So
I think we need to reevaluate what is a reasonable level of soil loss
from our fields.

     Morris:  I would like  to address the  Representative from  the
Legislature about  the  problem of  air pollution.  I  think it  may be
wise to include an air study which is being emphasized right now
by large  societies.

                                  CHAPTER 4 / V/ORKSHOP SESSION / 49

     Cochran:  In 1967 we did pass an air pollution bill set up by a
 State Air Pollution  Commission;  however,  our later  bill  concerns
 soil erosion and  water pollution from the standpoint  of water  and
 wind erosion.

     Culbertson:  I was happy to  hear Mr. Jones  say  what he  did.
 Very shortly we  are going to have to set standards for suspended
 sediment in  streams.  This will be extremely  difficult.  These  rivers
 have adjusted to a certain base—depending on slopes, widths, depths,
 etc., and if we cut off the  sediment, we will  have tremendous bank
 scours, bed scours, etc.  Now, our problem will be to determine from
 past records what the sediment supply should  be in a river.  We don't
 want the situation that occurs below dams.  So I think you have to
 start in the erosion phase on the land and see what you can allow to
 go into your streams and then those of us who work in the rivers (the
 transport phase) will have to take it from there.  Now you may have
 the erosion problem solved, but we certainly do not have the  transport
 problem solved. And this relates to what Mr. Kerr said. If we  had
 some standards for sediment concentration, if it were exceeded,  this
 could be considered pollution or damage.  Maybe this is the first step,
 before including it in the damage benefit ratio.

     Verduin:  You mean, for example, that  the clearing up of the
 water in the Missouri  is causing injury  to all those dams because
 there is not enough silt in the water?

     Culbertson:  Definitely. Certain types of  river  beds will readjust
 the entire regime of the system downstream to a delicate balance
 between sediment load, water discharge  and velocity, and stream
 width and depth.

     Verduin:  Yes, but before the dams the Missouri changed its
 course every year. The Missouri has been doing that since the glacier
 pushed  down.  I still think that the best  silt load is as near zero as
 you can  get.

     Culbertson:  This is unnatural.

     Verduin: I am not sure of that.  When  a country is well vege-
 tated, there is  not much silt in the water. The streams are clean.
 I would  say  that the base  of  most of the rivers in this country is
 pretty low in silt.

     Herpich: From what vantage point are  we defining this word
 "pollution"?  It seems  that we must establish this before we  can
 say this is or is not pollutant.

     Browning:  What  about  this? Does this mean that  we  have
 more than one goal?  When the erosion factor was developed, we
 had  in mind what you might lose from the  soil so that you  could
maintain production. It seems to me that you  probably have different
levels for different things you are trying to do.


      Kerr:  I wonder if we are using what we already know.  I think
 we have acquired a lot of knowledge.  What bothers me is that we
 don't get more action.  Why don't we get the tools that we need, as
 Iowa is trying to with legislation?

      Herpich: I can't help but feel that the action program is a little
 behind the research.  I think we need  something positive to start to
 catch up.

      Morris:  Some active research has already been done in which
 streams have had viscosity reduction, and when that happens, there
 has been a  higher velocity flow, and therefore a  change in stream
 beds. We at the University of Nevada are making  a hydrologic study
 of the Trucky River system from Tahoe to Clearming Lake.  I think
 that things like that are necessary to make the stream behave natur-
 ally over a large number of years.

     Broivning:  I have a letter from Roland E. Pine, Program  Co-
 ordinator, Water  Quality  and Environmental Programs, Washington
 State  Water Pollution Control Commission.  I will just read sections
 of the letter.

     Irrigation of agricultural lands is the largest, single consumptive use
     of  water, and the  resulting runoff from this  activity is one of the
     nation's major sources of water pollution.  The Yakima and Columbia
     River Basins in south-central Washington are devoted almost entirely
     to  agriculture and  comprise  one  of the  most extensively irrigated
     regions in the nation. Nearly the entire  summer flow in the lower 80
     to 90 miles of the Yakima River is composed of irrigation return water.

     The sediment load, and adsorbed nutrients and pesticides, is a signifi-
     cant contributor to water pollution problems associated with irrigation
    return flows. The control of irrigation return flows  and their associated
    contaminates must be  through proper land and  water management
    practices  which can and will reduce the quantity of such contaminates
    carried into the receiving water.
    It must be  shown that  such practices are highly economic practices,
    beneficial to the farmer and  to his neighbors with  respect to crop
    yield, quality of the crop and cost of production.

     Lafien:  The points that Pine made pertain to the failure of our
terrace practices to be  accepted. It is  very  difficult to  show to  the
farm owner  that conservation practices do  put  dollars in his pocket
within  a reasonable length of time.  What we need are conservation
practices  that will compete for his dollar as fertilizers do.

     Gattis:   Economic reason has brought about  more conservation
in  our  state  than  terraces have  in  the past 20 years.  The simple
reason is  that when the land was in row crops it had to be terraced.
It  got to the point where you couldn't make a living on those terraced
lands, so  those fields are now in pasture or in woodland.  With that,

                                  CHAPTER 4 / WORKSHOP SESSION / 51

we have done more to reduce the sediment stream than we have in
the previous 20 years.

     Shrader:  On this matter of the permissible rate  of erosion, I
think it is becoming increasingly clear that that is one of the  basic
things.  We have many fields with a loss of 20 tons per acre, because
it is in the farmer's economic interest to let that land erode.  He dis-
counts the future very heavily. His idea of what is permissible soil
loss is just another magnitude different from the person that is look-
ing at the reservoir problem.  We  obviously need some way to use
the land that will maximize the  return for the whole population for
the whole watershed.

     McGill:  Mr.  Culbertson has raised an entirely new thought to
me. Assuming that he is correct and assuming that desirable stand-
ards could be achieved or  set, what could you do about increasing
the suspension?

     Culbertson: It will increase by itself. If sediment is taken from
the river and there is an unlimited supply  in the  stream bed and
banks, the flowing water will pick up enough sediment to bring itself
back into balance and equilibrium.  We cannot stop it.  The only way
you can stop this entirely is to use a TVA-type network, where  the
backwater from one  dam goes right up to the foot of the upstream

     McGill:  Seventy percent of the terrain in my part of the state
lends itself well to the growing of forage crops  and pasture  land.
We have done wonders and we have hardly scratched the surface in
improving our pastures.  We can carry 2 to 3 head of cattle where
we have had only 1  before. In southern Iowa we would do well to
get the  tax off our breeding cattle—or reduce it—and  get this land
back into production of red meat.  That would take care of the ero-
sion problem.

     Amemiya: It has been mentioned about the detrimental effects
of decreasing the silt load in  the  streams, insofar as the stream
channel is concerned.  And it has been alluded to that there  is  an
equilibrium established in  these stream channels. But the  equilib-
rium is  changing, and as we alter  our land use patterns upstream
on the watersheds, we are going  to reduce the amount of water
getting into our streams, and the stream flow is going to be less and
more uniform; so we will  have to look for  a new  equilibrium.  I
can't see how we can assume that  the stream flow pattern  will re-
main the same when we put these upland treatments into effect.

     Morris; It might be possible to establish some kind of workable
equilibrium for each of the rivers that  are so investigated.

     Cochran: Does siltation  in water reduce its energy?  Is this


 why, if we take silt out of the water, we have more energy to cut the
 stream banks and stream bottoms?

     Culbertson: You can't destroy the energy.  Sediment uses some
 of the energy and it reduces the turbulence that does the cutting.
 The problem in erosion  and deposition, as  I see it, is not after it
 enters  the large  tributary or stream.  Sediment damages, as the
 farmer sees it, are those that are carried on within his land and in the
 immediate vicinity of the small channel. When we get to the large
 streams, I personally  wonder what the effect of additional erosion
 control would do to the stream system.

     Kirkham: In the past few years we have  thought of sediment
 as a pollutant because it carries nitrogen and phosphate and  other
 fertilizers and pesticides into the river.  It is these materials that the
 sanitary  engineers are concerned about, because they see  this tre-
 mendously excessive stuff over and above that that they find  from
 their waste disposal plants.

     Duncan:  I would like to ask Dr.  Shrader to comment on the
 recent geological development  of,  particularly, northern Iowa.  Ero-
 sion really isn't anything new.  We have had 600 or so  inches de-
 posited on the west  side of the state fairly recently.  It took quite a
 little wind to deposit that from somewhere  and there weren't very
 many people farming then.

     Shrader:  I think Duncan has answered  his question fairly well.
 I will try to put some of  this in perspective, as I see it. Yes, we've
 had about 100 feet of deposition in the western part of the  state, of
 loess piled up  there, blown out of  the Missouri bottom over  a period
 of  several thousand  years, stopping a few thousand years ago.  But,
 to keep our perspective, I think we have to accept the fact that we
 have grossly accelerated the rate of erosion  here in Iowa since the
 time of settlement.  We are eroding our landscape in tens of years, at
 a  rate that would ordinarily take  hundreds  or even thousands of
 years in certain parts of our landscape. Any way you look at it, some
 of  our lands are eroding  at an  astounding rate  and  the end is not
 yet in sight.

    Holt: I think Dr. Kirkham has a point  in  that  people  are now
 associating pollution and  siltation with nutrient enrichment.  How-
 ever, I am not convinced  that we have greater nutrient enrichment
 now than we  had before man came here.  We have more sediment
 going into our streams,  and this  sediment is carrying  nutrients.
 There is no question about this.  But the source of these nutrients
 is an unknown.

    Broiuning: Do you have some evidence  to prove this?

    Holt:  We have  some evidence that  grasslands are contributing
more nitrogen  and  phosphorus, generally, than  cultivated land, in

                                   CHAPTER 4 / WORKSHOP SESSION / 53

 terms of solubles and nutrients in the water supply.  We do  have
 evidence that the total amounts of nitrogen  and phosphorus coming
 off in the sediment  are greater.  We can't evaluate their availability,
 but so far evidence  is that they are not generally very available. The
 best  evidence we have is that  before man came,  the prairies  were
 supplying as many  nutrients or more than are presently being sup-

      Verduin:  Where does the  phosphorus  come  from  for  that
 prairie soil?

      Holt: There is  an abundant supply of native phosphorus in our
 soils, as Dr. Black pointed out.

      Browning:   We have talked about a good many problem areas;
 the  problems are undoubtedly  different in different states.  It  has
 been emphasized also that  we have a lot more knowledge than we
 are  using.  Little has been said  about very pressing  problems on
 which we need more research,  but we have  emphasized that we do
 need action.

      Cochran: I  have listened to a lot of experts the past few years—
 Dr. Morris from  the University of Iowa, for example. He didn't take
 us back to before the time of  civilization, but he did give us some
 results of the monitoring program  on various streams and  major
 rivers of Iowa, and  he pointed  out that the nitrogen content  in the
 water is  rising  each year—and rising rather rapidly—not too far
 from Des Moines. Dr. Morris  correlates it with the  ever-increasing
 use of fertilizer, and as that rate goes up, more and more nitrogen is
 getting into  our  streams  as we use more  cultivation and  less  con-

     Duncan: This  was an excellent study,  as  I recall it;  it  was a
 study of fertilizer applied on  a grass slope.  Measurement  of  the
 runoff showed a  very  high concentration—in fact,  a relatively high
 percentage—of the nutrients that had been  applied  ran off.  What
 the study did not indicate was  the probability of receiving precipita-
 tion in this amount.  It was a 21/4-inch rain in about an hour and a
half,  but this was not reported.  Certain kinds of monitoring will
 produce  the  same data, particularly out of small streams.

     Cochran: Dr. Kirkham raised an interesting question the first
day of the conference when he  asked, "Why  doesn't the government
just decide to allow  crops to be planted on land where erosion  isn't
a very serious problem and just  not  plant  crops  and not fertilize
heavily on the slopes where erosion takes place?"  I  thing  that is a
perfectly fair question.  Maybe  the government should reconsider its
policy on  this, if erosion and pollution are serious problems.

    Browning: This gets back, I think, to your real  basic question.
A  lot of this would  be solved if we would use the land the way it


 should be used.  We don't really have a countrywide land use policy.
 Your policy might say that any land having an excess of some pre-
 determined  slope isn't suitable for crops because it will wash away
 and fill streams, and if you  fertilize, it  is going to carry fertilizer
 with it  and contaminate water, etc., so that land can't be used for
 row crops.  We are a long way from that kind of regulation, but I
 think that when we realize our problems and what causes them, we
 will be  able to move people, and when they decide they want some-
 thing done we will get action. I think we are making real progress,
 but progress is slow.

     Morris:  Maybe it would be better if we had state  legislation,
 because the subdivision developers in the small watersheds of Sierra
 and the hills that you have out there are another erosion problem.

     Brmvning:  I  think the national, state, and  local governments
 should look  at this thing and try to develop procedures that would
 be as practical as possible.

     Kerr:  I don't think any one  of us wants the federal  govern-
 ment to dictate land use policy. But the thing that the United States
 does have and can use is an incentive based  on the power of the
 purse.  The  United States could  use  pollution as  an incentive  to
 assist in the control of siltation, because  the whole idea of pollution
 control is very, very popular nationwide at this time.  It seems that
 this is a very opportune time to get some of these things under a
 legislative umbrella.

     McGill: We  take the position that we own the land; we  hold
 title to it and we do about what we want to with it.  I think, how-
 ever, that if I am doing something  that affects my neighbors, maybe
 it is logical  that  I have  some compulsion about the way  I  use the
     There has been a bill in  the State  Senate the past two sessions
 that I have been  interested in. It would require anyone who shares
 in cost-sharing benefits under conservation practices to file an appli-
 cation for a farm plan at the local SCS office. Farm plans have been
 very beneficial and farmers have followed them voluntarily, in most

     Jones:  The  man from Nevada mentioned that we should have
 state legislation for land use plans.  I don't know about other states,
 but the basic law establishing Soil  Conservation  Districts in Illinois
 says that Soil and Water Conservation Districts shall establish land
 use plans.  Perhaps the  emphasis  from  the urban  population will
 put some pressure on these districts to move in this direction.

     Culbertson:  Last October a national  meeting of county officials
was held in Washington on the subject  of sediment control.  County
officials were urged to go back and try to set up  these types of con-
trols within  their counties. The counties in the Washington area

                                  CHAPTER 4 / WORKSHOP SESSION  / 55

 have actually had laws passed that require a builder who tears up
 the land to  erect a building (anything with a specified number of
 square feet or acres) to install sediment control measures to insure
 that the sediment that comes from that freshly dug ground would not
 go onto someone else's property.  This system appears to be working
 very well. The state highway departments are entering into  this in
 new construction. One or two counties started this and it moved from
 county to county.

     Herpich:  Was the real emphasis for this sociological, esthetic,
 or economic  in nature?

     Culbertson: I think it must be  a combination of  all factors.
 Economically, it is all tied together; it's like water and sediment—
 you can't separate them.

     Duncan: A year ago last spring, quite a bit of dust blew around
 Iowa and in Illinois and Minnesota.   We have increased  our soy-
 bean acreage from about 2Vo million acres to about 5i/>  million  the
 past 4 or 5 years and it will probably go up  another million.  Land
 following soybeans is easy to work on, but we need to find out why
 land following  soybeans  tends to blow  so much.  This  is an easy
 place to start.  I don't know why we don't work on some  of the easy
 problems instead of trying to work on the difficult research.

     Browning:  I am surprised that no one has talked about tillage,
 though some folks have  been  working on it.  It relates partly  to
 what Duncan said, because if you  leave that soybean straw on  the
 surface, it doesn't blow a heck of a lot; but we have to rake it in  the
 fall or plow  it under. Some work shows  that you can raise almost
 as many beans on land that you  disk or minimum till as when you
 plow.  I would  like  for someone to respond to what Duncan has

     Duncan: Iowa has an  absolute economic advantage producing
 soybeans in  the western part  of  the state.  Farmers  are finding
 this out; and as a result soybeans are moving in on this deeper soil.

     Hclt: Ten years ago  when I first went to western Minnesota to
 visit with the farmers out there,  they mentioned  two big problems.
 One was that they were going out of growing soybeans because it
 leaves  the land too loose and it is too difficult to control  weeds.  The
 other problem was controlling erosion on complex slopes. We have
 been working on tillage practices as an  approach to this. Generally
 speaking, some  form of multitillage is effective  on these  complex
 slopes, where other practices are not.  Back to  soybeans—I suspect
 that there may be something in the  microorganisms that  loosens this
 soil. We have been checking roots lately, and when a soybean plant
matures  and the leaves drop  off, the root system is completely gone—
so I  suspect there is some tie-in between the microorganisms  which
exist under a soybean crop and  this looseness and tendency to blow.


      Browning:  I figured this out 27 years ago and wrote a paper
 about it.  In the first place, soybeans have roots that grow a lot in
 about 6 weeks;  they use a lot of water  and then  dry out, and the
 wetting and drying and freezing and thawing have  a granulating
 effect.  The second answer is:  they have a lot of nodules in the roots,
 and that is quite a stimulant for microbiological activity—it is about
 40 times as  active in this area. We carried  on studies with corn and
 soybeans under identical conditions; if we left that residue on, it was
 loose and there was little runoff or erosion from the soybeans.

     Amemiya: We can, through tillage,  affect the infiltration  of
 rainwater into the  soil, and in so doing, cut down the runoff and
 soil loss problems.  Some of the work of my colleague Bill Molden-
 hauer indicates that corn following soybeans erodes much more than
 corn following corn.

     Moldenhauer: We had about 40% more water erosion from corn
 following  soybeans  than we  did from  corn following  corn.  We
 haven't measured wind erosion.

     Jores:  I think one of our primary needs in research right  now
 is good economics of soil conservation.  In Illinois we haven't had a
 good study since 1954.

     McGill:  Is it a reasonable assumption that the younger farmers
 are more susceptible to  these conservation  practices and new ideas
 and will voluntarily be more concerned with soil? Or will they look
 at the almighty dollar and go at it like our grandfathers  did  and
 plow the hills straight up and down, etc.?

     Jones:  I think they are more computer  based and look at the
 net return.  And I think they  are even  less prone than were their
 fathers  or their grandfathers  to use the  moral side as impetus for
 soil conservation.

     Holt:  There was a study in recent months in Illinois by Swan-
 son (Journal of Soils) on the economics of conservation work. Swan-
 son said that it doesn't pay, even up to 50 years' projection.

     Amemiya: I think this study was based on strip-cropping, crop
rotation, and contouring.  It didn't include some of the major slope
modification  practices that  we talked about.

    Verduin: We're getting hung up again as to whether or not  it
pays.  We should pay a man for terracing if we are convinced that
that's what should be done to hold that soil over the next 10 genera-

    Lafien:  There is a lot of competition for the tax dollar, and it is
going to  have  to be spread around  a  lot.  Where is  the money to

                                   CHAPTER 4 / WORKSHOP SESSION /  57

 come from?  It is going to cost 800 million dollars or more to terrace

     Verduin: That is the biggest problem, and I don't know whether
 any research has been done on it.  How can we get the thing going
 to do the things that almost anyone will admit need to be done?

     Wiersma: It's a matter of priorities  again, is it  not?   I think
 that if people who are going to pay for this are informed of the prob-
 lem in  the right manner, we will  get the priority and  I think we've
 got to talk in terms of "we" rather than "they."

     Browning:  How do you establish priorities on things?  In our
 process, we get people agitated and then we get things  done. So I
 think it behooves people in these areas to dig out the facts and present
 as well as they can alternative solutions  and what they're going to
 cost with and without our program.  Then the people can and  will
 decide.  I think we  are going to be forced to do more of this, and the
 better job we do of identifying the priorities and showing how the
 benefits will accrue against the cost, the better our chance will be of
 sharing in the short dollars that are available.

     Wiersma: How did they get the money to put this man on the
 moon?  They didn't do  that on economics, did they?  Haven't we a
 lot more concrete evidence in  agriculture than  they have?

     Browning: Agriculture has far more specific things to show  and
 put values on than practically any other area—and we have prob-
 ably done less of it than anyone.  We must do this.

     Herpich: Our  Congress has already established some priorities.
 They said "pollution."

     Browning: We must begin to establish some of these guidelines.
 We'll have some evidence but we  won't  have nearly  enough.  But
 we'd better get ourselves in that position or we won't get the support
 needed  to do the job that's in the best interest of the public.

     Kerr:  Right now, our gimmick is pollution. We're  in  a better
 position than we have ever been, if we  can just proceed correctly.

     Morris:  Why not use air pollution also to get at the urban pop-
 ulation, increasing  the  number of people  who might  be interested
 in pollution.  If we enlist the urban development  as well as the
 rural development, we might be able to make the package better than
 we could with the single increment, pollution.

     Broivning: This  is the key to obtaining support in these areas.
 Looking at legislation history, money comes in fairly large chunks—
usually  a good many years after somebody has been talking about it.


 You have to have a genuine problem that the public will recognize.
 The Congress will find the money, priority wise.  If we've done our
 jobs right, we'll get the kind of support we need.

     Kirkham: One of these  days, the American  public will  get so
 sick of that war in  Vietnam that it  will be stopped. Then  there
 should be some extra money,  and it's up to us to be ready for it.

     Gattis: For many years we've known what slopes should be to
 control erosion under  given  cropping systems.  Maybe we should
 work toward putting the slope on ground that will conserve the soil
 and keep it in place.  This is being done in some areas at a cost per
 acre that is less than the cost of terracing the land.  Your steep land
 here would cost  more,  but slope alterations would be something of
 a more permanent nature.

     Evans:  Sediment is a pollutant if man  thinks it is.  It is a re-
 source out of place.  Sediment can be useful, and it can be harmful.
 Economics is at the core of this.  We must find economical means of
 recovering our sediment  and  utilizing it.  We  need to take an en-
 vironmental  approach  to  pollution.   We  must enlist the help of
 ecologists.  This area has been neglected. We need to develop some
 type of land use plan that involves not only agriculture but also the
 urban  areas, and it's got to take  in transportation, manufacturing,
 agriculture—the  whole business.  We must  think in terms of the
 future and try,  as educators  and scientists, to get as many people
 as possible thinking in  this direction and trying to promote legisla-

     Amemiya:  Sediments are a problem, whether they're pollutants
 or not.  They are pollutants, not only  because of  their physical sig-
 nificance but because of their chemical significance. Sediment costs
 taxpayers money in maintaining irrigation ditches, canals  and estu-
 aries, ponds, lakes, and reservoirs.  I think that proper land use,
 especially on the upland areas, will go a long way toward minimizing
 the sediment that enters our streams.

    Pine:  If sediment interferes with another  beneficial use of
 water,  it is a pollutant.  When sediment gets into lakes, reservoirs,
 and other areas, it is a pollutant.

     Culbertson:  Many of you  realize there is a problem, but we need
 figures to work with.  The only way to get them is to collect  water
 and sediment data and evaluate the problem.  The  Geological Survey
is one federal agency that will put up half of the money and cooper-
 ate with any state agency or city, on a  fifty-fifty basis, to make these
investigations.  So I would suggest that if you  want definite values
 to attack a problem,  propose  this to my organization and we  can
make these evaluations.

                                   CHAPTER 4 / WORKSHOP  SESSION  / 59

     Verduin: If a river flowing into a lake contains  so much  silt
that it reduces the light penetration to a point where 1%  of surface
light doesn't reach the middle of the water, then silt is  a  real pollut-

     Manges:  I  think that within  5 years, society is going to tell us
sediment is a pollutant and we will be forced to do something about
it. We must be  prepared to suggest a program.

     Browning:  Thank you very much  for  your participation here
this  evening.  This is the end of the session.


       HE pollution of surface  waters by  domestic sewage and by
industrial waste has long been recognized as a serious  evil in  our
society.  Unfortunately,  action  to ameliorate these long-recognized
sources of water deterioration has lagged far behind the  recognition
of the  problem. But only in recent years has a more subtle problem
come to light.  It is the problem  of superabundant plant nutrients
in our surface waters.  These are  invisible; they do not show up in
the classic BOD test used  by sewage  treatment plant operators to
measure the efficiency of their treatment process.  And the presence
of these nutrients  is usually recognized  only after we see the nui-
sance  levels of aquatic  plant growth which they  support (Sawyer,
1947; Verduin, 1964, 1967, 1968,  1969).
     At first glance one would consider the addition of plant nutrients
to rivers and lakes as beneficial.  This is  especially true when  the
low level of plant nutrients in natural  waters is considered.   For
example, forest streams and lakes which receive no urban or agri-
cultural runoff have phosphorus concentrations of less than 7 Mg/liter
(parts  per billion)  and the waters of  our great Lake  Superior  are
mostly still in such condition.  Nitrogen also is scarce in such natural
waters, and one would imagine that a bit of enrichment would be
    However, to appreciate the  damage  done by even relatively
small quantities of fertilizer we must examine the  modus -Vivendi of
the aquatic community.  It has evolved,  of  course,  under  the  low
level of nutrients described above. It consists of  three major com-
ponents: (1) the microscopic autotrophic plants in  the water, (2) the
heterotrophic organisms in the water, and (3) the organisms  that
live on the bottom.  These form  a delicately balanced web of  life
which  can survive  only so long  as the  delicate balance is preserved.
There are of course many complex interactions in this community,
but as an example let us consider  the problem of oxygen  supply.
Oxygen has a low  solubility in  water (about 8 mg/liter  at  summer
     JACOB VERDUIN is Professor. Department of Botany, Southern Illinois


  temperatures), and it must be transported to the bottom by vertical
  mixing in the water column.  If we  add fertilizer to the lake, its
  plant population will increase, and the quantity of organic matter
  settling on the bottom will increase proportionately.  As it decays
  there, it consumes oxygen, but we have not increased  the vertical
  mixing rate, consequently the bottom  organisms  will be subjected
  to lower oxygen than ever before in their history—with catastrophic
  consequences.  Such a sequence  of events has been documented in
  Lake Erie. Dr. N. Wilson Britt (1955) has described the catastrophic
  extermination of the mayfly (Hexagenia) population in western Lake
  Erie in 1953.  This population was never reestablished, but a popula-
  tion of bloodworm larvae (Chironomus), more tolerant of low oxygen,
  has  displaced it.  Several other changes have  been documented in
  Lake Erie. Large numbers of dead clams have risen to  the surface
  during calm weather (the periods  of lowest mixing rates) and the
  dominant fish  species has changed from  walleye  (Stizostedion) to
  yellow perch (Perca).  There is little doubt that  all these changes are
  attributable primarily to enhanced  supplies of plant  nutrients.
      To focus  more specifically  on the levels of  plant nutrients,
  examine Table 5.1. It shows  what to an agriculturist  must  seem
  to be fantastically low  levels of plant  nutrients.  Hydroponic  solu-
  tions, for  example, are  made up in mg/liter instead  of the Mg/liter
  used in Table  5.1. But once the shock of this feature has been  over-
  come, there are two highly significant  features to notice:  (1) While
 nitrogen supplies increased by about 30% during the 20-odd years
 covered by the table, phosphorous supplies increased by 480%, and
 in doing so (2) the N/P ratio changed from a value  of  35 to 9.2.
 Even an elementary plant physiologist. will  recognize  that a  N/P
 ratio of 35 represents a medium in which phosphorus  is severely
 limiting,  but a ratio of 9  is a well-balanced medium because the
 ratio of N/P in protoplasm  is about 8.  Consequently the 20-year
 period covered in Table 5.1 was one  of greatly increased plant growth
 in Lake Erie as a response to nutrient enrichment, with phosphorus
 enrichment playing a  spectacular role.

     To  appreciate the extent  of the plant nutrient problem in  our
surface waters an examination of the present  phosphorus levels is
most revealing. Figure 5.1 presents such information for the years
1965-66. The data were  provided by the Federal Water Pollution
Control Administration.  Each figure on  the map is the average of 2
or more stations in the  vicinity of the number, with the exception
of the Sioux Falls, South Dakota, station (1,618 Mg/liter) which is in
a class by itself!  The data represent samples drawn  primarily from
drinkine water intakes.  But it is well known that for many cities
the drinking water intake draws samples of the  diluted sewage efflu-
ent from the city upstream. So it would not be unrealistic to regard
these numbers as representative  of the  diluted sewage effluents  of
our cities, plus the contributions from agricultural drainage.

                                 CHAPTER 5 / PHOSPHORUS IN WATER / 65

TABLE 5.1.   Comparison  of nitrogen and phosphorus data  in western Lake
            Erie for 1942 with data for 1965-66.
Available Nitrogen
NH,-N plus NO,-N
(ng/ liter)
(ng/ liter)
N/P Ratio
  Average of 28 samples, April through December.  Data of Chandler and
Weeks (1945).
f Average of 20 samples, June, July, August 1965, and March, April, May
1966.  Data of J.  Kishler (private communication).  Samples  analyzed by
the Great Lakes 111. River Basin Project Lab., Chicago.

     The degree of enrichment that our waters have experienced can
be  appreciated when we  compare the phosphorus levels in Figure
5.1 with those observable even today in streams of forested areas.
Sylvester (1961,  as  reported  by Mackenthun, 1965) reported soluble
phosphorus levels of 7 jug/liter for such streams. It seems likely that
our  prairie streams had  similar plant  nutrient  levels  before the
prairies were converted to  farmland. Therefore the aquatic communi-
ties  that  originally occupied  our  lakes and streams were adapted  to
such low nutrient levels. The data in Figure 5.1 reveal that such low
nutrient levels are found today only in the open areas of the Great
Lakes.  All of the major streams of the United States exhibit phos-
phorus levels five to thirty times higher  than this  "natural" plant
nutrient level.
                                        OPEN AREAS OF L SUPERIOR,--
                                                  &  HURON     ,  '
           *SIOUX FALLS  \
            SOUTH  DAKOTA
     FIG.  5.1.  Total  phosphorus  concentrations  (ortho-,  meta-,  and or-
     ganic)  in  water  supplies  of the  United  States.  Data  provided  by
     FWPCA, 1965-66, averages of 18 months' collections.


  TABLE 5.2.  Correlation  between metabolism  of  Cladophora  communities
            and  phosphorus concentrations in Lake  Erie.

                     „,   ,             Cladophora Community
   Location         Concentration    Photosynthesis      Respiration

                      (^g/liter)     (Mm CO2 absorbed   (um CO, evolved
                                        /9/hr)           /g/hr)
Lake Erie
Lake Erie






 * Average of 21 samples.  Unpublished data of B. A. Thumm, E. E. Klum,
 and D. Lentz, Dept.  of Chemistry, SUNY College at Fredonia.
 t Unpublished data, K. G. Wood, Dept. of Biology,  SUNY College at Fre-

     The influence  of such enrichment was recognized many years
 ago. More than 20 years ago a study of lake fertilization by tributary
 streams (Sawyer, 1947) revealed that nuisance blooms of algae arose
 when phosphorus concentrations exceeded  20  Atg/liter.  More  re-
 cently the  nuisance level of filamentous algal  growth (Cladophora
 glomerata) in the littoral zones of the lower Great Lakes has occa-
 sioned some investigation. Table  5.2 presents data (Verduin, 1968)
 showing a  correlation between the phosphorus supply and the meta-
 bolic rate of the Cladophora community in Lake Erie.  The western
 Lake Erie community, which is bathed in water having about 35 Mg
 of phosphorus per  liter,  exhibits a photosynthetic  rate  about 3.7
 times that of the eastern Lake Erie community (Dunkirk, New York),
 which is bathed in waters having about 10 Mg of phosphorus per liter.
 The respiration rate of the western Lake Erie Cladophora community
 is also  distinctly  higher than  that of the eastern Lake  Erie  com-
 munity.  These data  suggest that the metabolism of the Cladophora
 community increases almost  in  linear proportion to  increases in
 phosphorus supplies, within  the range  of values presently encoun-
 tered in the Great Lakes.
     The influence of enhanced plant nutrient levels is widespread
 today. They are responsible for foul tastes and odors in  our drink-
 ing water, clogging of water intake filters, and windrows of decaying
 algae on our beaches. They also result in oxygen depletion in deeper
 parts of  our lakes,  with catastrophic  destruction of fish and of
 bottom-dwelling organisms, and they support the weed-choked con-
 dition of shallow areas.  It is significant that the 1966 annual report
 of the Division of  Health and Safety of the Tennessee  Valley Author-
ity devotes  three  pages  to the problem  of  counteracting nuisance
 growths of aquatic plants in the large TVA reservoirs.  The following
quotation from  that report is pertinent:  "In  Cherokee Reservoir, in
June, the last of three surveys showed oxygen nearly depleted below
the thermocline. Analysis of survey results over the past  few years
revealed  earlier oxygen  depletion  each ensuing  year.  The Holston
River below Kingsport was found  to be highly eutrophic  .  . .  sup-

                                CHAPTER 5 / PHOSPHORUS IN WATER / 67

porting dense masses of aquatic weeds."  The report postulates that
organic contributions by the aquatic weeds may have influenced the
oxygen depletion in the hypolimnion. No monitoring of phytoplank-
ton crops in the epilimnion is mentioned.  It seems likely that en-
richment of the Holston River, which flows into Cherokee Reservoir,
has supported greater phytoplankton growth there, with  a resultant
increase in the organic  load settling into the hypolimnion. The TVA
report describes countermeasures taken to reduce nuisance growths
of aquatic plants. These include scouting of dense plant concentra-
tions  by aerial  surveillance,  and the helicopterized application  of
2, 4-D pellets. Thus we counteract the pollution due to plant nu-
trients by adding another chemical pollutant, a plant toxin!

     Up to this point we have been concerned primarily with estab-
lishing the significance of plant nutrients as water pollutants, and
with  an inspection of the level  of  a key plant nutrient,  namely
phosphorus.  Because this  symposium  is concerned about the role of
agriculture in clean  water, we can best advance that concern  by
trying to evaluate agriculture's contribution to the phosphorus levels
now prevailing in our surface waters.  Such evaluation can be made
by examining the concentrations in  streams whose watershed rep-
resents agricultural land and  does not include urban  runoff. Table
5.3 presents data of this kind compiled from  several sources. These

TABLE 5.3.  Soluble  phosphorus  concentrations  reported for waters  from
           agricultural watersheds.

      Author                      Watershed             Phosphorus
Engelbrecht and            Kaskaskia  River
  Morgan, 1960*              (111-)                           60
Sawyer,  1947*             Watershed,  farmlands
                             agricultural drainage
                             around Lake Mendota            48
Putnam  and Olson,         St. Louis and Black
  1960*                      rivers, tributaries
                             of western Lake Superior         40
Harlow,  1966t             Raisin River
                             (Mich.)                         60
Owen, 1965f               Ontario agricultural
                             watershed                      33
Hardy, 1966t              Big Muddy
                             (111.) river system,
                             upstream portions               110
                                               Average      58
* As reported by Mackenthun (1965).
t Private communication, plus papers presented at the 9th Conf. on Great
Lakes Res., Chicago, 1966.


  TABLE  5.4.   Comparison of upstream station phosphorus values (assumed to
             represent agricultural  runoff  and drainage) with values  at the
             mouth of the Big Muddy River (representing the combined con-
             tribution from agricultural and urban effluents).

  Four upstream stations, average of 4 samples
     at each station 	   110 /ig/liter
  River mouth station, average of 4 samples	   350 /ig/liter
  Percentage of total attributed to agriculture	     31%

  Source: Private communication.  Hardy supervised collection of samples.
  Chemical analyses  were performed  by the Great  Lakes  111. River  Basin
  Project Lab.  in Chicago.  Statistical analyses were made by Richard Rowe
  of the Southern 111. Univ. School of Technology.

  data show that runoff from rural watersheds represents a significant
  fraction  of  the  phosphorus  appearing in our  surface waters.  The
  data in Table 5.3  represent rural watersheds from the upper Missis-
  sippi and Great Lakes region, the  same region for which phosphorus
  levels of 175 Mg/liter are shown in Figure 5.1.  The average value of
  58 /ug/Hter in Table 5.3 suggests that  approximately one-third of the
 phosphorus  contribution  may come from  agricultural watersheds.
 Some data collected by George  Hardy, a sanitary engineer with the
 Illinois Department of Health during the summer of 1966, permit
 a similar evaluation  for  a single  river system,  the  Big  Muddy in
 southern Illinois.  Table 5.4  presents these data.  This analysis also
 indicates that about one-third of the phosphorus concentration found
 at the  mouth of the Big Muddy  may be attributed  to agricultural
 sources.  Upon consulting with fertilizer dealers in the  Big Muddy
 watershed,  it was  learned that farmers  apply  fertilizers  in  such
 quantity that the PO4 added amounts to about 100 Ib/acre. If 1% of
 this addition is dissolved in the  annual runoff and drainage it would
 create  a  phosphorus  concentration of about 50 Mg/liter, which is
 similar to the average value in  Table 5.3.  Obviously the farmer is
 not  going to be impressed by  the fact  that 1%  of his  phosphate ap-
 plication  is  lost in runoff and  drainage; neither do any practical
 measures for reducing this contribution come to  mind.
     A somewhat encouraging aspect of the  above analysis, at  least
 for the agriculturist,  is that  agriculture appears to  be responsible
 for less than half of the phosphorus supplies  in our waters.   The
 supply from urban sources seems  to represent the major fraction.
 In urban  sewage effluents, detergents seem to contribute about three
 times more phosphate than is contributed by the organic matter in
 sewage, according  to Engelbrecht and Morgan (Mackenthun, 1965).
 Consequently,  detergents would appear to be the most significant
 single  source  of phosphates  enriching our  waters  today.  Unfor-
 tunately their use is still increasing. I am told  that some weed  and
orchard sprays contain these detergents to prevent clogging of spray
nozzles.  And our northern cities  are  now  adding these detergents
to salt applications on icy streets.  The detergents presumably act
as rust inhibitors.

                                CHAPTER 5 / PHOSPHORUS IN WATER / 69


     There is a fairly recent development in agricultural practice that
is most distressing. For hundreds of years the successful  farmer has
been spreading animal manures and decayed vegetable composts on
his land.  But in recent years feedlot  operators have been installing
lagoons to decompose the  animal manures. These ancient, tried and
true soil improvers are now being digested in lagoons, and the plant
nutrients  that remain in solution after the digestion is completed are
discharged to our surface waters  in  the lagoon overflow!  Prepare
yourselves for a dogmatic statement:  Manure belongs on the land.
To  be  sure, the  agricultural economist may be able to demonstrate
that it is cheaper to buy chemical fertilizers than to spread manure.
But it  certainly  cannot  be a great deal  cheaper, and if our  Agricul-
tural Stabilization Service can pay the farmers for not raising feed
grains or  cotton, then it should certainly consider paying the feedlot
operator for not lagooning animal manure, because manure belongs
on  the land.  Manure contains many trace  elements, vitamins, soil
conditioners, etc., which  chemical fertilizers do  not provide, and
organic manures represent the only economically feasible source of
CO., fertilization. With our modern crop densities the COo content of
air among the leaves drops spectacularly, especially in quiet  weather
(Verduin and Loomis, 1944).  A healthy layer of  decaying  organic
matter will serve to augment the atmospheric CO2 supply  significant-
ly.  The method we use in  solving a problem is profoundly influenced
by the initial  conception of the problem.  If we regard a concentration
of animal manure as a disposal problem, we are likely to adopt the
least expensive means of disposal available, and the lagoon may well
fit the  bill. But  if we regard a concentration of animal manure as a
valuable source  of fertilizer and soil conditioner, the problem is one
of transportation and application—it is not a disposal problem  at all,
and a lagoon will never be considered a solution to the problem.

    Phosphorus is a key element in the problem of pollution by plant
nutrients because it is present in such low concentrations in natural
waters and because it  has undergone much more  spectacular in-
creases than any  other plant  nutrient.  The agricultural practices
which would tend to reduce the agricultural phosphorus contribu-
tion are simply sound  soil  conservation practices: (1) methods of
cultivation which minimize runoff,  (2) insuring intimate mixture of
fertilizer with soil, (3) improving soil texture by addition of  animal
manures and  ploughing  in legume  stands,  and (4)  particularly,
abandoning the practice of lagooning animal manures.  But the fact
that the major  phosphorus contributions come from nonagricultural


 sources shifts the burden of problem  solution elsewhere.  However,
 the solution is of interest to the agriculturist as well as to any other
 citizen, and more so because agriculture can make distinct contribu-
 tions to the solution of this problem.
      The contribution  of agriculture to the solution of the problem
 has two facets:  (1) The same condemnation made above of the de-
 composition of animal manures in a lagoon applies to the decomposi-
 tion  of organic matter in urban  sewage treatment plants.  This or-
 ganic  matter, considerably enhanced  today  by the use of garbage
 disposal units in our  kitchens, is as  good for  the land  as animal
 manure, and it should be so utilized.  Again, if we stop thinking of
 the problem as a waste disposal problem and look at it as a problem
 in processing, transporting, and applying a valuable soil conditioner,
 our attack on it will  be drastically  altered. Because the organic
 matter  includes  human  excrement  we have  a  pasteurization
 problem, but the fact that the organic matter is suspended in water
 introduces a drying  problem. If the drying is  done at sufficiently
 high  temperature, pasteurization will  be automatic.  Moreover,  we
 may  well be able to utilize the "waste"  heat from thermal electric
 power  plants—heat which is now being widely decried as thermal
 pollution! Once the  organic  matter is  pasteurized  and  dried  we
 should call in the chemical  fertilizer  industry  to add  as much of
 their product as is needed to provide a maximally advantageous fer-
 tilizer.  Then we should pelletize this product so the farmer can dis-
 pense it from attachments on his plough, disc, drill, and planter,
 thus  avoiding extra trips over the landscape.
     (2) Even after  all settleable  solids are removed  from the sew-
 age,  a  high  level of dissolved organic matter  and plant  nutrients,
 especially detergent phosphorus, will remain  in  the urban effluents.
 The most promising method of treating such effluents is  again an
 agriculture-related treatment.  It  is the "living  filter" described by
 Kardos in the AAAS symposium on Agriculture and  the Quality of
 our Environment (Kardos,  1967).  If such  sewage  effluents are
 allowed to percolate  through the  root zone of crop plants or trees,
 the dissolved materials are removed effectively and diverted to pro-
 mote  valuable plant growth. The water emerging from tiles beneath
 these  root zones can be released to our surface waters without fear of
 serious pollution.
    It is obvious that agriculture has a primary role to play in the
 solution of  the  pollution problem.  Where it is  contributing plant
 nutrients directly, it should attempt to  minimize such contributions,
 but wherever plant nutrients are  entering our surface waters from
 nonagricultural sources, we should recognize  the agricultural poten-
 tial of such plant nutrient sources and attack the problem  of restor-
ing them to the land.  In the problem of removing concentrated nu-
 trients from  water, agricultural technology can make a major con-
 tribution in the application of the living root zone filter to the process
of plant nutrient removal.

                               CHAPTER 5 / PHOSPHORUS IN WATER / 71


Britt,  N.  W.  1955. Stratification in western Lake Erie in summer
    1953:  effects on the Hexagenia (Ephemeroptera) population.
    Ecology 36:239-44.
Chandler, D.  C., and Weeks, O. B.  1945.  Limnological  studies  of
    western Lake Erie. V. Relation of limnological and meteorolog-
    ical  conditions to the production  of phytoplankton in  1942.
    Ecol. Monographs 15:435-56.
Engelbrecht, R. S., and Morgan, J. J.  1959.  Studies on  the occur-
    rence  and  degradation  of condensed  phosphate in  surface
    waters. Sewage Ind. Wastes 31:458-78.
Kardos, Louis T.  1967. Waste water renovation by the land—a liv-
    ing filter. In Agriculture and the quality of our environment, ed.
    Nyle C. Brady, pp. 241-50.  Norwood, Mass.:  Plimpton Press.
Mackenthun,  K. M. 1965. Nitrogen and phosphorus in water.  U.S.
    Health, Education and Welfare Publ.
Sawyer, C. N. 1947. Fertilization of lakes by agricultural drainage.
    /. New Engl. Water Works Assoc. 61:109-27.
Tennessee  Valley  Authority.  1966. Annual report of Division  of
    Health and Safety.
Verduin, J. 1964. Changes in western  Lake Erie during  the period
     1948-1962. Verhandl. Intern. Ver.  Limnol. 15:639-44.
	.  1967.  Eutrophication and agriculture in  the United  States.
    In Agriculture and the quality of our environment, ed.  Nyle C.
    Brady, pp. 163-72. Norwood, Mass.: Plimpton Press.
	.  1968.  Reservoir management  problems created by increased
    phosphorus levels of surface waters.  Am.  Fish. Soc.  Symp.,
    pp.  200-206. Athens: Ga.:  Univ. of Georgia Press.
       1969.  Man's influence on Lake  Erie. Ohio J. Sci. 69:65-70.
Verduin, J., and Loomis, W. E. 1944. Absorption of carbon dioxide
     by maize. Plant Physiol. 19:278-93.

 C.  A.  BLACK
       HE principal objective of this chapter is to present an account
of selected aspects of the behavior of soil and fertilizer phosphorus as
a basis for understanding how phosphorus from these sources may
contribute to the phosphorus content of waters in the soil and leaving
the soil.  An attempt is made to place  these matters in perspective
in the  broad picture without undue encroachment on the aspects of
the subject covered by other contributors to the symposium.
     Although the basis for the subjects discussed is mostly chemical,
an exhaustive review of current knowledge of the chemistry of phos-
phorus in soils and fertilizers will not be  attempted because such a
review would lose sight of the objective.  Chemically oriented reviews
have been published by Dean (1949), Wild (1950), Hemwall (1957),
Larsen (1967),  Mattingly and  Talibudeen (1967),  and  Huffman
(1968). Taylor (1967) published a review on phosphorus  and water
pollution with emphasis similar to that in this chapter.

     Vertical cycle

     Plant roots continually absorb small amounts  of phosphorus
from soil, generally less than 15 kg per hectare annually.  The major
portion of the phosphorus is transported to the above-ground organs.
The  phosphorus not contained  in harvested  parts is returned to the
surface of the soil in the plant residues.
     The phosphorus added to  soil in plant residues is constrained
against downward movement by a mechanical sieving  action of the
soil,  which is effective on the solid residues,  and by a chemical siev-
     C. A. BLACK is Professor,  Department  of Agronomy, Iowa  State
     Journal  Paper No. J-6373 of the Iowa Agriculture and Home Eco-
     nomics Experiment Station, Ames. Project No. 1183.

                        CHAPTER 6 / SOIL AND FERTILIZER PHOSPHORUS / 73

ing action, which is effective on the phosphorus  that has  been re-
leased from the residues to the water in the soil.  The existence of a
chemical sieving action is  suggested by  data  by Ponomareva  et  al.
(1968) showing that the concentrations of phosphorus in micrograms
per rnilliliter in drainage water from successively deeper layers in a
soil from the USSR were 0.005, 0.001, 0, 0, and 0. Similarly, Barber
et al. (1962) measured an average concentration of 0.18 Mg of inor-
ganic phosphorus as orthophosphate per milliliter of the saturation
extract of the 0- to 15-cm layer of soils of midwestern  United States
and 0.08 p.g per ml of the saturation extract of the 46- to 61-cm layer
of the same soils.
     The combination of upward transport of phosphorus in the soil
profile by plants and the retention of phosphorus by the  soil against
downward transport by water may significantly  alter the vertical
distribution of phosphorus in the soil.  In soils that have been sub-
jected to moderate weathering and leaching, a minimum in the con-
centration of total phosphorus in the soil may be found a small dis-
tance below the surface (Fig. 6.1).
                                Dilute- |
                                acid-  11
                                soluble  I x
                               inorganic £ |
                    Phosphorus content of soil,%
     FIG. 6.1.  Vertical distribution of phosphorus in a soil developed  on
     loess under  grass vegetation  in Iowa. (Pearson and Simonson, 1939;
     Pearson et al., 1940.  Reproduced by permission  of John Wiley  & Sons,
     Inc., New York.)


      Chemical cycle

      The phosphorus plants absorb from soil is presumably inorganic
  orthophosphate.  In plants, perhaps half of the  phosphorus occurs
  as inorganic orthophosphate and almost all the remainder as various
  organic forms.  Plant residues therefore  return to the soil  some in-
  organic phosphorus and some organic phosphorus.
      Inorganic phosphorus, present in relatively high concentration
  in the plant sap, diffuses readily from the dead plant material into
  the soil.  In the  soil it reacts with the soil minerals, and the concen-
  tration in solution is much reduced, as may be inferred from the ex-
  periment by Ponomareva et al. (1968) discussed previously.
      A small proportion of the organic phosphorus probably diffuses
  out of the plant residues into the soil, but most of it is  not readily
  soluble in  water and presumably must be  acted upon by microor-
  ganisms before  it is released.  Despite rapid decomposition in  the
 first few months, however, complete disappearance of added organic
 matter requires  a  long time.  A consequence is that during soil de-
 velopment  organic phosphorus is produced  at the expense of inor-
 ganic phosphorus.  The accumulation of organic phosphorus parallels
 the accumulation of organic carbon, nitrogen, and sulfur (Jackman,
 1964), and the content of organic phosphorus is usually  greatest at
 the surface and decreases with depth, as is true also of other organic
 constituents (Pearson and  Simonson,  1939). Figure 6.1  shows  the
 vertical distribution of organic phosphorus in one soil  profile.  In
 time,  presumably,  a steady state is reached in which organic phos-
 phorus changes to  the inorganic form as rapidly as it is produced.
     When soils  are cultivated, the previously existing balance  be-
 tween formation and decomposition of organic phosphorus is upset.
 Generally the content of organic phosphorus decreases (Haas et  al.,
 1961; Cunningham, 1963).
     One other aspect of the chemical cycle worthy of particular men-
 tion is that inorganic orthophosphate ions in the soil solution  ex-
 change continuously with inorganic orthophosphate ions held by the
 soil  solids (but not with organic orthophosphate).  The classic paper
 on this subject was published by McAuIiffe et al. (1948). In each soil,
 some of the solid-phase phosphorus exchanges readily with added
 radioactive  phosphorus,  some  exchanges  more slowly, and  usually
 most exchanges extremely slowly, if at all.  Phosphorus is supplied
 to the solution  from the readily exchanging fraction in response to
removal of  phosphorus from the  solution and is transferred from
the  solution to  the readily  exchanging  fraction  in  response  to
addition of phosphorus to  the  solution from  external sources. The
readily exchangeable fraction, in turn, gains phosphorus from other
sources in  the  soil when its level is decreased,  and it loses phos-
phorus to other forms when its  level is increased by phosphorus addi-
tions.  Larsen (1967) gave  an exceptionally  clear picture of these

                        CHAPTER 6 / SOIL AND FERTILIZER PHOSPHORUS / 75


     Soil contributes to the  geologic phosphorus cycle by  supplying
phosphorus in solution to ground-water and surface water and by
supplying phosphorus in suspended solids to surface water and air.
Emphasis here will be on the parts involving water.
     Loss of phosphorus from soil by drainage into the groundwater
is a normal part of the  geologic phosphorus cycle. Indirect evidence
of various kinds (Clarke, 1924; Weir, 1936; Wild, 1961; Ludecke,
1962)  indicates that during the time required  for soil to  develop
from parent material, a substantial part of the original phosphorus
may have disappeared,  presumably as a result of downward move-
ment of water through  the soil. The annual losses are so small in re-
lation  to  the  amount  present, however, as  to be undetectable by
analyses made of the soil over a  span of a few years or perhaps even
a lifetime.
     Loss of  phosphorus due to  downward movement  of water
through the soil on a  short-time basis is  commonly determined by
analyzing the drainage  water from a  lysimeter,  in which  the depth
of soil is usually no more than a meter, or by analyzing water from
tile drains.  The concentrations are usually less than 0.1 ^g of phos-
phorus per milliliter (Voelcker, 1874; Kohnke et.al., 1940;  Morgan
and Jacobson, 1942; Sylvester and Seabloom, 1963).  Analyses  for
phosphorus are often omitted because the concentrations are so con-
sistently low.
     Two  difficulties in interpretation of  values obtained  as  just
described are that  (1)  part  of the water in  tile  drains has passed
through  strata beneath the  tiles and  (2)  the  phosphorus filtering
process that goes on in  soil  proper takes place even more  effectively
in unconsolidated material underlying the soil. Occurrence of a zone
of relatively high  phosphorus content  in  the unconsolidated  ma-
terial below the  soil (Huddleston, 1969) is evidence that some of  the
phosphorus leached from the soil is retained by the material beneath
and that the estimate of loss of phosphorus  to the groundwater by
leaching may depend on the depth at which the water is collected.
     Because of the effectiveness of soil and underlying material in
retaining  phosphorus,  the phosphorus  content  of groundwaters is
normally low. A value of 0.011 ^ig of  phosphorus per milliliter is
obtained by averaging 63 of the 65 analyses reported by White et al.
(1963) in a survey of  data. Two  high values  not  included in  the
average were  0.15  and  0.36 ng  of phosphorus per milliliter. Juday
and Birge (1931) reported an average  of 0.016 Mg of phosphorus  per
milliliter in water from 17 wells near lakes in northeastern Wiscon-
sin (2  additional wells  had phosphorus contents of 0.086  and 0.197
Mg per  milliliter) and an average of 0.023 Ag  of phosphorus per mil-
liliter of lake water. Groundwaters may thus be expected to be low
in phosphorus in most instances.
     Surface waters present  a different sort of problem because they
contain phosphorus in  both dissolved  and particulate form.  The
solids  are derived primarily from  surface  soils (Gottschalk, 1962),

 FIG.  6.2.  Concentration of
 phosphorus in solution af-
 ter  equilibration  of  soils
 of France with superphos-
 phate  versus  calculated
 concentration   in  solution
 due to phosphorus added.
 Two  parts of water  were
 equilibrated with one part
 of soil.  Each line  repre-
 sents   a   different  soil.
 (Demolon   and  Boischot,
                                          5          10
                                       Phosphorus added per milliliter,
 most of which have, in the soil solution, concentrations of phosphorus
 exceeding the value of 0.015 vg per milliliter, quoted by Mackenthun
 (1965) as a concentration of phosphorus sufficient to produce a sub-
 sequent nuisance growth of algae in water.  Data on phosphorus in
 soil solutions were published by Pierre and Parker (1927) and Barber
 et al. (1962).  Another source of phosphorus in surface waters is dead
 plant  residues  on the surface of the soil.  These residues release
 phosphorus readily, and the initial phosphorus  concentrations  are
 much above those generally found in soil solutions.
     The suspended solids impart to the stream a phosphorus-buffer-
 ing  quality, illustrated in principle in Figure 6.2. That is, when the
 solids are initially suspended in rainwater or when the stream is later
 diluted by low-phosphorus  water, release  of phosphorus from  the
 solids will make the concentration of phosphorus in the final mixture
 closer to that  in the original soil solution  than would  be predicted
 from a simple dilution effect.  Conversely, if a stream receives high-
 phosphorus water from another  source, such as sewage,  the  soil-
 derived solids will take up phosphorus from the water and will reduce
 the concentration of phosphorus in solution.
     Whether the  entrance of groundwater into streams increases or
 decreases  the concentration of phosphorus in the stream water de-
 pends on the relative concentrations of phosphorus in the two. If the
 stream is  one  that carries  substantial amounts  of  suspended solids
 from surface soils and receives sewage effluent at intervals, it seems
 unlikely that entrance of the groundwater will raise the  concentra-
 tion of phosphorus. But,  even if the groundwater originally has a
 lower concentration of phosphorus than the stream water, the ground-
 water may not much lower the concentration in the stream because
 of the buffering effect of  the solids.  Groundwater enters streams
 mainly through the sides and bottom of the channel, and it must pass
 through  the previously deposited  sediments in the stream bed  and
must be substantially at equilibrium with them by the time it enters
 the stream proper.
     Concentrations of inorganic orthophosphate in the  water  of
streams  and lakes are  extremely low by conventional  standards.
Plants are extremely efficient in absorbing phosphorus, however,  and

                        CHAPTER 6 / SOU AND FERTILIZER PHOSPHORUS / 77

if other conditions are favorable, will reduce the external concen-
tration of phosphorus essentially to zero. Absorption of  dissolved
phosphorus by aquatic plants starts a biological cycle in which  ani-
mals feed on the  plants and  the residues of both decompose,  with
release of inorganic orthophosphate that starts around the biological
cycle again.
     This biological cycle continues after the water has reached the
oceans.  But the depth  and circulation of  oceans introduce some
changes.  Photosynthesis occurs  only near the  surface  because of
the requirement for light.  The residues of both plants and animals
sink and  decompose at great  depths or on the bottom, where there
is little synthesis.  Consequently the inorganic phosphorus content of
surface water is low, and that of deeper water is higher.  Circulation
of the oceans brings up water from the depths and renews the supply
of phosphorus for the  biological cycle.
     Despite  the annual addition of an  estimated 2 million metric
tons of dissolved phosphorus to the oceans (McKelvey et al.. 1953),
the phosphorus concentration in  ocean  water remains low because
of continuous loss of phosphorus from the biological cycle, the princi-
pal loss being due  to formation of the mineral apatite.  According to
Kazakov's theory (McKelvey et al., 1953), the  cold ocean water from
great depths, which contains a relatively high concentration of carbon
dioxide and inorganic orthophosphate, becomes supersaturated  with
respect to apatite as it flows upward, warms, loses carbon dioxide,
and increases in  pH.   Solid-phase apatite is then slowly  formed.
Apatite is forming now off the coast of California under these condi-
tions, according to Dietz et al. (1942).  If the apatite is formed in a
place that receives little extraneous sediment, a substantial and high-
grade  bed of "phosphorite" or "phosphate rock" may be developed
over geologic time.  If, later, the  bed of phosphorite is uplifted  and
occurs above sea level, the geologic phosphorus cycle  begins again
with  loss of phosphorus by leaching.  Phosphorus in phosphorite
reenters the geologic cycle in  another way in that beds of this sub-
stance located now on land supply almost all the phosphorus  used
for fertilizers and other purposes.

     Nature of fertilizer  phosphorus

     The phosphorus in phosphorite  is  present as  orthophosphate
(PO4- - -), and it remains as orthophosphate when phosphorite is proc-
essed to form the more soluble  phosphate compounds that contain
the bulk of the fertilizer phosphorus. In  some modem fertilizers,
however, a part or most of the phosphorus is now appearing as con-
densed  phosphates, in which two  or more  orthophosphate  groups
are joined  through an oxygen atom.  The solubility of condensed
phosphates decreases with an increase in size of the molecules.
     The chemistry of condensed phosphates is somewhat different
from that of  orthophosphates (Huffman,  1968).  For present pur-

FIG. 6.3.  Crystalline phosphates formed from the interaction of phos-
phate  fertilizers with soils.  A.  Crystals of  calcium  ammonium  phos-
phate  [CadslHiWHPCXkHaO (dimorph B)]  formed  on calcium  carbonate
in a calcareous soil to which dibasic ammonium  phosphate was added
as fertilizer.  The fertilizer was added in a thin  layer  and moved  up-
ward into the soil.  The surface of the calcium carbonate shown in  the
picture  was oriented perpendicular  to the layer of fertilizer and paral-
lel to the direction of movement. (Bell, 1968.) B.  Variation  in content
of calcium  ammonium  phosphate  [Ca(NHi)o(HPOi)2.H2O  (dimorph  B)]
with original  pH of a soil  when dibasic  ammonium phosphate  was al-
lowed  to move upward into  a column of soil from a thin layer at  the
bottom.  The soil is dark colored,  and the intensity of the sprinkling of
white indicates the  relative amount of calcium  ammonium  phosphate
formed.  The sample  in  each  case was taken  from a 2-mm layer of  soil
adjacent to the fertilizer and  had been pressed  into  a  brass ring pre-


paratory to examination  by X-ray diffraction.  (Photograph courtesy of
L.  C.  Bell.)  C.  Crystals of magnesium ammonium  phosphate hexahy-
drate  (MgNH4PO4.6H2O)  developed  in  a  soil  high in  exchangeable
magnesium when monobasic ammonium  phosphate moved upward  into
a column of the soil from a thin layer  of the salt at the bottom.  Values
for cation-exchange capacity, exchangeable calcium, and exchangeable
magnesium in the soil were 54, 38, and 12 m.e. per  TOO g, respectively.
(Photograph courtesy of L.  C.  Bell.)  D.  Cross  section  of an  originally
neutral,  high-calcium soil, adjacent to a  granule of  concentrated super-
phosphate,  showing development  of crystals  of dibasic  calcium phos-
phate  dihydrate (CaHPCX.2H«O).  Although the soil contains a large
amount of the newly formed crystalline phosphate, most of the crystals
are too small  to be identified  at this magnification.  Only a  few rela-
tively large crystals may  be seen.


 poses,  however, it is perhaps  sufficient to say that (1) condensed
 phosphates  spontaneously  decompose  in  soil,  gradually  forming
 orthophosphates, and (2) in the meantime  they are present in forms
 that are probably no more readily lost from the soil than are ortho-
     Reactions at high phosphorus concentrations

     The quantities of phosphate fertilizers added to soil are rarely
 great  enough  to produce a high  phosphorus concentration in  the
 soil solution if the fertilizer phosphorus were uniformly distributed,
 but uniform distribution is never accomplished in practice.  Initially
 the solution is usually saturated with the fertilizer salt at the imme-
 diate site of application, and the concentration  of phosphorus is of
 the order of 1 million times greater than the concentration of phos-
 phorus in soil solutions and streams.
     Soils  invariably  contain cations that  form  phosphates of low
 solubility (calcium, magnesium, aluminum, and iron are of principal
 importance), and such phosphates form rapidly in soil in the presence
 of  the high concentrations of phosphorus found near the site of  the
 fertilizer.  Some  of the phosphates are crystalline, and the crystals
 may be seen under  a microscope  and occasionally even with  the
 unaided eye.  Figure 6.3 shows  some examples.  The  kinds  and
 amounts formed depend  on the nature of the soil and fertilizer and
 on  other factors  as well (Bell,  1968; Huffman, 1968).  Formation of
 these compounds greatly decreases the tendency of the  phosphorus
 to move in the soil water by either mass movement or diffusion.
     The crystalline phosphates that form  quickly in soil when sol-
 uble phosphate fertilizers are  added disappear with time when  the
 concentration of phosphorus  decreases.  They may  simply dissolve
 (Larsen et al., 1964),  or they may leave a less soluble phosphate as a
 residue (Bell, 1968).  In either case the phosphate released does not
 stay in solution but is retained in some way by the soil solids. There
 is some evidence for  eventual formation of crystalline phosphates of
 extremely low  solubility (Nagelschmidt and Nixon, 1944; Australia,
 1956; Bell, 1968). On the other hand, if the phosphorus concentration
is maintained,  the quickly forming phosphates may be stable indefi-
 nitely.   This situation will be  discussed in the  section on  reaction
     1.  Scott  (1958) investigated the reaction of orthophosphate  and
     condensed  phosphates  (from  calcium  metaphosphate  fertilizer,
     vitreous calcium metaphosphate,  sodium metaphosphate, and  am-
     monium metaphosphate) with soils and found that the soils tested
     sorbed  the condensed phosphate more strongly than the orthophos-
     phate.  Sample  (1965) was  quoted by Huffman  (1968) as having
     found that pyrophosphate was taken up by soil  more rapidly than
     orthophosphate but was retained less strongly.  Gunary (1966) found
     that most soils he tested had a higher "adsorption maximum"  for
     pyrophosphate than for orthophosphate.  The adsorption concept is
     discussed in a subsequent section.

                        CHAPTER 6 / SOIL AND FERTILIZER PHOSPHORUS / 81

     Reactions  at  low phosphorus concentrations

     Reactions at low phosphorus concentrations are important at the
 perimeter of the zone of soil containing fertilizer phosphorus and also
 in stream  waters, where  suspended and  sedimented  solids interact
 with waters having low concentrations of phosphorus.  Figure 6.2
 shows that as phosphorus was added to suspensions of soil in water
 the concentration of phosphorus in solution increased,  slowly at first
 and then more rapidly, but the concentrations of phosphorus in solu-
 tion with no addition and the rates of increase with phosphorus ad-
 ditions differed among soils.  These  observations signify that soils
 react most strongly with the first increment of added phosphorus and
 less strongly with succeeding increments  and that the reaction has
 both an intensity aspect and a quantity aspect.
     The Freundlich and Langmuir equations used in  colloid chem-
 istry to describe adsorptions have both been used to express the reac-
 tions of soil with low concentrations of  inorganic orthophosphate.
 Recently attention has been focused on the  Langmuir  equation.
 Olsen and Watanabe (1957) used the equation in the form
                           C     1    C
                          	—	1	
                          x/m   kb   b
 in which C is the equilibrium phosphorus concentration, x/m is the
 quantity of phosphorus adsorbed per unit weight of soil, b is  the maxi-
 mum quantity of phosphorus that can be held by adsorption per unit
 weight of soil, and k is a parameter related to the bonding energy of
 the soil for phosphorus. If experimental data fit the equation, a plot
 of C/x/m  against C should yield a straight line with  slope 1/b and
 intercept 1/kb, from which  b and k may be evaluated.  The quantity
 of phosphorus  found in  the soil  by isotopic  dilution  of  radioactive
 orthophosphate was used as an estimate  of the phosphorus  already
 present in  adsorbed form.
     Figure 6.4 shows Olsen and Watanabe's data for two soils. From
 the equations, it may be seen that Pierre  clay had a higher  adsorp-
 tion capacity  (b = 25.9  mg P/100 g  soil) than Owyhee  silt  loam
 (b = 13.3 mg P/100 g soil)  and that Pierre clay bonded phosphorus
 more  strongly  (fe — 1.32 x  104 liters/mole)  than did Owyhee silt
 loam (k - 0.94 x 104 liters/mole).
     The constants in the Langmuir  equation provide  a  convenient
 way to represent the phosphorus-adsorbing properties of soils in the
 presence of low concentrations of phosphorus in solution and  provide
 reasonable bases for comparing different  soils if the procedures are
 standardized.  Moreover, the Langmuir equation may be used to de-
 scribe phosphorus release or desorption from soil as shown by Fried
 and  Shapiro (1956).
     The experience obtained in use of the Langmuir equation with
soils suggests that it  may be useful also  for describing phosphorus
adsorption  and release by solids suspended in streams.  Nevertheless,
the results  should not be taken too seriously because the equation is
empirical as applied to interaction of phosphorus with soil, and the

                                             Owyhee silt loam
             FIG.  6.4.   Plot of phosphorus adsorption data for two soils according to
             the Langmuir equation. In each case the first five points fall close  to
             a straight line, indicating conformance  to  the equation,  and the sixth
             point deviates from the line.  (Olsen and Watanabe, 1957.)

                         CHAPTER 6 / SOIL AND FERTILIZER PHOSPHORUS / 83

 equation does not fit  data from all soils (Fried and Shapiro, 1956;
 Olsen and  Watanabe,  1957;  Thompson et al., 1960).  Deviation of
 the data from a linear plot at high concentrations, illustrated in Fig-
 ure 6.4, is a common problem; this means that the adsorption capacity
 must be calculated and not determined directly. Olsen and Watanabe
 (1957) quoted other work suggesting that the deviation  might be due
 to formation of crystalline phosphates.
     Reaction capacities

     Of great importance  in  the behavior of phosphorus in soil in
 relation to water pollution is  the capacity of soil to react with phos-
 phate.  There is much confusion on this  matter because soils are so
 complex, conditions are  so  many,  and measurement  capabilities
 are so limited.
     For present purposes, it seems reasonable to describe, conceptu-
 ally,  three kinds of capacities. Each is  significant under different
     First is  the capacity of the soil to react with phosphorus at low
 concentrations.  This is the  so-called  adsorption  capacity discussed
 in the preceding section. It is of significance in both soils and streams.
     Second  is the  capacity of soil to  react rapidly with phosphorus
 added at high concentrations, as when water-soluble phosphate fer-
 tilizers  are added as solids. This capacity is significant in determin-
 ing the capability of soil to capture fertilizer phosphorus in new solid
 phosphate species and  to retain the phosphorus near the site of its
 introduction  into the soil.  This capacity  could be defined operation-
 ally in  many different ways, yielding  many different values.  Under
 conditions such as those of practical concern in the field, this capac-
 ity or these capacities far  exceed the  adsorption  capacity discussed
 in the preceding paragraph.
     Third is the ultimate  capacity of soil to react with phosphorus.
 The ultimate capacity is equal to the phosphorus retained by the total
 amount of cations in the soil capable  of  forming phosphates of low
 solubility.  This capacity is far greater than the capacity of soil to
 react rapidly with phosphorus added at high concentrations.
     The ultimate capacity is evoked when soil has been in contact
 with  a  solution of high phosphate concentration for  a long time.
 The original  carbonate, hydrous oxide, and silicate minerals are then
 decomposed,  with release of soluble silica from the silicates, and the
 product is a  bed of phosphates.
     There is no known instance in which soil has been thus altered
 by addition of phosphate fertilizer, but there is no doubt of the validity
 of the concept.  In  a classic paper, Gautier (1894) traced  a layer of
 clay that had entered a cave in France through a fissure in the rock
 and found that the clay had been altered  to an aluminum phosphate
 where it had  been contacted by water derived from bat guano.  Many
instances are known of alteration of rocks  to phosphates under the in-
fluence  of leachings from guano  in  caves  and  on ocean  islands.
Teall's (1898) photomicrographs of thin sections of trachyte slightly

FIG. 6.5.  Photomicrographs  of  thin  sections  of trachyte altered by
phosphate from  overlying  guano on Clipperton Atoll.  A. Altered tra-
chyte,  showing  phenocrysts of sanidine  set  in a  groundmass of micro-
litic feldspars and brown interstitial matter.  In the central  lower por-
tion of the photomicrograph is a crystal of the feldspar crowded  with
brown  inclusions.  The phosphorus is  present in  the  brown  substance.
B.  Highly  altered  trachyte, showing  the  replacement  of  feldspar by
phosphate with  concretionary  structure.  The groundmass  is  replaced
by  a similar material, but  without concretionary  structure.  The outline
of one of  the feldspar crystals is clearly seen  in  the  lower right por-
tion of the photomicrograph,  but the original  substance  has  been re-
placed  by the  phosphate.  (Teall, 1898.)

                        CHAPTER 6 / SOIL AND FERTILIZER PHOSPHORUS / 85

 TABLE 6.1.   Chemical composition of trachyte  at different degrees of altera-
            tion under guano on Clipperton Atoll.

K o 	
Loss on ignition 	
Matter insoluble in HC1 	

5^ 0

43 7
17 0


2 8
38 5


  Total  	     99.4           . . .           99.8

Source: Teall (1898).

altered and strongly altered by bird guano are reproduced in Figure
6.5, and his data showing the change in chemical composition of the
rock with degree of alteration  are given in Table 6.1. The inverse
relationship between phosphorus content and silicon  content is par-
ticularly noteworthy.
     Laboratory work has verified  that soils, clay, and minerals may
indeed be altered to phosphates. Gautier (1894) demonstrated the al-
teration of gelatinous alumina, clay, siderite, and chalk to phosphates.
Tamini et al. (1964) reported recent work  on gibbsite and soils and
reviewed some of the  previous work.  Clarke (1924)  reviewed early
work.  Modern  researchers have better tools than their predecessors
and now can determine more  easily the nature of the phosphates
formed.  Figure 6.6 shows, for  example, a  cross section of a crystal
of calcite, the surface  of which had been altered by a sodium  phos-
phate  solution to a calcium phosphate identified by X-ray diffraction
as apatite.
     The  second and  third  kinds of capacities are  usually  great
enough to enable soil to retain  a tremendous amount of phosphorus
near the site of application  of  soluble phosphate fertilizer.  At the
same time, the  combined effect of all three kinds of capacities  keeps
the concentration of phosphorus in the soil solution at a low value in
soil only a few  centimeters away.
     Addition of phosphate  fertilizers in agriculture  is never con-
tinued to the stage at which the  ultimate capacity of soil to react
with phosphate is satisfied because such additions would be accom-
panied by unfavorable effects  on plants.  The maximum favorable
effects are achieved with comparatively small additions.
     In terms of the concentration of phosphorus in solution, the con-
sequence of adding so  much  phosphate fertilizer that the  soil is con-
verted  to a phosphate bed would depend on the circumstances. One
solid figure—the value of 0.15 Mg of phosphorus per milliliter of
groundwater from the Phosphoria phosphorite formation of Garrison,

  0,1 mm
   ^* • -iwar-^

FIG. 6.6.  Cross  section  of  a  crystal of  calcite  showing a layer  of
apatite developed on the surface when the calcite  was immersed  for 2
weeks in a  0.1-molar solution  of tribasic  sodium  phosphate.  (Ames,

                         CHAPTER 6 / SOIL AND FERTILIZER PHOSPHORUS / 87

Montana—was recorded by White et al. (1963).  This value is to be
regarded as a minimum that might be reached some years after appli-
cation of soluble phosphate had been discontinued.  Higher concen-
trations would be expected as long as more soluble phosphates of the
type illustrated in Figure 6.3 remained.

     Inorganic phosphorus

     Most phosphate fertilizer is added to soil as a solid.  The highly
soluble phosphates attract water from the surrounding soil and form
a saturated solution of the  fertilizer, first in the fertilizer itself and
then in the surrounding soil as  the solution is drawn into the soil by
capillarity.  If the bulk soil is relatively dry, the soil around the ferti-
lizer is visibly wetted by the water that has accumulated (see Fig.
6.7).  This process was described by Lehr et al. (1959) and Lindsay
and  Stephenson (1959).
     During outward movement of the solution  the concentration of
phosphorus decreases because  of reaction with the soil, exhaustion
of the soluble salts in the fertilizer, and dilution of the solution with
water in the  soil.  Eventually  the concentration of the solution be-
comes low enough so that water is no longer drawn to any appreciable
extent from the surrounding soil. Within a few weeks the concen-
                                            FIG. 6.7.  Wetted zone of
                                            soil  around  a granule of
                                            concentrated   superphos-
                                            phate. The granule of fer-
                                            tilizer  was  imbedded  in
                                            the  smooth  surface  of  a
                                            dry  soil, and the soil  was
                                            exposed  to an atmosphere
                                            saturated  with  water va-
                                            por. The fertilizer took up
                                            the  water vapor, forming
                                            a  solution, and  the  solu-
                                            tion moved into the soil by


      FIG. 6.8.   Autoradiographs,  showing distribution of radioactive  phos-
      phorus along water-saturated columns of soil  2 weeks  after placement
      of phosphate fertilizer  at the lower end of the columns.  The darker
      areas represent higher  concentrations of radioactive phosphorus.  A —
      Cecil  sandy loam,  KHsP^CX  source; B = Elliott  silt  loam,  Ca(H2P3-Qt)2
      source; C = Fargo silty clay loam, KHsP^CX source; D = Miami silt loam,
      KH2P32O4 source; E = Miami silt loam, Ca(H2P'?O4)2 source.  (Bouldin and
      Black, 1954.)

tration of phosphate in solution is so low that little further movement
occurs over a much longer time  by  either diffusion or mass movement
in moving water.
     Generally, the concentration  of total phosphorus in soil a few
weeks after addition of a soluble fertilizer is greatest at the site of, or
immediately adjacent to, the fertilizer and gradually decreases  with
distance from  the site of the fertilizer.  The distribution pattern, how-
ever, is not always like this.  Figure 6.8,  for example, shows an in-
stance (autoradiograph E) in which there were two maxima in the
distribution of phosphorus with  distance from the source. Bell (1968)
observed occurrences of bands  of  crystals of  dibasic calcium phos-
phate dihydrate in glass-fiber filter paper  imbedded in soil in which
phosphorus was slowly moving from a layer of  soluble phosphate fer-
tilizer. The phenomenon of periodic precipitates or Liesegang rings

                        CHAPTER 6 / SOIL AND FERTILIZER PHOSPHORUS / 89
                                       135 kg  of  P
                                       added per hectare
               20        40        60        80        100
             Extractable phosphorus per gram  of  soil,jjg
     FIG. 6.9.  Extractable phosphorus  at  different  depths in  unfertilized
     and phosphate-fertilized  silt loam soil  in Wisconsin.  The soil received
     a surface application of  superphosphate equivalent to 135  kg of phos-
     phorus per hectare  on April  25  and  was sampled on October  15  of
     the same year for analysis  by Truog's (1930) method.  (Midgley, 1931.)

thus seems to have application in soil as well as in more homogeneous
media usually studied by chemists.
     Results of three field  experiments on movement of fertilizer
phosphorus in soil will be  cited.  Figure  6.9 shows an instance  in
which the  increase in extractable phosphorus in  the soil in the
autumn following an early spring topdressing of superphosphate was
confined to the surface 6 cm. Results such as this are characteristic
of soils  with moderate capacities  to react rapidly  with  phosphate
added at high concentrations.
     In the second experiment  (Ozanne, 1962), the  equivalent  of
225 kg of P32-labeled superphosphate per hectare was broadcasted on
a fallow siliceous sand in  the winter season in Western Australia.
After 38 days, during which  a total  of 23 cm of rain was received,
more than 50% of the labeled phosphorus had penetrated  more than
1 meter below the surface of the soil. These results are characteristic
of soils that have little capacity of any  kind for reaction  with  phos-
     The third experiment (Fig.  6.10) shows the measurable accumu-
lation  of phosphorus that occurred with time  when repeated  addi-
tions of superphosphate were made to a soil with moderate  capacity
to react  quickly with fertilizer phosphorus. The downward  penetra-
tion  was such that after 31 years the plots receiving 60 kg  of  phos-
phorus per hectare at presumably  annual intervals could be clearly
distinguished from the control plots by .analyses of samples of soil
from the 40- to 60-cm depth.  Plots receiving 180 kg could be clearly

_- 40

                        CHAPTER 6 / SOIL AND FERTILIZER PHOSPHORUS / 91

 than of phosphorus derived from inorganic fertilizers.  The findings
 have further relevance to attempts to use soil to remove phosphorus
 from sewage or livestock wastes. The efficiency of soil for this pur-
 pose may not be as great as it is for removing inorganic phosphorus.
 Koelliker  and Miner (1969) reported  that water from drain tiles at a
 depth cf 122 cm contained 0.5 ng of phosphorus per ml during a sea-
 son in which the soil was irrigated with livestock wastewater contain-
 ing 552 kg of phosphorus per hectare.  The chemical oxygen demand
 of the tile water was 37 jug/ml, which suggests that much of the rela-
 tively high concentration of phosphorus was organic.

 Alekseeva, E. N. 1968. Migration of phosphorus down the soil pro-
     file during long-term use of fertilizers.  (Translated title.)  Agrok-
     himiya, 1968, No. 8, pp. 78-82.
 Ames, L.  L., Jr. 1961. Anion metasomatic  replacement reactions.
     Econ. Geol. 56:521-32.
 Australia.  1956. Commonwealth Scientific and Industrial Research
     Organization, Ann. Kept. 8:18-19.
 Barber, S. A., Walker, J. M., and Vasey, E. H.  1962.  Principles of ion
     movement through the soil to the plant root. Trans. Joint Meet-
     ing Com. IV &• V, Intern. Soc.  Soil Sci. (New Zealand, 1962),
     pp. 121-24.
 Bell,  L. C.  1968.  Nature  and transformation  of  crystalline phos-
     phates produced  by  interaction of phosphate fertilizers  with
     slightly acid and alkaline soils.  Ph.D. Thesis, Iowa State Univ.,
 Bouldin, D. R., and Black, C. A. 1954.  Phosphorus diffusion in soils.
     Soil Sci. Soc. Am. Proc. 18:255-59.
 Clarke, F.  W.  1924.  The  data of geochemistry. U.S. Geol.  Survey
     Bull. 770.
 Cunningham, R. K.  1963.  The effect of clearing a tropical forest soil.
     /. Soil Sci.  14:334-45.
 Dean, L.  A.  1949.  Fixation  of  soil phosphorus.   Advan.  Agron.
 Demolon, A., and Boischot, P.  1951.  Reaction des sols a 1'apport de
     phosphates solubles.   Doses  isodynames.  Compt. Rend. Acad.
     Sci. 233:509-12.
 Dietz, R. S., Emery, K. O., and Shepard, F. P. 1942.  Phosphorite de-
     posits on the sea floor off southern California.   Bull. Geol. Soc.
     Am. 53:815-47.
Dyer  B.  1902.  Results of investigations on the Rothamsted soils.
     USDA, Office of Exp. Sta. BuU. 106.
 Fried, M.,  and Shapiro, R. E.  1956.  Phosphate supply pattern of
     various soils. Soil Sci.  Soc. Am. Proc. 20:471-75.
 Gautier, A.  1894.  Sur un gisement de phosphates de chaux et d'alu-
     mine contenant des especes rares ou nouvelles et sur la genese
     des phosphates et nitres naturels.  Ann. Mines (Ser. 9) 5:1-53.
Gottschalk, L. C.  1962. Effects of watershed protection measures on
     reduction of erosion and sediment damages in the United States.
    Intern. Assoc. Sci. Hydrol. Publ. 59, pp. 426-47.
Gunary, D.  1966.  Pyrophosphate in soil; some physico-chemical
     aspects. Nature 210:1297-98.


 Haas, H. J., Grunes, D. L., and Reichman, G. A.  1961.  Phosphorus
     changes  in  Great  Plains soils as influenced by cropping and
     manure applications. Soil Sci. Soc. Am. Proc. 25:214-18.
 Hannapel, R.  J., Fuller, W. H., Bosma, S., and BuUock, J. S. 1964a.
     Phosphorus movement in a calcareous soil: I. Predominance of
     organic forms of  phosphorus in phosphorus movement.  Soil
     Sci.  97:350-57.
 Hannapel, R.  J.,  Fuller, W.  H., and Fox, R. H. 1964b.  Phosphorus
     movement in a calcareous soil:   II.  Soil rnicrobial activity and
     organic phosphorus movement. Soil Sci. 97:421—27.
 Hemwall, J. B.  1957.   The  fixation of phosphorus by soils.  Advan.
     Agron. 9:95-112.
 Huddleston, J. H.  1969. Local soil-landscape relationships in eastern
     Pottawattamie County, Iowa. Ph.D.  Thesis, Iowa State Univ.,
 Huffman, E. O. 1968. The reactions of fertilizer phosphate with soils.
     Outlook Agr. 5:202-7.
 Jackman, R. H.  1964.  Accumulation  of organic matter in some New
     Zealand  soils under permanent pasture.  I.  Patterns of change
     of organic  carbon, nitrogen, sulphur, and  phosphorus.  New
     Zealand J. Agr. Res.  7:445-71.
 Juday, C., and Birge, E. A. 1931.  A second report on the phosphorus
     content of Wisconsin lake waters.  Trans.  Wis.  Acad. Sci. Arts
     Letters 26:353-82.
 Koelliker, J. K, and Miner, J. R.  1969. Use of soil to treat anaerobic
     lagoon effluent: renovation as a  function of  depth and applica-
     tion rate. Paper presented at meeting of Am. Soc. Agr. Engrs.,
     June 1969, Purdue Univ., West Lafayette, Ind.
 Kohnke, H., Dreibelbis, F. R., and Davidson,  J. M.  1940.  A survey
     and discussion of  lysimeters and a bibliography on their con-
     struction  and performance. USDA Misc.  Publ. 372.
 Larsen, S.  1967.  Soil phosphorus. Advan. Agron. 19:151-210.
 Larsen, S., Gunary, D., and Devine, J.  R.  1964.  Stability of granular
     dicalcium phosphate dihydrate in soil. Nature 204:1114.
 Lehr, J. R., Brown, W. E.,  and Brown, E. H.  1959. Chemical  be-
     havior of monocalcium phosphate monohydrate in soils.  Soil
     Sci. Soc. Am. Proc. 23:3-7.
 Lindsay, W. L.. and Stephenson, H. F.  1959. Nature of the reactions
     of monocalcium phosphate monohydrate in soils:  I. The solu-
     tion that  reacts with the soil.   Soil Sci. Soc. Am. Proc. 23:12-18.
 Ludecke, T. E. 1962.  Formulation of  a rational fertiliser programme
     in tussock country. Proc. Neiv Zealand Grassland  Assoc.  24:
 McAuliffe,  C. D., Hall, N. S., Dean, L.  A., and  Hendricks, S. B. 1948.
     Exchange reactions between  phosphates and soils: hydroxylic
     surfaces of soil minerals. Soil Sci. Soc. Am. Proc. 12:119-23.
 McKelvey, V. E., Swanson, R. W.,  and  Sheldon. R.  P.  1953. The Per-
     mian  nhosphorite  deposits of western United  States.  Congr.
     Geol. Intern. Compt.  Rend.  19th Sess.  11:45-64.
Mackenthun, K. M. 1965. Nitrogen and phosphorus in water. U.S.
     Dept. Health, Education, and Welfare, Public Health Serv., Div.
     Water Supply and Pollution Control.
 Mattingly, G. E. G., and Talibudeen. O. 1967. Progress in the chem-
     istrv of fertilizer and soil phosphorus.  Topics  Phosphorus Chem.
 Midgley, A. R. 1931.  The movement and fixation of phosphates in

                        CHAPTER 6 / SOIL AND FERTILIZER PHOSPHORUS / 93

     relation to permanent pasture fertilization.  /. Am. Soc. Agron.
Morgan, M. F., and Jacobson, H. G. M.  1942.  Soil and crop interre-
     lations of various  nitrogeneous fertilizers.  Windsor lysimeter
     series B.  Conn. (New Haven) Agr. Exp. Sta. Bull. 458.
Nagelschmidt, G., and Nixon, H. L. 1944.  Formation of apatite from
     superphosphate in  the soil. Nature 154:428-29.
Olsen, S. R., and Watanabe, F. S.  1957.  A method to determine  a
     phosphorus adsorption maximum of soils as  measured by  the
     Langmuir isotherm. Soil Sci. Soc. Am. Proc. 21:144-49.
Ozanne, P. G. 1962.  Some nutritional problems characteristic  of
     sandy soils. Trans. Joint Meeting Com. IV & V, Intern. Soc. Soil
     Sci. (New Zealand, 1962), pp. 139-43.
Pearson, R. W., and Simonson, R. W. 1939.  Organic phosphorus in
     seven Iowa soil profiles: distribution and amounts as compared
     to  organic carbon  and nitrogen.  Soil Sci.  Soc. Am.  Proc. 4:
Pearson, R. W., Spry, R.5 and Pierre, W. H. 1940. The vertical distri-
     bution of total and  dilute acid-soluble phosphorus in twelve Iowa
     soil profiles. J. Am. Soc. Agron. 32:683-96.
Pierre, W. H., and Parker, F. W. 1927.  Soil phosphorus studies:  II.
     The concentration  of organic  and inorganic phosphorus in  the
     soil solution and soil extracts and the availability of the organic
     phosphorus to plants. Soil Sci. 24:119-28.
Ponomareva, V. V.,  Rozhnova, T.  A.,  and Sotnikova, N. S.  1968.
     Lysimetric observations on the leaching of elements in podzolic
     soils.  Trans. 9th Intern. Congr. Soil Sci. (Australia) 1:155—64.
Scott, C.  O.  1958.  Sorption of orthophosphate and nonorthophos-
     phate phosphorus by soils.  Ph.D. Thesis, Iowa State Univ., Ames.
Stephenson, R. E., and Chapman, H. D. 1931. Phosphate penetration
     in field soils.  /. Am. Soc. Agron. 23:759-70.
Sylvester, R. O., and  Seabloom, R. W. 1963. Quality and significance
     of  irrigation return flow.   /. Irrig.  Drain. Div.,  Proc. Am. Soc.
     Civil Eng. 89, No. IR3, pp. 1-27.
Tamini, Y. F., Kanehiro, Y., and Sherman, G.  D.  1964. Reactions of
     ammonium phosphate with gibbsite and with montmorillonitic
     and kaolinitic soils.  Soil Sci. 98:249-55.
Taylor,  A. W. 1967. Phosphorus and water pollution.  /. Soil Water
     Conserv.  22:228-31.
Teall, J. J. H. 1898. A phosphatized trachyte from Clipperton Atoll
     (northern Pacific).  Quart. J. Geol. Soc. London 54:230-32.
Thompson, E. J., Oliveira,  A.  L. F.,  Moser,  U.  S., and Black, C. A.
     1960.  Evaluation  of  laboratory indexes of absorption  of  soil
     phosphorus by plants:  II.  Plant Soil 13:28-38.
Truog. E.  1930. The determination  of readily available phosphorus
     of soils.  J. Am. Soc. Agron. 22:874-82.
Voelcker, A.  1874.  On the composition of waters of land-drainage.
     J. Roy. Agr. Soc. Engl, 2nd Ser. 10:132-65.
Weir, W. W. 1936. Soil science. Chicago: J. B.  Lippincott.
White, D. E., Hem, J. D., and Waring, G A. 1963.  Data of geochem-
     istry, sixth edition. Chapter F.  Chemical  composition of sub-
     surface waters.  U.S. Geol. Survey Prof.  Paper 440-F.
Wild, A. 1950. The retention of phosphate by soil. A review. J.  Soil
     Sci. 1:221-38.
	.  1961.  A pedological study of phosphorus in  12 soils derived
     from granite.  Australian J. Agr. Res. 12:286—99.

        ITROGEN, like silicon, carbon, and phosphorus, has the unique
 ability to  act as a Lewis acid or base.  It has three p electrons, pre-
 sumably unpaired, each capable of entering into chemical reaction.
 Also, with all  three valence electrons tied up, the  inner  2s shell
 electrons  act as an electron pair donor and gave nitrogen its electro-
 negative character in secondary amines and other such compounds.
 Nitrogen is a gas at standard temperature  and pressure; its density
 is 0.81 g/ml.  When  combined  in compounds,  nitrogen  exhibits
 oxidation states from —3 to  +5. The elemental state is extremely
 electronegative with a value  of 3 on the Pauling scale.  Only a few
 of the nonmetals,  for example, O, F, and Cl, have electronegativity
 of equal or greater value.
     Nitrogen concentration in the dry atmosphere of the  earth is
 757.4 g/cm2 of earth's surface, according to Hutchinson (1954).  The
 total  atmospheric  mass  is  38.65 geograms.  (A geogram  is  1020
 grams.) The mass of water in the hydrosphere is 14,000 geograms.
 The amount of dissolved nitrogen is 0.26 geograms, or equivalent to
 5.2 g/cm2 of the earth's surface.  The average nitrogen concentration
 in igneous rocks is about 0.005% by weight.  It occurs in the form of
 ammonium substituted for potassium in mineral lattices.

     Ammonia  and other reduced forms such as nitrous oxide,  ni-
trites, etc., are oxidized by nitrifying bacteria to nitrates in water.
Organic nitrogen is primarily formed and degraded by biological  ac-
tion. Common species of organic nitrogen are proteins, protein deriva-
tives, purines, pyrimidines, and urea.  Some of these materials are
readily degradable and some are not.  Pyrimidines and purines are
important components of nucleotides and eventually may end up as
genetic components such as DNA and RNA.  Urea, on the other hand,

     MAHVIN C. GOLDBERG is Research Hydrologist, U.S. Geological Survey,
     Denver Federal Center, Denver.
     Publication authorized by the Director, U.S. Geological Survey.

                         Chapter 7 / SOURCES OF NITROGEN IN WATER / 95

 is a decomposition product of proteins or amino acids and is readily
 hydrolyzed enzymatically in  natural waters into ammonia and car-
 bon dioxide.  Urea is a highly available form of nitrogen for biological
     Another species  of bound nitrogen in water is the  solute from
 geochemical organic deposits.  Most of this type of material, which re-
 mains undissolved, does not enter the nitrogen cycle. Usually the non-
 reactive character of  geo-organic nitrogen is due to adsorption onto
 clay minerals or formation in complex forms which are polymeric and
 exist in water as polyelectrolytes.  Only about 5 to 10%  of this ma-
 terial is in the form  of nucleic acids, 30 to 40%  is  in  the form  of
 proteins, and  10 to 15%  is amino  sugars. The remainder has been
 uncharacterized. The majority of soluble organic nitrogen in lakes
 is present in the form of amino groups.
     For each 15 atoms of available nitrogen in water, there are 510
 atoms  of dissolved molecular nitrogen  and  a relatively unlimited
 supply  of elemental nitrogen both  in the atmosphere and the sedi-
 ments.  Hence, nitrogen becomes a limiting nutrient in water  bodies
 only because of the slow rate at which atmospheric nitrogen is fixed
 or the slow rate at which organic nitrogen deposits are degraded.

     Water supplies can be categorized as surface waters or ground-
waters.  This discussion will examine representative  studies of ni-
trate entrance to both  types of water supplies, with  summaries  of
some of the many  laboratory and field studies described in the cur-
rent literature. As the literature is voluminous, only some exemplary
studies are mentioned.

     Mechanisms for the  introduction of various fixed forms of ni-
trogen into water are categorized  as nitrogen fixation from  the air,
ammonia entrance from "rainout," entrance of organic nitrogen from
decomposing plants and  animals, and  land drainage. Water solu-
tions usually contain nitrogen in either organic or ionic form.  Am-
monium,  nitrite, and nitrate are the most common  ionic forms  of
nitrogen  found in water.  In  water itself,  as the  nitrogen  cycle
illustrates, proteinaceous material is decomposed by bacterial action.
The inorganic ions which result from this decomposition are in turn
used as nutrients to form new  cell material.  The  forms of the new
cell material are controlled by the environmental conditions imposed
upon the biological systems involved.   Grill and  Richards (1964)
examined nutrient regeneration from phytoplankton and observed
that at the end of  an experiment dealing with phytoplankton de-
composition, 33%  of the  total nitrogen was  ammonia, 39%  was  in
particulate matter, and 28% was in dissolved  organic compounds.



     A catalogue of some sources of nitrogen in water supplies would
     Rural runoff
     Tile  drainage
     Air  pollutants  from industrial sources
   Industrial  wastes
   Lake sediments
   Pond water
   Rural  waste
   Storm water
   Urban waste
     Leaking sewers
     Sanitary landfills
     Septic tanks
     Sludge  lagoons
     Waste stabilization  ponds
     Water treatment plants

     Precipitation might be the most important single source of ni-
trogen in surface waters (Feth, 1966), and thus it would also be avi
important source of nitrogen in groundwater.
     Most of the nitrogen in the atmosphere is in the molecular form
of N2; however, there are small amounts of  ammonia as well as
various oxides  of  nitrogen and their hydration products,  such as
nitric acid.  Much atmospheric ammonia is attributed  to industrial
air pollution. Additional sources are released  from soil decomposi-
tion  products and  photochemical reactions occurring in the strato-
sphere.  The most abundant oxide of nitrogen is  probably N2O.} pro-
duced by internal combustion engines. Ion molecule reactions which
occur in the stratosphere and upper ionosphere  account for forma-
tion  of nitrogen molecules other than N2.

                        Chapter 7 / SOURCES OF NITROGEN IN WATER / 97

     Stable aerosols composed of ammonium sulfate and ammonium
persulfate occur at altitudes between 15 and 25 kilometers above the
earth.  Particles are in a  constant condition  of  fallout from  the
stratosphere to lower layers of the atmosphere where they are in-
corporated into falling rain or snow. In a study by Junge (1968), it
was reported that nitrate and ammonium concentrations in rainwater
were low near coastlines.  During April 1958  to March 1959 about
59%  of  total inorganic  nitrogen  in rainwater at Yangambi,  Bel-
gian Congo, was ammoniacal nitrogen (Meyer and  Pampfer, 1959).
Nitrous nitrogen did not exceed 3%  of the nitric nitrogen.  Of special
interest is the fact that examination of individual downpours showed
that  the  smaller the downpour, the  higher the concentration of ni-
trogen, especially  ammoniacal nitrogen.  Matheson (1951) reports
6.5 kg nitrogen per hectare per year as accumulate nitrogen fall
contained in precipitation  and atmospheric sediments  collected  at
Hamilton, Ontario, with 61% of the total nitrogen  collected on 25%
of the days when precipitation occurred. The balance is  due solely
to sedimentation of dust. Fifty-six percent of the total was ammonia
nitrogen. In a New Zealand experiment (Miller, 1961) it was observed
that  total nitrogen collected at  the Taita Experimental Station from
rainwater was double the concentration of inorganic and  aluminoid
nitrogen.  Contributions from rainwater to nitrogen in soil would
probably be not less than 3.36 kg/ha/yr.
     Feth (1967) lists tables of data  indicating bulk precipitation of
nitrate in the Mojave Desert Region, California, between  March 1965
and March 1966. Values for nitrate nitrogen ranged from a trace to
as high as 16 mg/1 of rain, depending upon time of year and location.
     Wind-borne  sources   of  nitrogen  also  exist.   For  example,
McGauhey et al. (1963)  have  shown the amount  of nitrogen  con-
tributed by pollen may be  as high as 2 to 5 kg/ha/yr in  a forested

     Examples of geological sources of nitrate are the estimated 227
teragrams of nitrate of soda on the plateau of Tarapaca in Chili and
the significant amounts of nitrate in  the  Amargosa Valley, Inyo
County, California.  Nitrate deposits  have  been found in  soils  or
geologic formations in all of the 11 western states. Most of the states
in the Ozark and Appalachian plateaus have natural nitrate accumu-
lations in caves.  Other geological sources  of  nitrate  are  igneous
rocks, coal, peat beds and muck soils,  cave deposits, caliche deposits,
and playa deposits. In addition, all the world's organic matter, both
living and dead, plus that in sedimentary rocks, are potential sources
of nitrogen.
     According to notes  on the Conference on  Nitrogen Chemistry
held by the U.S. Geological Survey in  Menlo Park, California (1965),
fixed nitrogen in rocks may amount to a total  50 times as great as
the amount of fixed nitrogen in the  atmosphere. Rocks,  however,
are not a  ready  source of nitrogen  because  of  access  problems,


 attributed to small exposed surfaces.  Only limited zones near the
 surface are in position to yield their nitrogen freely  to circulating
 air and water.  They can be considered, however,  as groundv/ater
 nitrate sources under certain circumstances (Smith, 1967)—for ex-
 ample, in those localities where conditions have altered the geologic
 strata in such manner that there is collapse of cavern roofs or burial
 of ancient playa deposits which  become  geographically  placed in
 the zone of saturation.
      Organic-rich shales can  also be a source of nitrogen.  It has
 been reported that in sedimentary rocks,  concentrations as high as
 600  mg/kg dry weight  of nitrogen may be present.  Miocene  shale
 from the  Los Angeles Basin, California,  can  contain up to 8,600
 mg/kg nitrogen.
      The largest  geologic  concentrations  of  nitrogen seem to  be
 present in the  younger rocks; they are highest in  clay,  slate, and
 argillite,  and generally  low in metamorphic rocks.  Water released
 during metamorphism tends to be high in ammonia.   Igneous and
 sedimentary rocks may  contain nitrogen in amounts  ranging  from
 40 to 500 mg/kg, but organic shales contain nitrogen in much higher

     Nitrogen from  Lake Sediments

     In 14 samples  from the upper 10 cm of Lake Tahoe sediments
 analyzed  by the Kjeldahl method, nitrogen  concentrations  ranged
 from 0.06 to 16.6 mg/g dry weight and carbon nitrogen  ratios from
 3.7 to 28.4 (McGauhey et al., 1963).

     Decomposition  Processes in a Lake  as a Source of Nitrogen

     Koyama and Tomino (1967) studied  the mineralization of nitro-
 gen-containing materials in a lake. Their results are typical of many
 such  studies and show primarily (1) that during the early stages of
 the stagnation period, nitrogen fixation is generally more  active than
 denitrification.  Denitrificarion gradually exceeds nitrogen  fixation
 with  progressive stagnation and  (2) at  the end of the stagnation
 period, the  amount  of denitrified nitrogen is large  when compared
 with  other mineralized nitrogen compounds.   Denitrification is  the
 dominant process determining nitrogen metabolism in the lake water.
 The ratio  of mineralized carbon to nitrogen at the end of  the stagna-
 tion period is 3.5, considerably smaller than the value for plankton
 of 5.7.  Mineralization rates of carbon and nitrogen in the  organic
 detritus of the lake  studied were 51%  per year and 76% per year,

    Nitrate Metabolism  in  Lakes

    The following regime has developed in  Sanctuary Lake, Penn-
sylvania (Dugdale and Dugdale, 1965):  (1)  a spring bloom when

                        Chapter 7 / SOURCES OF NITROGEN IN WATER / 99

ammonia nitrogen,  nitrate  nitrogen, and molecular  nitrogen  are
assimilated strongly, in that order of importance; (2) a midsummer
period when weak assimilation of ammonia nitrogen and molecular
nitrogen, but not nitrate nitrogen, occurs; and (3) a fall bloom with
intense nitrogen  fixation and  some  ammonia nitrogen uptake,  but
characterized by  a low nitrate nitrogen  activity. Nitrogen fixation
and ammonia nitrogen uptake appear to proceed at  the  same time,
although ammonia uptake dominates in the spring and nitrogen fixa-
tion dominates in the fall.
    Biogenic interactions affect nitrogen sources in waste—for ex-
ample, fixed  nitrogen entering  a reservoir  is synthesized into  the
biomass  as protein and liberated upon death of the biological entity.
As much as  40% is  released to  the  aqueous  environment,  some
diffuses to the surface and escapes as a volatile gas,  some is denitri-
fied (gcing from the +5  oxidation  state to  —3) and  some is per-
manently incorporated into the bed sediments.
    Examination of a nitrogen cycle (Ehrlich and  Slack, 1969) re-
vealed that nitrogen  assimilation in a laboratory study, where  the
sole nitrogen source was a stream of calcium nitrate, followed  the
characteristic  pattern.  The  nitrogen was assimilated by plant life,
in this case algae, with slight demtrification occurring at high nitrate
concentrations. The organic nitrogen was converted  to ammonia by
proteolytic bacteria, with the possible escape of some ammonia. The
ammonia from the  organic compounds  was partly assimilated  by
algae  and partly nitrified by bacteria.   The nitrate, of bacterial
origin, was  assimilated by algae.  Nitrification  apparently was  not
of major importance in converting organic nitrogen to algal biomass.
    Analysis of surface and subsurface samples from western Lake
Superior (Putnam and Olsen,  1959)  showed that ammonia nitrogen
was present in trace amounts only,  usually less than  0.1 mg/1.  It
was found that  the  range of  organic nitrogen during  the year was
from  0.08 mg/1  in  the hypolimnion to  0.28  mg/1  at the surface.
The b"lk of nitrogen in the lake existed in the form of  nitrate  and
ranged from 0.93 mg/1 at  the  surface to 1.15 mg/1 in the  hypo-
limnion.  Nitrite  was practically indetectable.  Waters that entered
Lake  Superior from  its tributary streams contained very little free
ammonia. Nitrate concentrations in the  rivers were  lower than that
observed in the lake and varied from 0.16 mg/1 to  0.47 mg/1.  Ni-
trite was either absent or present only in trace amounts. In a second
publication (Putnam and Olsen, 1960)  it was  stated that nitrate-
nitrogen concentrations were  directly  related to the  depth of the
sample and in no case was the concentration lower in the deeper
water layers  than near the surface. As  expected,  nitrate nitrogen
in all tributary streams except one was considerably lower than that
observed in the lake. In August the  nitrate nitrogen range was 0.01
to 0.44 mg/1.

     Agricultural sources  of nitrogen result primarily from organic
and inorganic materials added to soils for crop nutrition.  Movement


  TABLE 7.1.   Estimate of nutrient contributions from various sources.
Domestic waste ....
Industrial waste ....
Rural runoff:
Agricultural land .
Farm animal waste .
Urban runoff 	

per year
tion in
per year
tion in
 Source: Task Group 2610-P Report (1967). Reprinted from the March 1967
 issue of J. Am. Water Works  Assoc.  Copyright 1967 by the Am.  Water
 Works Assoc., Inc.
 * Considers rainfall contributed directly to water surface.
 t Insufficient data available to make estimate.

 of these materials has been traced from their soil origin to entrance
 into  surface and  groundwater supplies.  Several of the  following
 studies indicate the fate of agricultural nitrogen-containing materials
 after entrance into the environment.
     A review paper (Smith, 1967) describes the use of fertilizer salts
 to supplement nitrogen in soils. This nitrate source feeds vegetation,
 is lost to the atmosphere by denitrification, and is removed from the
 soil by erosion and leaching.   Organic-matter  nitrogen lost from soils
 is usually attributed to mineralization.
     Estimate of Nutrient Contributions from Various  Sources

     Table 7.1 characterizes nitrate sources in water.  The relative
magnitude of runoff from agricultural land is noticeably large.
     Rural Runoff  as a Nitrogen  Source

     Approximately 742 million hectares of rural land in the United
States produce runoff. Major factors in rural runoff are amount of
water applied and  land use.  For example (McGuinness et al., 1960),
runoff is greatest  for a  corn crop, somewhat  less for wheat,  and
least when the land is in meadow. The mean concentration given in
milligrams per liter of total nitrogen constituents per  storm event
with land planted  to wheat varies between 6 and 9 (Weidner et al.,
1969). These data  are losses from a watershed varying in agricultural

                         Chapter 7 /  SOURCES OF NITROGEN IN WATER / 101

     Agricultural  drainage waters  contain nitrogen concentrations
 ranging from 1 to 60 mg/1, mostly in the form of nitrate. Sediment
 suspended in flowing water may carry  relatively high  amounts of
 ammonium nitrogen as  well as particulate organic nitrogen.  Dis-
 tribution of nitrogen in river waters in the United  States  ranges
 roughly from 0.1 mg/1 to 3.0 mg/1.
     Annual average nitrogen loss from  a watershed drainage of an
 apple orchard in Ripley, Ohio, was  1.0 kg total nitrogen  per hectare.
 The mean-runoff nitrogen concentration per  storm for the apple
 orchard was 4.9 mg/1 total nitrogen.  The results of this work indi-
 cate that rural runoff is a factor in  stream pollution and  must be
 considered as a source of nitrogen  in water supplies.
     Timmons et al. (1968) have conducted a definitive study on
 the loss of crop nutrients through runoff. As can be  seen from Table
 7.2, a study of the Barnes-Aastad Soil, Water, and Conservation Re-
 search  Association farm near Morris, Minnesota, showed 29.1 kg/ha
 nitrogen loss in the  year  1966, with a high of 9.65 cm of runoff, and
 100.7 kg/ha nitrogen loss in 1967, with  a high of 11.76  annual cen-
 timeters of runoff. Nitrate accounted for the majority of  the nitrogen
 loss and in 1966 was 0.89 and in 1967,  2.9 kg/ha. '

     Several materials used as  agricultural fertilizers are salts  of
nitrogen. One of the materials used commercially in large quantities
is ureaform. When evaluated as a nitrogen-loading material in soils,
TABLE 7.2.   Annual nutrient loss for two seasons for the natural-rainfall ero-
            sion plots.
Cropping Hectare
Treatments Soil Loss
Corn-continuous . .
Corn-rotation ....
Oats-rotation ....
Avg Kg per Hectare Nutrient Loss
Corn-continuous . .
Oats-rotation ....
Hay-rotation . ...












Source: Timmons et al. (1968).  Reprinted with permission.
* Excludes NH4- and NO.-N.

 it was found to  be relatively long  lived (about 1 year) and  stable
 (Brown and Volk, 1966).  Losses did occur once the nitrogen entered
 the soil biological cycle.  If water-soluble nonurea portions of urea-
 form were incubated anaerobicaliy  with  soil, appreciable quantities
 of hydrogen were produced.  Simultaneously NH,NOo was evaluated
 and in  studies where  the initial rate of  application  was  168 kg cf
 nitrogen per hectare it was found that 2 to 5%  of  the ammonium
 nitrate was in the soil to a depth of 0 to 15 cm.  Ureaform was found
 in amounts  of 18 to 249c  in soil from 0  to 15  cm in depth.  Of the
 1 68 kg nitrogen per hectare application it was noted that the percent
 excess was 9.091  of ammonium nitrogen and 8.973 of nitrate nitro-
 gen, whereas the ureaform  excess produced  was 9.907 ureaform
      The loss  of  urea  nitrogen  on leaching as traced by lysimcter
 studies  was  examined  (Overrein,  1968) during  a 12-week  experi-
 mental period, and  it  was shown that at urea application rates of
 less than 250 kg  of nitrogen per hectare, loss  was slight. At appli-
 cation rates  of 1.000 kg of urea nitrogen  per hectare, treatment was
                      o             o    i          '
 followed by  a leaching loss equivalent to 5rc  of the  added fertilizer
 nitrogen.  The volatile  ammonia  gas loss  was also characterised.
 The  highest  total of accumulated  loss  of ammonia was  equal to
 3.5r'  of the added urea nitrogen.  No gaseous nitrogen oxides were
 produced.  Trace   amounts of  tagged molecular  nitrogen were re-
 covered in the atmosphere above lysimeters receiving the urea nitro-
 gen during tests in which the higher  application rate was used.

     Ammonium nitrate and urea differ considerably in  the extent
to which they are adsorbed by the soil.  It is reasonable to expect
that they vary in their susceptibility to loss from the soil by leaching
into surface runoff water.  Urea, however, is hydrolyzed to ammonium
ion and ammonium ion is  nitrified to nitrate in soils. A time interval,
therefore, must be allowed between fertilizer  application  and occur-
rence of rainfall before computing nitrate runoff into surface waters.
Such a study was  conducted (Moe et al., 1968) and the ammonium
nitrogen losses  from urea-treated plots were approximately equal to
those from the  ammonium-nitrate-treated plots during artificial rain-
fall  applications.  In a second set of artificial rainfall applications,
losses  of ammonium nitrogen averaged 40% less.  It was concluded
that the  ammonium nitrogen is less susceptible to runoff loss in the
urea-treated  plots than in the ammonium-nitrate-treated plots.  An
explanation is that ammonium nitrate, because of its high ionization.
wculd  be adsorbed and held near the  surface of  the  soil.  A non-
ionized urea  would be carried farther down into the soil with the first
increment of rainfall and would be less subject to surface runoff loss.
     It was found that urea is rapidly hydrolyzed to ammonia in  the
soil: the only measurable amounts of urea occurred in the runoff from
socl plots and resulted from the  direct washing of urea from the sur-
face vegetation.  Total nitrogen  losses from all plots ranged between
2.4 and 12.7% .  These results are very similar to  those of Mce et al.

                          Chapter 7 / SOURCES OF NITROGEN IN WATER /  103

 (1968). As a general conclusion,  the amount of nitrogen in the run-
 off water from soils treated in this  manner would not contribute
 appreciably to  nitrate pollution of surface-water resources.
      White et al.  (1967)  studied  movement of NH,NO., applied  on
 a soil surface at the concentration of 224 kg/ha.  Six and three-tenths
 centimeters of artificial  rainfall  were applied in  a 2-hour period,
 during which the runoff from fallow soil was 4.9 mg/1 and from sod
 2.1 mg/l. After a few moments, most of the soluble nitrogen moved
 into the soil and was inaccessible to erosion processes.
      On the southern high  plains (Lotspeich et  al.,  1969) estimates
 are that 0.2%  of  the fertilizer nitrogen applied is found in surface-
 water runoff.  In nearly all the playas of the southern high plains,
 the nitrate content is less than  1.0 mg/1, revealing the fact that ni-
 trogen fertilizer applied to the farmland adds little nitrate to the sur-
 face water.
   A clue to nitrogen runoff from  soil, resulting in enrichment  of ni-
 trogen in water, may be discovered in the data of Pratt et al. (1967).
 It was found that the ratio  of  nitrogen removal to total crop yield
 was higher with Ca(NO;J)o treatment than with ammonium sources.
 The largest  amount  of  nitrogen removed by drainage water  as well
 as the highest  nitrogen depletion occurred in a soil relatively  high
 in organic content. Lack of organic matter may explain low nitrate
 runoff on the southern high plains.
     A study of 82,029  irrigated hectares (Carter et al.,  1969) illus-
 trated that subsurface drainage water contains more nitrate nitro-
 gen than does the irrigation w*ater. but concentration rarely exceeds
 5.0 mg/1 of nitrogen.  Concentrations of nitrogen in surface drainage
 waters are only slightly higher than concentrations in the irrigation
     A study in Britain demonstrated that the amount of nitrogenous
 fertilizers applied to agricultural land has doubled in the last 10 years.
 Land drainage has been shown to contribute much inorganic nitrogen
 to rivers. In the Great Ouse River (Owens and Wood, 1968)  sewage
 effluents contributed  a small proportion of the total concentration  of
 nitrogen, silicon,  chloride,  and sulfate;  however,  the bulk  of the
 phosphorus could  be  attributed to the effluent  sources.  Table 7.3

 TABLE  7.3.   Ranges  of some selected  nutrients in sewage effluents and land
           drainage entering the  Great Ouse:  Concentrations in the  river
           water are also included.
Carbon (soluble) 	
Arnmoniurn-N • ...
Total soluble phosphorus . .

. 16.0-32.0


3 5 12 4
3 0-14 9
0 01-04
0.0-2 9
6.8-9 0
0.07-5 0

Source:   Owens and Wood  (1968). Reprinted with  permission  from M.
Owens.  Copyright 1968,  Pergamon Publ. Co.

 (Owens and Wood, 1968) lists the nutrients entering the Great Ouse
 River yearly and their sources.
      Figure 7.1 shows the estimated total quantity of nutrients sup-
 plied by sewage effluents in comparison with the  total load of  nu-
 trients in the river.  Only about 10%  of the nitrogen  in  the river
 could be accounted for by the amounts discharged in sewage efflu-
 ents.  It was assumed that increase in river flow between the influent
 streams and the downstream limit of the reach, other than that from
 the sewage effluent, was  derived from  land  drainage and  that this
 land drainage would have  the  same average concentration of  nu-
 trients as the land  drain sampled.  Estimated nutrient loads  are
 between  1.0 and 2.3  times  greater than those actually determined.
 Hence, the assumption that the increase in  load results from land
 drainage may be in error.
     About 3.15 x 106 kg of nitrogen were carried by the Great Ouse
 River in  1966.  If  all of the material flowing down the river were
 derived from land drainage, the flows of nitrogen per unit  of catch-
 ment area would be  18.5. The amount of nitrogen applied per unit
FIG.   7.1.   Nutrient   bal-
ance in Great Ouse, March
1967. (Owens and Wood,
1968. Reprinted with  per-
mission from M. Wood.)
       N -  Mass  Flow of Nutrients in River

       N  =  Accumulated Flow of Nutrients
            From  Sewage Effluents
                      g 200
                               135  140   145   150  155   160  165

                                 CATCHMENT AREA  (thousand ha)

                        Chapter 7 / SOURCES OF NITROGEN IN WATER / 105

area in the Ouse basin, according to statistics supplied  by the Min-
istry of Agriculture, Fisheries and Food, was 65 kg/ha.

     One  study  (Olsen et al., 1969) related that more  leaching  of
nitrate nitrogen occurred between fall and spring than during the
growing season  and more under fallow than cropped conditions.

     Effluents from a tile drainage system in irrigated areas in the
San Joaquin Valley of California  showed  that  initial  tile  effluent
frcm a previously unirrigated, noncropped area had a nitrogen con-
centration  of  1.0 mg/1.  Another system that  had been planted to
alfalfa had a low discharge over a year's period and yielded  a range
of nitrogen between  2.0  and  14.3  mg/1.  In  systems  where  high
rates of nitrogen fertilizer were applied,  the concentrations ranged
up  to  62.4 mg/1.  Concentrations of nitrate in  all systems ranged
from 1.8 to 62.4 mg/1, with a weighted  average of 25.1 (Johnston
et al.,  1969).
     Nitrogen can be carried directly into surface drains with tail-
water  from fields  where fertilizer is  being applied in the irrigation
water.  In  addition, nitrogen can also come from nonirrigated  agri-
cultural land.  A  further  source of  nitrogen  is  soil  from  erosion,
with resulting increase  in  sediment load, plant nutrients, and pesti-
cides.  Soil nitrogen is  sporadically released to water, such  release
being produced  and greatly influenced by intensity of precipitation.
     It was shown that in sand columns (Preul and Schroepfer, 1968)
the breakthrough curve for ammonium nitrogen  occurs  between 0.5
and 1.0 in units of throughput volume per column weight measured
in liters per kilogram of soil. Flow rates varied from 200 to 1,170
ml/day, nH ranged from 7.1 to 7.6, and nitrogen absorntion ranged
from 24.7 to 126.0 /j.g/g of ammonium nitrogen, at equilibrium. Dis-
charge  velocities were approximately 4.93 cm/day at flow  rates of
1,000 ml/day.  The conclusions of this study are that movement of
nitrogen in soils is controlled by adsorption  and biological action,
and that where nitrogen  is in  a  nitrate form at the  pH  of usual
wastewaters there is no impairment to nitrogen movement. Biological
interference is minimized under flow conditions with limited oxygen
tension. These results indicate a minimal vertical flow.  In the study
indicated, the major part of nitrification was restricted to within 0.61
meter  of the surface.  The possibility of lateral flow, however, would
tend to move nitrogen  into surface  waters. Leachates of nitrogen
(NH4NOo)-treated  soil (Krause and Batsch, 1968) contained 4 to 7
mg/1 ammonium nitrogen; the soil lost 88% of  its treated nitrogen
between September and December.  Untreated soils slowly lost  ni-
trate nitrogen.



      A study  of irrigation return flow in the Yakima River Basin,
  Washington (Sylvester and Seabloom, 1963), was made for an irri-
  gated  area  of 151,870 hectares during an irrigation season extend-
  ing  from April through  September.  Average water diversion was
  20,000 m3 per hectare per year of which approximately 5,240 m3 was
  applied to land; the remainder was lost in canal seepage  and canal
  evaporation wastage.
      It was established that the evapotranspiration loss in  the irriga-
  tion water return would  result  in  a salt  concentration of 1.7.  The
  nitrate content  of the return  due  to evapotranspiration, leaching,
  and  ion exchange was 10 times greater than in the applied water.
  Removal of 37  kg/ha  of nitrate resulted from  irrigation leaching.
      Salts and sediments  are of great concern to water users (Peter-
  son et al., 1969).  Salt  and silt create the most  difficult problems to
 irrigation, agriculture, and subsequent users of return flow.  Under
 certain conditions, however, animal waste, plant nutrient,  and toxic
 elements become equally important.
      Nitrogen  Movement  in Soils

      In a leaching experiment (Sinha and Prasad, 1967), urea was
 found to be distributed mainly within the top 10 centimeters of the
 soil column and  very little of it  leached down below this  depth.
 Retention of urea in soils appears  to be due to its  conversion to
 ammoniacal form.
     In a typical  situation in  the  San Joaquin Valley  from 1962 to
 1966, it  was  found that the subsurface drainage  water of  fields
 irrigated  with  water containing 1.7  mg/1 of nitrate  nitrogen had
 nitrate levels averaging 44.5 mg/1.  Fertilization consisted of adding
 187 kg nitrate nitrogen per hectare per year (Doneen,  1968).
     It was also found that nitrate  concentrations of drainage waters
 from different  areas  generally paralleled the  amount of fertilizer

     Smith (1967) cites data to illustrate the diminishing concentra-
tion of nitrates 61 to 91 meters from the source origin.  A second
conclusion was that  at  the site under observation, leaching of fer-
tilizer nitrogen was relatively insignificant in  comparison  to other

     Macrae et al. (1968) in a study of submergence of tropical soils
determined  that considerable proportions of  applied nitrate nitrogen
had  been  immobilized into the soil organic fraction.  Six Philippine

                         Chapter 7 / SOURCES OF NITROGEN IN WATER / 107

 soils were used in the study to trace the fate of added nitrate nitrogen
 after submergence.
     Occurrence of Nitrates  in Wei!  Water

     Shallow wells frequently contain greater nitrate concentrations
 than deep wells.  As has been shown by many workers, this may be a
 result of  improper well  construction.  Shallow wells and deep  wells
 can be polluted by nitrates, either leached from the aquifer or trans-
 mitted to the aquifer by percolating  waters.  Sources of nitrogen in
 deep wells are strata through old wells which have  been abandoned,
 pumped wells with rusted or perforated  casings, improper  sewage
 and waste  disposal,  natural  sinkholes,  and river valleys  recharge.
 Nitrate levels fluctuate in wells on an annual basis.
     A survey of nitrates in private water supplies in Morgan County,
 Missouri, made by the  Missouri Division of Health (Inglish, 1967)
 showed that in 157 well waters tested,  40 contained no nitrate, 27
 contained less than 1 mg/1, 44 contained between 1 and  20 mg/1,
 20 contained between 20  and 45  mg/1, and 23  contained over 45
 mg/1, with  the highest  value being 200 mg/1.  Of  these wells only
 13 were cased to a depth of 30.5 meters or more and only 4  were free
 of nitrate.
     Nitrates  in Groundwater  Supplies

     Appearance of excessively high amounts of nitrates in ground-
waters has  been considered an indication of wastewater infiltration
into the supply.  The wastewater may originate  from septic tank
effluents, waste stabilization ponds, waste treatment plant effluents,
sludge lagoons, sanitary landfills, privies, barnyards, leaking sewers,
irrigation systems,  and  similar  sources.  Of course, these sources
carry public health implications.   Comly (1945), Metzler (1950), and
Whitehead and Moxon (1952) have reported on the hazard of nitrogen
in water supplies. One manifestation of nitrates is the disease  infant
methemoglobinemia. Livestock, chiefly hogs and cattle, are affected
adversely and exhibit poor growth characteristics.  Nitrates can cause
gastroenteritis and diarrhea. In some instances, high nitrogen levels
in water can be lethal.

     Forty-five mg/1 nitrate is the upper limit set by the  U.S. Public
Health Service for city potable water supplies.
     Some of the  most immediate sources of nitrate and nitrite in
groundwaters are domestic sewage effluents, fertilizers,  and wastes
from corrals.  Mean concentrations of nitrate nitrogen from wells in
nonirrigated  and irrigated regions of southern  Oahu were  1 ± 0.22
mg/1 and 8.2 ± 2.4 mg/1, respectively (Mink, 1962).  Mink attributed


  TABLE 7.4.   Characterization  of  waters in the  San Luis Valley,  Colorado.

         Water Characteristics
           Rio Grande River                Aquifer Characteristics
  Total dissolved solids    41-120 mg/1  Unconfined to depth of 100 feet
  Character              Ca(HCO3)2    Wells are 35-100 feet in depth
  pH                    6.7-7.4       Total dissolved solids  100-400 mg/1
                                     pH 7.0-7.8

  Source:  R. K. Glanzman and J. M. Klein.  (Data to be published.) Private
  this difference to percolation of nitrate materials previously added
  to the system in the form of fertilizer.
      A study made by  the U.S. Geological Survey in the  San Luis
  Valley, Colorado, conducted by Glanzman and Klein1 of the Colorado
  District found relatively high nitrate concentrations in wells at depths
  of 11 to 30 meters.
      The San Luis Valley has an arid  high-altitude climate with 15
  to 18 cm of rainfall annually. The Rio Grande is the major surface-
  water source; its nitrate concentration  ranges from 0.0 to 2.3 mg/1.
  Table 7.4 gives the other characteristics. Soil characteristics are given
  in Table 7.5.
      Nitrogen application on the surface soils of San Luis Valley was
  112  kg of nitrogen per hectare  per year in the form of ammonium
  sulfate.  The nitrogen was applied by disking, banding, or sideband-
 ing.  These treatments were followed by copious applications of water.
 Considering that the water table was within 30 cm of the surface,
 it is  likely that the high nitrogen content in wells was  due to percola-
 tion  from surface applications and especially from irrigation ditches.
 which  are  used  to disperse the  fertilizer in liquid form to soil sur-
      Figure 7.2 is a contour  map  showing concentrations of nitrate
 in the San Luis Valley. The irrigation ditches correspond to the lines
 of high nitrate concentration.  It is reasonable  to conclude that some
 of the nitrate  infiltration is coincident with the surface route of the
 main irrigation ditches.
      The fertilization  practices  followed in this valley  incorporate
 large amounts of ammonium sulfate, applied as dissolved solute, in
 irrigation water.   Considering the high  water table and the general
 lack  of  other  nitrogen sources such as feedlots, septic  tanks, and
 sewage-processing tanks, it appears that in this case the  nitrogen
 source is inorganic fertilizer.

     Nitrates tend to accumulate at the top of a groundwater column.
As leaching through a soil is a function of physical parameters, such

     1. Glanzman, R. K., and Klein, J. M. (Data to be published.) Private

                         Chapter 7 / SOURCES OF NITROGEN IN WATER / 109

 TABLE 7.5.  Characterization of the  principle  on the fan,  San Luis Valley,

    Gunbarrel loamy sand       0—48 inches course loamy  sand
                             48-60 inches sand to sand and gravel

Percent Passing Sieve Size pH meability
No. 4 No. 10 No. 200 (in./hr)
90-100 90-100 15-30 7.9-9.0 2.5-5.0
75-90 65-80 5-15 7.9-9.0 >5.
(in. /in.)
Depth to sand and gravel 24-60 inches
Water table range 1-5 feet
  Sand 83-92%                   pH
  Silt   4-11%                   Salinity millimhos/cm     1.3-1.8
  Clay   4-7 %                   Organic matter           0.4 or less
                                 CaCO3 equivalent percent  0.7-2.3
 Moisture at saturation 19-20
 Cation exchange capacity 5.0-8.3

 Source: R. K. Glanzman and J. M. Klein. (Data to be published.) Private

 as soil permeability, soil porosity, temperature, rainfall, snow melt,
 or volumes of irrigation water, it is obvious that nitrate concentration
 in a water supply will vary according to the season of the year and
 the amount of water flow at  any given time.

     Numbers  of livestock vary  throughout  the  United  States  as
follows: 74% of the hogs, 42% of the cattle, and 39% of the poultry
are contained in the north-central region (Loehr,  1969); the south-
central and  western regions contain 41%  of the  cattle;  and  the
poultry population is  evenly  divided throughout the country. Dra-
matic increases in  numbers of cattle have been noted in the United
States.  For example,  an increase of 36 million head has occurred
during  the past 25 years, and 17 million of this  increase occurred
during  the last 8 years.  The  poultry industry  today  is a $3.4 billion
industry.  Cattle feedlots  have expanded rapidly also in the last few
years.  Livestock on American farms produce about  1,814 teragrams
of manure each year. In units of population equivalents, the nitrogen
contribution  of domestic  wastes is estimated to be 3.6 to 5.4 kg  per
year, or 0.015 kg per capita per day.  For chickens, the nitrogen con-
tribution per  animal per day is 0.001; for swine, 0.02; for dairy cattle,
0.18;  and for beef cattle, 0.14. Swine,  dairy cattle, and beef cattle
on a population-equivalent basis  produce more nitrogen per capita
per day than that  derived  from domestic sewage  wastes.  The pro-
duction of animal  wastes  in the United  States exceeds the waste
produced  by  the  human population by about 5 to  1  on a BOD (bio-
logical oxygen  demand) basis, 10 to  1 on a total-dry-solids basis, or
7 to 1 on  a total-nitrogen basis (Table 7.6).

     FIG. 7.2.  Concentration  of nitrate, Rio Grande Fan, San  Luis  Valley,
     Nitrates  in Wells

     Analysis of 6,000 rural water supplies (Keller and Smith, 1967;
Smith,  1967) indicated that the sources  of nitrogen were  animal
wastes, improperly constructed shallow wells, and septic-tank drain-
age.  There was some evidence of nitrogen infiltration from heavy
annual applications of nitrogen fertilizer.   The soil was an  alluvial
sand.  Clay soils generally do not transmit nitrogen.  In this study,
livestock were considered a more important source of contamination
than  nitrogen fertilizer.   None of the  reservoirs sampled  showed
increases in nitrate due to fertilization.  It was thought that nitrate
infiltration  is relatively slow.  The infiltration  mechanism involves

                         Chapter 7 / SOURCES OF NITROGEN IN WATER /111

 TABLE 7.6.   Average animal waste characteristics.

               Kilograms per Animal per Day  Per Capita Equivalent*!
Dairy Cattle . .
Beef Cattle ....
0 136
0 027
0 001
0 023
0 11
1 7
0 09
1 7
0 11
1 5
 Source: Loehr (1969). Reprinted with permission from R. C. Loehr.
 * Based on average characteristics in municipal sewage:  0.077 kg BOD;
 per capita per day; 0.25 kg total solids per capita per day; and 0.015 kg
 total nitrogen per capita per day.
 "i" Number of people equivalent to one animal.
 4. Total Kjeldahl nitrogen.

 trapping soil fissures during drought  and further infiltration washed
 during times of  heavy rains.  Indeed, without microorganism pop-
 ulations to reduce nitrates to ammonia, the nitrates persist.
     Nitrogen from  sinkholes and cave leaching is  a source  of  ni-
 trate in wells.  It was estimated that as many as 1,450 known caves
 in Missouri contain bat  guano.  In Cooper County, Missouri, nearly
 50%  of drilled wells, over  85%  of dug wells, and 80% of springs
 contained more  than 5 mg/1 nitrogen.   Keller and Smith  (1967)
 attribute this to the following sources:  fertilizers, feedlots, bat guano,
 and biological waste materials.
     Waterfowl  as a Source of Nitrogen

     Duck wastes are quoted as being as high as 0.95 kg  of fixed
nitrogen per duck per year. Estimating 100 million waterfowl in the
United  States, they would produce  91 to 227 million kilograms of
nitrogen per year. Wild duck nutrient contributions to 1,416-hectare
Lake Chautauqua in Illinois were 14.3 kg of total nitrogen per hectare
of water (Paloumpis and Starrett, 1960).
     Feediots as  a Source of Nitrogen

     Many workers have shown that feedlot runoff pollutes streams;
such runoff has high ammonia concentrations and reduces the oxygen
content.  At 4.41 cm  of rain per hour, nitrogen concentrations in
the form of ammonia  can run as high as  400 mg/1 within  an hour
after  the rain starts.  Water  which  has moved  through  feedlots
commonly contains nitrates and ammonium  compounds, and has
an offensive odor. The animal wastes in feedlots and in other areas
of containment can, under the  proper conditions, act as sources of


 nitrogen in both surface waters and groundwaters.  Typical examples
 are  given below.
     It has been confirmed that  groundwater under feedlots is usu-
 ally contaminated by nitrate (Stewart et al., 1967, 1968). It has also
 been shown, however,  that nitrate  levels in the range  of 10  to  30
 mg/1 are found in groundwater beneath irrigated fields.
     Atmospheric  ammonia measured near feedlots (Hutchinson and
 Viets, 1969) was as much as 20 times greater than near control sites.
 The conclusion was that surface waters in the immediate vicinity of
 a feedlot  can become enriched in nitrogen by absorption of atmos-
 pheric  ammonia volatilized  from the feedlot.   These data seem to
 indicate that not only are runoff and percolation sources of nitrogen
 from feedlots,  but atmospheric pollution is  a serious consideration
 as well.
     Data from the above study show that at a  sampling station 0.4
 km west of a 90,000 unit feedlot,  2.8 kg/ha of ammonia were ab-
 sorbed  each week, which would  be  146 kg/ha  on  an annual  basis.
 In other sites where no feedlots  were located, the  weekly ammonia
 nitrogen absorption was 0.15 kg/ha.
     Hanway et al. (1963) found evidence for the  fact that nitrates
 are more  concentrated  below or near the area of a waste accumula-
 tion or disposal, such as manure  piles, feedlots, septic tanks, disposal
 fields, cesspools, and privies, than in other areas of a fertilized field.
 Nitrate also may be marshalled in water under low areas and water-
 ways that convey runoff from higher ground.  Water which percolates
 through feedlots, decomposing peat  soils, heavily mineralized soils,
 or other nitrogen sources moves  nitrates to the groundwater.
     Stewart et al.  (1968)  found that nitrate concentrations in  soil
 under feedlots ranged from none to  more than 5,604 kg/ha in  a 6.1-
 meter profile.  They found that  even though the  ratio  of irrigated
 lands to feedlots was 200:1, calculations based on the average con-
 tent  of the irrigated fields, excluding alfalfa, and the rate of  water
moving through these profiles suggested that  28 to 34 kg of nitrogen
 per hectare were lost  annually to  the water table.  This indicated
 that  feedlots contribute very large amounts of nitrate to  the soil pro-
file with respect to irrigated  land. An important observation is that
feedlots are usually located  near homesteads and  thus  have a pro-
nounced effect on rural water supplies.
     The amount  of nitrate found  under cultivated dry  land was
 significant in relation to historic loss of total nitrogen during cultiva-
tion.  Studies have shown that total nitrogen in dry-land soils de-
creases about 50% during 30 to 50 years of cultivation.  A  large
part  of this decrease  cannot be accounted  for by crop removal.
Lotspeich  et al. (1969)  pointed out the negligible loss of nitrogen in
the Great Plains because of low rainfall.  Losses by volatilization and
erosion were emphasized but also seemed to be minimal.
     Collected  data  suggest  that leaching losses  may  have  been
greatly  underestimated.  There is an accumulation of nitrate in the
2.4- to 3.0-meter depth just below the rooting depth of most dry-land
crops.  The rainfall in the study area averaged about 38 cm per year.
     Stewart et al. (1968) list the chemical data for water samples

                        Chapter 7 / SOURCES OF NITROGEN IN WATER / 113
     	_	  x
      FIG.  7.3.  Feedlot  sources  of  nitrogen  and groundwater.  (Engberg,

 taken from  beneath  feedlots and  adjacent irrigated  fields.  The
 average concentration of ammonium nitrate of the waters beneath
 28 irrigated  fields was 0.2 mg/1. On the other hand,  water from
 beneath  29  feedlots averaged 4.5  mg/1  ammonium  nitrogen.  It
 was also  observed that samples  high in organic carbon  contained
 high amounts of ammonium nitrogen. Nitrite  was usually high un-
 der feedlots. These results indicate the  kinds  and amounts of ma-
 terials moving through soil to groundwater.
     The  importance of well  location with respect to  feedlots was
 demonstrated by Engberg (1967).  In a  study in Holt County,  Ne-
 braska, high nitrate concentrations were observed in domestic wells.
 In Figure 7.3A, an example  is given of undisturbed lateral move-
 ment of high nitrate water in the direction of groundwater move-
 ment. Figures 7.3B and  7.3C illustrate  well pumping  that induces
 movement of high nitrate water  into wells.  Figure 7.3D  illustrates
 a properly located  well that will be free of nitrate. (Also see Fig.  7.4.)
     The  aforementioned  studies make  it apparent  that livestock-
 feeding operations are becoming more concentrated  and  that  their
 effect on  groundwater and surface water  is indeed noticeable  as  a
 source of nitrogen.

     Barnyard Wastes
     The California  State Water Pollution Control Board (1953) re-
ports 1,300 mg/1 of ammonia and  organic nitrogen in percolate
from refuse.

      FIG. 7.4.  Wells A and B yield low nitrate water, C and D yield high
      nitrate water.  (Engberg, 1967.)


     Nitrates are likely to be found in feeds including forages, hay,
 weeds, fodder, silages, or pasture  grasses grown on soils that have
 received heavy applications  of manure or nitrogen fertilizers (Han-
 way et al., 1963).  This is especially true when  drought, shade, dis-
 ease, herbicide applications, or other interfering factors affect normal
 growth and development of  the plant. Nitrate  concentrations are
 highest in immature plants. Stems of plants  concentrate the majority
 of nitrate; intermediate levels are found in the leaves, and low7 levels
 in the grain.

     Domestic Wastewater

   Domestic wastewater effluents range in concentrations from 18 to
28 mg/1 of nitrogen  (exclusive of molecular nitrogen) without spe-
cific treatment for nitrogen removal, according to the Task Group
2610 Report (1967).
     Ammonia nitrogen is the predominant form of nitrogen in efflu-
ents from primary and high-rate treatment plants.
     Although it is not normal to add sewage to a water supply, it
is possible that by filtration  and proper  sewage treatment, waters
that once contained sewage could safely be added to a water supply.
The concentration of nitrogen in a settled domestic sewage is approxi-
mately 80 to 120 mg/1.
     The urban runoff which  accumulated from three streams in an
area containing large reservoirs,  roads, and some logging, but no
human  habitation, is shown in Table  7.7  (Sylvester,  1961).  The

TABLE 7.7.
             Chapter 7 / SOURCES OF NITROGEN IN WATER / 115

Mean  nutrienf concentrations from runoff sources in parts  per

Urban street drainage 	
Urban street drainage (median) .
Streams from forested areas ...
Subsurface irrigation drains 	
Surface irrigation drains 	
Green Lake 	

. 208
. . 154
. . 216
. . 251




Source:  Sylvester (1961).
sources were the Yakima River irrigation return flow drains and the
Green Lake in Washington near Seattle. Nitrate nitrogen levels were
generally  above 200  /*g/l.  The  mean nutrient  concentration was
about 800 /ig/1.
     The average per-capita refuse  originating from food and other
materials imported into the Lake Tahoe watershed is 0.9 kg per day
of which 1% is nitrogen (McGauhey et al.,  1963).
     Nitrogen compounds  used for crop fertilization  and  disposal
of sewage and industrial  wastes  were pinpointed as  sources of
groundwater nitrate (Navone et al., 1963).
     Groundwater Infiltration

     Nitrate movement downward in a silt loam is relatively small
(Herron et al., 1968) and corroborates the fact that there is a  lack
of downward movement of nitrogen from the surface.
     Fifty wells were examined in a 31-square-kilometer area (Behnke
and Haskell, 1968) in the northeastern section of Fresno, California.
Most of the wells were unperforated, open-bottomed casings 27 to 43
meters long.  Behnke and Haskell identified areas of increased nitrate
concentrations on a contour map and drew a nitrate concentration
map of the area.  The background nitrate concentration was 11 to 15
mg/1.  In a zone directly  under the  Clovis sewage treatment plant,
a concentration of 35 mg/1 nitrate  was  found. A second area of
25  mg/1 concentration extended in a zone in a southwesterly direc-
tion and coincided with the strike of the flow line in the same vicinity
as determined by a water-table contour map. A third zone of nitrogen
underlies a  subdivision containing  individual septic  tanks.   The
nitrate concentration again  ranged  from 35 to 25 mg/1.  It  was
noted that nitrate concentrations decreased from 50 to 25 mg/1  in a
lateral distance of eight-tenths of a kilometer.  This study also found
the nitrate concentrations in the top 3  meters of the  groundwater
body was one-third greater than in the rest of  the  water column.


      Leaching  and Erosion

      A study  conducted in  California  (Stout  and  Bureau, 1967)
 showed that a  major part  of  the nitrate reaching  underground
 aquifers was from urban areas and sewage fields.  In addition, it
 was  found that  agricultural crops reduce the amount  of nitrate in
 irrigation water  that returns  to lower soil depths.
      In Missouri, where  the  precipitation normally  exceeds evapo-
 transpiration, soils are acid on the surface. Salts that weather from
 soil minerals are regularly leached and are a normal constituent of
 drainage  water.  Hence,  elements from  the land  that can pollute
 streams are derived more from erosion sediments  than from leach-
 ates. The main sources of sediment that enter water courses are soil
 eroded from urban developments, from highway construction areas,
 and  from  agricultural land.  Losses of  essential mineral nutrients
 during the past half century in the United States have been greater
 from erosion than from  crop removal.  On soils that have low  ex-
 change capacities, leaching of nitrate can be serious.
     Storm Water

     Storm-water runoff, because  of  storm overflows,  can supply
 some nitrogen to rivers.  For example, analysis  of River Erwell at
 Ratcliffe showed that ammoniacal  nitrogen was  in the order of 4.9
 mg/1 and albuminoid nitrogen in the  order of 32.6 mg/1 (Klein et
 al., 1962).
     Storm runoff measured over a year's period  from  an 11-hectare
 residential,  commercial, urban  area indicated that  phosphorus (as
 phosphate) and total nitrogen are  9 and 11%, respectively, of the
 estimated raw sewage  content  from sources in the  area.  At Co-
 shocton, Ohio, two  storms with 5.61 and  12.93  cm of rainfall per
 storm produced runoff of 61.7 to  714 kiloliters per  hectare.  Phosphate
 in the  runoff water ranged  from 0.06  to 0.47 kg/ha and  total
 nitrogen ranged from 0.22 to 6.86 kg/ha (U.S. Public Health Service,

     Several substances containing nitrogen are commonly found in
industrial wastes. For example, ammonia is a waste material from
gas  and coke  manufacturing  and  other chemical manufacturing
processes.  Cyanide is evolved during gas manufacture, plating, case
hardening, and  metal cleaning.  Nitrogen compounds also originate
from explosive factories and other chemical works.
     In  November 1966  a severe ammonia infestation  reached the
Becva River in Czechoslovakia.  The pollution was caused by a chem-
ical  plant  where the equipment for the  production of granulated
superphosphate  was taken  out  of  operation.  Simultaneously, the
ammonia water storage  tanks were  cleaned.  During cleaning, the

                        Chapter 7 / SOURCES OF NITROGEN IN WATER / 117

inlet to the chemical sewage systems became clogged, and the water
overflowed into  the normal sewage system.  This resulted in severe
contamination of the river and led to poisoning of fish along a 20-
kilometer length from  the site  of entrance (Dockal and  Varecha,
     A study conducted on Lake Norrviken in central Sweden during
the years  1961  to 1962 indicated that  wastewater from  a yeast
factory was  responsible for more than  80% of the  nitrogen and
70%  of the phosphorus found in the lake.  Only about 40% of the
nitrogen and  50%  of the phosphorus input to  the  lake leaves the
lake through  its outflow. The rest of the phosphorus accumulated
in the sediments, but a large fraction of  the nitrogen was presumed
to undergo denitrification to free nitrogen. On the basis of the ratio
of the content of nitrogen to  phosphorus in surface  sediment, this
fraction was calculated to be about 37%   of the  total nitrogen input
to the lake or 60%  of the amount which does not leave the lake
through the outflow (Ahlgren, 1967).

     In areas never touched by man for purposes of building or culti-
vation, nitrate from natural  deposits and normal decomposition of
organic matter is present in soil profiles  and groundwaters. Nitrate
accumulates in salty  areas of semiarid and arid regions where sur-
face waters  evaporate.  Also irrigation without adequate  drainage
accelerates nitrate  accumulation.  These  natural  sources cannot be
neglected in any appraisal of a nitrate infiltration problem.  Unless
such additions of nitrogen to a basin or  watershed are balanced by
withdrawals  or denitrification losses, soluble nitrogen  will accumu-
late in surface and soil profiles.
     In many cases, natural sources of nitrogen are sufficient to cause
large nitrogen inputs  into an area. A good example of  this situation
was  reported  by  Frink  (1967) wherein the nutrient input from  a
largely forested watershed with  no overt source  of pollution was
found  to  be adequate to  support abundant  vegetative growth.  In
addition,  a  reservoir  was  noted in which the upper centimeter of
the bottom sediment of a lake contained at least 10 times the esti-
mated  annual input of nitrogen and phosphorus to  the lake. A nitro-
gen budget  of this  lake indicated that in  kilograms  of nitrogen per
lake, the annual input from  the watershed was  30,700; the  output
from the lake was  27,500, resulting in a mean  net input  of 3,200
kg/yr for the lake.

     Ponded water also received nitrate from the  sources mentioned
above. Ponded water that contains abundant algal or other growth,
however, has  less nitrate than water  which does not contain this
growth.  Apparently the  plant growth  uses  excess nitrate  about  as


 rapidly as the nitrate enters  the pond.  Ponded  water also loses ni-
 trate  by denitrification  and  anaerobic decomposition  of  organic
 matter in the ponded water.  This nitrate may eventually escape as
 molecular nitrogen.
     Autotrophic  organisms,  particularly  those  of  green  algae,
 metabolically produce hydroxylamine in highly eutrophicated water
 (Koprivik and Burian, 1966).  This biological origin of hydroxylamine
 is confirmed by the fact that higher concentrations  were observed
 at night than during the day. The occurrence of hydroxylamine in
 pond water is influenced by its relation to the oxygen content.  The
 authors find  great differences between origin and existence of this
 chemical in  water.
     Infiltration from Ponds

     Passage of water through the ground by infiltration from ponds
 has little or no effect on nutrient concentrations in the Tahoe Basin
 near Lake Tahoe, Nevada (McGauhey et al., 1963), and subsequent
 flow through  the ground affects only partial removal  of  nutrients.
 Nitrate appears to be transported by groundwater without significant
 reduction  by   earth materials.   Percolation of  water  through  the
 ground dees materially reduce the concentrations of  other chemical
 constituents.   Wisconsin  stabilization  ponds indicate  annual  per-
 capita contributions of 1.9 kg of inorganic  nitrogen.  Nine  Springs
 sewage treatment plant, serving a 135,000 population with primary
 and secondary filtration,  had an annual per-capita contribution  of
 3.9  kg inorganic nitrogen.  In  contrast, surface runoff in  one in-
 stance was found to contain 43 kg/ha of inorganic nitrogen on a
 20% slope and 20 kg/ha on an 8% slope (Eck et al..  1957).  The
 annual contribution of inorganic nitrogen per  hectare of drainage
 area loading Lake Monona was 4.9 kg, Lake Waubesa 5.5 kg, and
 Lake Kegonsa  7.2 kg (Sawyer et  al., 1945).  (See Table 7.8.)


     Nitrate concentration in stream water from an experimentally
deforested watershed (Likens et al.,  1969) increased from 0.9  mg/1
before removal of the vegetation to 53 mg/1 two years later.  The
nitrate mobilization was attributed to increased microbial nitrifica-
tion and was  equivalent to all the  other  net cationic increases and
anionic decreases observed in the  drainage water of the Hubbard
Brook Experimental Forest in central New Hampshire.

     Enlargement  of  the  Chesapeake and Delaware  Canal and its
approaches required that  large  volumes of bottom material be re-

TABLE  7.8  Distribution of nitrogen in  ponds.

Org.-N (mK/1) .
NIL-N (ms/1) •
NO.-.-N (mg/1) .
Al^uc (no. /ml)
. . 26.3
. . 32.4
. . 0.0
. . Nil
7 X 10 :
7.1 X 10':
2.4 x 107
1.4 x 10;
1.2 X 10"
7.2 /. 10:
2.9 x 10
8.5 / 10
 Source:   Parker (1962).


 moved from channel areas and relocated.  In  the  Chesapeake  ap-
 proach  area,  about  7.4 x 10° cubic  meters of  silt  and clay were
 scheduled to be dredged.  As a result  of data gathered and projected
 by Biggs, it was concluded that  such  action would increase the total
 phosphate and nitrogen by a factor of 50 or  100 over ambient levels
 in the immediate vicinity  of the proposed disposal plant (Biggs,

      It has been shown that there are multiple sources of nitrogen
 to water  supplies.  These include atmospheric, geologic, biogenic,
 rural runoff,  urban runoff, sewage, irrigation, return flow, animals,
 sinkholes, caves, feedlots,  pollen, rural waste, industrial waste, pond
 waters,  deforestation, and land  stripping, among others.
      Generally, salts of nitrogen applied as fertilizer do  not  move
 either vertically or laterally to any significant extent. Movement is
 a function of soil type, soil  saturation,  applied water volume, and
 temperature.   A few  examples  of  such  movement  to  groundwater
 and  surface-water  supplies  and the relative significance of such
 movement were discussed.
      Nitrate in a nonsalt form seems to  have higher soil infiltration
 capacity than salt nitrogen.  This is  dependent, however, upon  the
 physical conditions of the soil and the hydrology  of the  region.
      Other sources not directly used as nutrients to plants, such as
 soil erosion, urban and industrial  wastes, natural  soil nitrogen loss,
 land renovation, deforestation, and atmospheric fallout,  were eval-
 uated with respect  to  their importance  as  a source of nitrogen in
 water.  In general,  industrial waste, rural runoff (including agri-
 cultural land and nonagricultural  land runoff),  farm animal waste,
 and domestic waste are the dominant sources in  surface  waters.
     In  groundwater supplies, specifically wells, the usual sources
 of nitrogen are feedlots, privies,  septic tanks, or other waste forms.
 A few examples were  given  of  geologic  sources within the aquifer
 either as nitrate deposits or nitrate minerals.
     Dissemination of nitrogen from  a plant-nutrient source  is  de-
 pendent upon the geology  and hydrology  extant at the nitrogen
 origin.  With  sufficient data describing these variables  it should be
 possible to characterize the potential for retention or loss of nitrogen
 from the point of origin and the possibility of entrance into a surface
 or groundwater supply. Certainly, with the varied sources of nitrogen
 now  available, and the increasing amount of man-made nitrogen
 materials added to the environment each year, a careful check is
 necessary  on  the  amounts and  sources of nitrogen entering  water

Ahlgren, I.  1967.  Limnological studies of Lake Norrviken, a eutroph-
     icated Swedish lake.  Schweiz. Z. Hydrol. 29:54-90.
Behnke,  J. J., and Haskell, E. E., Jr.  1968. Ground water  nitrate

                         Chapter 7 / SOURCES OF NITROGEN IN WATER / 121

     distributions beneath Fresno, California.  /. Am. Water Works
     Assoc. 60(4): 477-80.
 Biggs, R. B.  1968. Environmental effects of overboard spoil disposal.
     J. Sanit. Eng. Div. Am. Soc.  Civil Engrs. 94 (SA-3): 477-87.
 Brown, M. A., and Volk, G. M. 1966.  Evaluation of ureaform fertil-
     izer using nitrogen-15 labeled materials in sandy soils. SoiZ Sci.
     Soc. Am. Proc. 30 (2): 278-81.
 Calif.  State Dept. of Public Health, Bur.  of Sanit. Eng. 1963.  Occur-
     rence of nitrate in ground water supplies in southern California.
 Calif.  State Water Pollution Control Board.  1953.  Field investiga-
     tion of waste water reclamation in relation  to  ground  water
     pollution.  Calif. State Water Pollution Control Board Publ. 6.
 Carter, D. L., Bobbins, C. W., and Bondurant, J. A. 1969.  The effects
     of irrigation on water quality  and  pollution  in  south central
     Idaho.  In Western  Soc.  Sot7 Sci., 1969 Meetings, Wash. State
     Univ.,  Pullman.
 Comly, H. H. 1945.  Cyanosis in infants caused by nitrates in well
     waters. /. Am. Med. Assoc. 129:112-16.
 Dockal, P., and Varecha, A. 1967. Destructive pollution of the Becva
     River by ammonia.  Vodni Hospodarstvi 17 (9): 388-91.
 Domogalla, B. P., Juday, C., and Peterson, W.  H. 1925.  The forms
     of nitrogen  found in certain lake waters.  /.  Biol. Chem. 63:
 Doneen, L. D.  1968. Effects of soil salinity and  nitrates on tile
     drainage in San Joaquin Valley, California. Water Sci. and Eng.
     Paper 4002.   Sacramento, Calif.  (1966) and San Joaquin Master
     Drain, Appendix Part C.  Fed. Water Pollution  Control Admin.,
     Southwest Region.
 Dugdale, V. A., and Dugdale, R. C.  1965.  Nitrogen metabolism in
     lakes.  III. Tracer studies of the assimilation of inorganic nitro-
     gen sources.  Limnol. Oceanog.  10(1): 53-57.
 Eck, P.,  Jackson, M. L., Hayes, O. E.,  and Bay, C. E.  1957. Runoff
     analysis as a measure of erosion  losses and potential discharge
     of minerals  and organic  matter into lakes and streams.  Sum-
     mary Rept. Lakes Investigation, Univ. of Wis.
 Ehrlich,  G.  G., and Slack, K.  V.  1969.  Uptake and assimilation of
     nitrogen in microecological systems. Am. Soc. Testing Materials.
     Spec. Tech. Publ. 448, pp. 11-23.
 Engberg, R. A.  1967. The nitrate hazard in ivell water. Nebr. Water
     Survey  Paper 21.  Univ. of Nebr. Conserv. and Survey Div.
 Feth, J.  H.   1966. Nitrogen compounds  in water-A review.  Water
     Resources Res. 2(1): 41-58.
 	.  1967.  Chemical characteristics of bulk precipitation in the
     Mojave Desert Region, California. U.S. Geol. Survey Prof. Paper
     575-C, pp. 222-27.
 Frink,  C. R.  1967. Nutrient budget:  rational analysis of eutrophi-
     cation in a Connecticut lake. Environ. Sci. Tech. 1 (5): 425—28.
 Grill, E.  V., and Richards, F.  A.  1964.  Nutrient regeneration from
     phytoplankton decomposing in sea water. /.  Marine Res. 22 (1):
 Hanway, J. J., Herrick, J. B., WiUrich, T.  L., Bennett,  P.  C., and
     McCall, J. T. 1963.  The  nitrate problem.  Agronomy 615:1.
	.  The nitrate problem. Iowa State Univ. of Sci. and Tech. Spec.
     Rept. 34.
Herron, G. M., Terman, G. L., Drier, A.  F., and  Olsen, R. A.  1968.


      Residual nitrate nitrogen in fertilized deep loess-derived soils.
      Agron. J. 60:477-82.
 Hutchinson, G. E. 1954.  The biogeochemistry  of the terrestrial at-
      mosphere.  In The earth as a planet, ed. G. P. Kuiper, pp. 371—
      433.  Chicago: Univ. of Chicago Press.
 Hutchinson, G.  L., and Viets, F. G., Jr.  1969.  Nitrogen enrichment
      of  surface water by absorption of  ammonia volatilized from
      cattle feedlots. Science  166 (3904):  514-15.
 Inglish, H. J.  1967. Nitrates in  private water supplies in Morgan
      County, Missouri.  Milk Food Technol. 30 (7): 224-25.
 Johnston,  W. R.,  Ittihadieh, F., Daum, R.  M., and Pillsbury, A. F.
      1969. Proc. SoilSci.  Soc., p. 287.
 Junge, C.  E.  1958.  The distribution of ammonia  and nitrate in
      rainwater over the United States. Trans.  Am. Geophys.  Union
      39(2): 21-248.
 Keller, W. D., and Smith, George E.  1967. Ground water contami-
      nation by dissolved nitrate. Geol. Soc. Am. Spec. Papers 90:48-
 Klein, L., Jones, J. R. E.,  Hawkes, H. A.,  and Downing, A. L.  1962.
      River pollution. II. Causes and effects. London:  Butterworth.
 Koprivik, B.,  and  Burian, V.  1966.  Origination and  occurrence of
      hydroxylamine in pond water. Cesk.  Hygiena 11 (5): 268—75.
 Koyama, T., and Tomino,  T.  1967.  Decomposition process of organic
      carbon and nitrogen  in lake water. Geochem. ]. 1 (3): 109—24.
 Krause, H. H., and Batsch, W.  1968. Movement of fall-applied nitro-
      gen in sandy soil.  Can. J. Soil Sci. 48:363-65.
 Likens, G.  E., Bormann, F. H., and Johnson, N. M.  1969.  Nitrifica-
     tion:  importance  to nutrient  losses from a cutover  forested
     ecosystem.  Science 163 (3872): 1205-6.
 Loehr, R. C.  1969.  Animal wastes, a national problem. /. Sanit. Eng.
     Div. Am. Soc. Civil Engrs., 95 (SA-2):  189-220.
 Lotspeich,  F.  B., Hauser,  V.  L., and Lehman, O. R.  1969.  Quality
     of water from play as on the southern High Plains.  Water Re-
     sources Res. 5 (1):  48-57.
 McGauhey, P. H., Eliassen, R., Rohlich, G., Ludwig, H. F., and Pear-
     son, E. A.  1963.  Comprehensive study on protection of water
     resources of  Lake  Tahoe Basin through controlled  waste  dis-
     posal.  Prepared for  the  Board of Directors,  Lake Tahoe Area
     Council, Al Tahoe, Calif.
 McGuinness, J. L., Harrold, L. L., and Dreibelbis, F. R.  1960.  Some
     effects of land use and treatment on small single crop water-
     sheds.  /. Soil Water  Conserv. 15 (2): 65-69.
 Macrae, I.  C., Rosabel, R. A., and Salandan, S.  1968. The fate of
     nitrate nitrogen in some tropical soils following submergence.
     SoilSci. 105(5): 327-34.
 Matheson,  D.  H. 1951.  Inorganic nitrogen in precipitation and  at-
     mospheric sediments. Can. J. Technol. 29:406-12.
 Metzler, D. F., and Stoltenberg, H. A. 1950. The public health sig-
     nificance of high nitrate waters as a cause of infant cyanosis
     and methods of control.  Trans. Kansas Acad. Sci. 53:194-211.
 Meyer, J., and Pampfer, E. 1959. Nitrogen content of rainwater col-
     lected  in the humid central Congo Basin. Nature 184:717.
 Miller. R. B. 1961.  The chemical composition of rainwater at  Taita,
     New Zealand, 1956-1958. Nezv Zealand J. Sci. 4:844.
Mink, J. F. 1962. Excessive irrigation in the soils and ground water
     of Oahu, Hawaii. Science 135 (3504):  672-73.

                        Chapter 7 /  SOURCES OF NITROGEN IN WATER / 123

Moe, P. G., Mannering, J. V., and Johnson, C. B. 1967.  The loss of
    fertilizer nitrogen in surface runoff water. Soil Sci. 104 (6): 389—
	.  1968.  A comparison of nitrogen losses from urea and am-
    monium nitrate in surface runoff water.  Soil Sci. 105 (6): 428—
Navone, R., Harmon, J. A., and Voyles, C. F. 1963.  Nitrogen content
    of ground water  in southern  California. /.  Am.  Water Works
    Assoc. 55 (5): 615-18.
Olsen, R. J., Hensler, R. F., Attoe, O. J., Witzel, S. A. 1969. Effect of
    fertilizer nitrogen, crop rotation and other factors  on amounts
    and movement of nitrate nitrogen through soil profiles.  Agron.
    Abstr. Am. Soc.  Agron., 61st Annual Meeting, p. 104.
Overrein, L. N. 1968.  Lysimeter studies on tracer nitrogen in forest
    soil. I. Nitrogen losses by leaching and volatilization after addi-
    tion of urea-N15.  Soil Sci.  106  (4): 280-90.
Owens, M., and Wood, G. 1968. Some aspects of the eutrophication
    of water.  Water Res. 2:151-59.
Paloumpis, A. A.,  and Starrett, W.  C. 1960. An  ecological study of
    benthic organisms in the  three  Illinois river flood  plain lakes.
    Am. Midland Naturalist 64(2): 406-35.
Parker, C. D. 1962.  Microbiological aspects of lagoon treatment.  J.
    Water Pollution Control Federation 34:149-61.
Peterson, H. B., Bishop, A. A., Law,  J. P.,  Jr.  1969.   Problems of
    pollution  of irrigation waters  in arid regions.  In  AAAS  inter-
    national conference on arid lands in a changing world (preprint).
Pratt, P. F., Cannell,  G.  H., Garber, M. J., and Blair,  F. L.  1967.
    Effect of three nitrogen fertilizers on gains, losses, and distribu-
    tion of  various  elements  in  irrigated  lysimeters.  Hilgardia
    38 (8): 277.
Preul, H. C., and Schroepfer, G. J.  1968. Travel of nitrogen in soils.
    /. Water Pollution Control Federation 40 (1): 30-48.
Putnam, H. D., and Olsen, T. A. 1959.  A preliminary investigation of
    nutrients in western Lake Superior, 1958-1959.  School of Pub-
    lic Health, Univ. of Minn.
	.  1960.  An investigation of nutrients in ivestern Lake Superior.
    School of Public Health, Univ. of Minn.
Sawyer, C.  N., Lackey, J. B., and Lenz, R. T. 1945. An investigation
    of  the  odor  nuisances  occurring in the Madison lakes, par-
    ticularly Monona, Waubesa and Kegonsa from  July  1942 to
    1944.  Report of Governors Committee. Madison, Wis. 2  vol-
Sinha, H.,  and Prasad, K. 1967. Performance, transformation and
    movment of urea in acid  soils.  J. Indian Soc. Soil Sci. 15(4):
Smith, G. E.  1967.   Fertilizer nutrients  as contaminants in  water
    supplies.  In  Agriculture and  the quality of our  environment.
    Publ. 85, pp. 173-86. Am. Assoc. for  the Advancement of Sci.
Stewart, B. A., Viets, F. G., Jr., Hutchinson, G. L., Kemper, W. D.,
    Clark, F. E., Fairbourn, M. L., and Strauch, F. 1967. Distribution
    of nitrates and other water pollutants under fields and corrals in
    middle South Platte Valley of  Colorado. USDA, ARS 41-134.
Stewart, B. A., Viets, F. G., Jr., and  Hutchinson, G. L. 1968. Agricul-
    ture's effect on  nitrate pollution of ground water.  /. Soil Water
    Conserv. 23 (1): 13-15.
Stout, P. R., and Burau, R. G.  1967. The extent and significance of


     fertilizer build-up in soils as revealed by vertical distribution of
     nitrogenous matter between soils and  underlying  water reser-
     voirs.  In Agriculture and quality of our environment.  Publ. 85,
     pp. 283—310.  Am. Assoc. for the Advancement of Sci.
 Sylvester, R. O. 1961. Nutrient content of drainage water  from for-
     ested,  urban and  agricultural areas.  Algae  and Metropolitan
     Wastes. U.S. Public'Health Serv.  SEC TR  W61-3, pp.  80-87.
 Sylvester, R. O., and Seabloom, R. W. 1963.  Quality and significance
     of irrigation return flow.  /. Irrigation Drainage Div.  Am. Soc.
     Civil Engrs. 89  (IR-3).  Proceedings  Paper 3624, pp.  1-27.
 Task Group 2610-P Report.  1966.  Nutrient-associated  problems in
     water  quality and treatment.  /. Am. Water  Works  Assoc.
     58 (10): 1337-55.
 Task Group 2610-P Report.   1967.  Sources  of nitrogen and phos-
     phorus in water  supplies.  J. Am. Water Works Assoc. 59:344—
 Timmons, D. R., Burwell, R. E., and Holt, R. F. 1968. Loss of crop
     nutrients through runoff.  Minnesota Sci. 24 (4): 1.
 U.S. Geological Survey, Water Resources Division.  1965. Conference
     on Nitrogen Chemistry,  1965. Menlo Park, Calif., Sept.  21-22.
 U.S. Public Health Service. 1964. Basic and Applied Sciences Branch,
     Division of Water  Supply and Pollution  Control. Activities Re-
     port July 1, 1963-June 30, 1964.
Weidner, R. B., Christiansen, A. G., Weibel, S. R., and Robeck, G. G.
     1969.   Rural runoff as a  factor in stream pollution.  /. Water
     Pollution Control Federation 41 (3): 377-84.
White, A. W., Burnett, A. P., Jackson, W.  A., and Kilmer, V. J. 1967.
     Nitrogen fertilizer  loss in runoff from crop land tested.  Crops
     Soils 19 (4): 28.
Whitehead,  E. L, and Moxon, A. L. 1952.  Nitrate poisoning. S. Dak.
     State College Bull. 424.


      I  HE importance of N from the standpoint of soil fertility has
long  been recognized, and our knowledge concerning the nature,
distribution, and transformations of N compounds in soil is exten-
sive.  Early work dealt largely with practical aspects of maintaining
a reserve of humus N for plant growth; more recently, interest has
been centered on the efficient use of fertilizer N. Increasing attention
is now being given to problems associated with the disposal of nitrog-
enous wastes on farmland and of the fate of applied N  as related
to water quality.
     A schematic diagram  depicting the cycle of N in soil is given
in Figure  8.1.   Ammonium (NH4+) added  as fertilizer, or formed
from decay of plant and  animal residues, is temporarily held by
the exchange complex of the soil but is eventually oxidized to nitrate
(NO3-) unless it becomes fixed by humus  or clay minerals.  Immo-
bilization by microorganisms leads  to conversion of NH4+ and  NO3~
to the humus form.  The NO3~ is subject to leaching,  and it can be
converted to gaseous products through a process called denitrification.
Losses of N can also occur through chemical reactions involving ni-
trite (NO.,-).
     The  above  considerations emphasize  that a close relationship
exists between inorganic and organic forms of N, and that the sub-
ject of soil N deals not only with the nature and distribution of vari-
ous inorganic and organic  compounds but  of their interactions with
each other and with mineral matter. An understanding of the chem-
istry of soil N complexes,  and of  the reactions  they undergo,  is of
considerable importance  from the standpoint of evaluating agricul-
tural practices as they relate to the occurrence of NO,- and nitrog-
enous organic  substances in natural waters.
     The purpose of this chapter is to summarize our knowledge of
the kinds and amounts of  N compounds in soil.  Brief mention will
be made  of  chemical transformations involving NH4" and  NO2~.
The  subject of biological transformations will be mentioned only as
     F. J. STEVENSON is Professor of Soil Chemistry, Department of Agron-
     omy, University  of Illinois. G.  H. WAGNER is Associate Professor,
     Department of Agronomy, University of Missouri.

S, *

.Xl\ON —
1^1 MATTER fnATTtn
     FIG.  8.1.   The N  cycle in soil. (From Stevenson, 1965.)

far as it  contributes  to our understanding of the  chemistry of  soil

     Examination of Figure 8.1 shows that several mineral forms of
N other than NH4+ and NO3- are possible in  soil.  They included
nitrite (NO2-),  elemental N (N2), and nitrous oxide (N2O).  Nitrite
and NoO, along with nitric oxide (NO) and nitrogen dioxide (NO2),
can be found in soil only under very special circumstances (see sec-
tions dealing with  denitrification and nitrite reactions).  Other in-
organic N compounds, such as hydroxylamine  (NH2OH) and hypo-
nitrous  acid (HON = NOH), may occur as intermediates in biological
transformations of N  but for the most  part they  are unstable and
have  only a transitory existence.  Elemental N is a  common con-
stituent of  soil air; unfortunately, it cannot be  used directly by

                      CHAPTER 8 / CHEMISTRY OF NITROGEN IN SOILS /  127

     Although plants are capable of utilizing organic N compounds
(for example, amino acids), practically all of the  N taken up from
the soil exists in inorganic forms (as NH,+ and NO:; ).
     Exchangeable NH4h  and NO:f

     Several recent reviews (Bremner,  1965;  Harmsen  and Kolen-
brander, 1965; Stevenson,  1965) have emphasized that only a small
fraction of the N in  soils, generally less than  0.1%, exists in avail-
able mineral compounds (as exchangeable NH4+ and NO3~).  Thus,
only a few pounds of N may be  available to the plant  at any one
time, even though 2 or 3 tons  may be  present in combined forms.
The slow  conversion of nitrogenous organic substances to available
mineral forms by microorganisms has been attributed to their sta-
bilization  by ligninlike substances and to the protective action of
clay minerals. The formation of stable complexes can be considered
beneficial, because the N  is protected against decomposition  and
subsequent leaching  as NO;?-.
     Levels of exchangeable  NH4+ and  NO:5- vary from  day to day
and from  one season to another, and will depend upon such factors
as climate (temperature, rainfall), organic matter content, presence
or absence of  growing plants, C/N ratio of added residues, and time
and rate  of application of nitrogenous fertilizers.  Some important
aspects regarding available N in soils are itemized below.

1.  The quantity of available N in unfertilized soil at any one time
    is markedly influenced by climatic patterns (Harmsen and van
    Schreven,  1955;  Harmsen  and Kolenbrander,  1965).  For ex-
    ample, in soils of the temperate humic climatic zone,  the content
    of inorganic N in  the  surface layer is lowest in winter due to
    leaching,  rises in  spring as mineralizat'on  of  organic  N  com-
    mences, decreases in summer through consumption  by plants,
    and increases once again in the fall when plant growth ceases
    and the dead residues start  to decay. The level in winter seldom
    exceeds 10 ppm but may increase 4- to 6-fold or more during the
    spring (Harmsen  and van  Schreven,  1955).  The  winter mini-
    mum is usually ascribed to leaching.
2.  Biological  turnover leads to  the interchange  of  NH4*-N  and
    NO3"-N with the N locked up in organic forms.  Accordingly, the
    amount of mineral N in. the soil at any one time represents a
    balance between the opposing processes of mineralization  and
    immobilization, and will be determined to a large extent by the
    activity of the  soil  microflora and the  C/N ratio  of decomposing
    residues. A C/N ratio above a. critical value of 20 to  25 (equiva-
    lent to 1.5 to 2.0%  N)  results  in a net immobilization  of N
    whereas a  ratio below this value leads to net miner ah" zation.
3.  Growing plants have a depressing effect on  the level of mineral
    N in soils.  The decrease when soils are cropped cannot be ac-
    counted for entirely  by plant  uptake,  and may be  due to one
    or more of the following: (1) inhibition of nitrification by excre-
    tion products  of plant roots, (2) immobilization of mineral N by

 FiG. 8.2.  Nitrate-N in the
 upper 8 feet of 4 soil types
 after the  annual applica-
 tion of N  fertilizer for  7
 years fo continuous corn in
 Missouri.  (Adapted  from
 Smith,  1968.)
                       CHAPTER 8 / CHEMISTRY OF NITROGEN IN SOILS /  129
               NOs (ppm)
              2OO      4OO
       \     /

        \  /

                                           FIG.  8.3.   Distribution  of

     (Micrococcus denitrificans; Thiobacillus denitrificans) are also capa-
     ble of converting NO3~  to N2 but they are not believed to be im-
     portant in most soils.
          The following pathway represents  the probable mechanism of
     bacterial denitrification.

                                                   + 2H
                                      N = N =O	»N = N
+ 2H
- 2H.O
         x, O    + 4H                        + 4H
2HO - N ^  	> 2HO - N = O 	> (HO _ N ~ N - OH)
         ^ O    - 2H,O                       - 2H,O
Nitrate                      Nitrite                        Hyponitrite

         Nitrous oxide represents an intermediate in the  denitrification
     process and is  normally reduced further to  N2; consequently, the
     N20 has  only a  transitory existence in the soil.
         Optimum conditions for denitrification are as follows:

     Poor drainage:   Moisture status  is of importance from the stand-
         point of its effect  on  aeration. Denitrification is negligible at
         moisture levels below  two-thirds of the water-holding capacity
         but is appreciable  in flooded soils.  The  process  may occur in
         anaerobic microenvironments of well-drained soils, such as small
         pores filled  with  water, the rhizosphere of  plant roots, and the
         vicinity of decomposing plant and animal residues.
     Temperature of 25° C and  above: Denitrification proceeds at a pro-
         gressively slower rate at temperatures below 25° C and  practi-
         cally ceases at 2° C.
     Soil reaction near neutral: Denitrifying bacteria are sensitive to high
         hydrogen  ion  concentrations.  Their  activity in  acidic  soils
         (< pH 5) is  limited.
     Good supply of readily  decomposable organic matter:  The amount
         of  organic  matter  available  to denitrifying microorganisms is
         generally appreciable in the surface horizon  but  negligible  in
         the subsoil.  Significant amounts of  soluble organic matter may
         be  found under feedlots, as well as in the  lower horizons of soils
         amended with  large quantities of organic wastes.

         Denitrification  can be considered a desirable process when it
    occurs below the  rooting zone, because of reduction in  the NO3- con-
    tent of  groundwater.  Dentrifying microorganisms  are known  to be
    present  at considerable depths in soil, and it is possible that some of
    the NO3- leached into  the subsoil may be volatilized before reaching
    the water  table.  Meek et al. (1969) concluded that much of the NO3-
    leached into the subsoil in irrigation waters was  lost through denitri-
    fication.  Stewart et  al. (1967) found that NO-r  levels  in soil under

                       CHAPTER 8 / CHEMISTRY OF NITROGEN IN SOILS / 131

 feedlots decreased sharply with increasing depth and concluded  that
 the decrease was due to denitrification.
     Under optimum conditions,  NO...--N can be volatilized  quanti-
 tatively in a comparatively short time (24 to 36 hours).  This suggests
 that  the  denitrification process can be xitilized  to  eliminate excess
 NO3- from soil,  thereby  reducing the  NO3- content  of percolating
 water.  For example, the  disposal  of nitrogenous wastes on farmland
 results  in the generation  of large  quantities of NO:i-, which must be
 removed  if  groundwater contamination is  to be avoided.  In fine-
 textured soils, reduction in NO:t-  content could  be  accomplished by
 artificially subjecting the soil to  successive  cycles of submergence
 and drying.  The anaerobic  conditions  created during waterlogging
 would result in a significant loss  of NO3- produced by oxidation of
 NH4+ during the  aerobic cycle. A  similar procedure may prove effec-
 tive in  reducing  the NO:!~ content of soil under  feedlots.
     Fixation  of  NH4+ by Clay  Minerals

     The NH4- produced in soil through microbial activity, or added
 as fertilizer, can  be fixed by clav minerals (Nommik, 1965\  Fixation
 results from a replacement of NH4+ for interlayer cations that expand
 the lattice (Ca2+, Mg2+, Na% H+), but not by those that contract  the
 lattice (K>, Rb\ Cs+). Soils containing large amounts of vermiculitic-
 or illitic-type minerals have the capacity for fixing 1 to 6 m.e. of NH4+
 per 100 g, or  from about 280 to 1,680 Ib per acre plow depth.  Prac-
 tically no fixation will occur when the clay fraction is predominantly
     The availability of NH4+ to both nitrifying microorganisms and
 higher plants can be reduced by fixation. However, various studies
 have shown that  fixation is usually not a serious problem under nor-
 mal fertilizer practices. Potassium, being a fixable cation, is effective
 in blocking the release of fixed NH.,+; thus, the  application of large
 amounts of K+ simultaneously or immediately following an NH4+ addi-
 tion may diminish the availability of the fixed NH4+ to higher plants.
     Naturally Occurring Fixed NH4*

     For many years, soil scientists assumed that the major inorganic
forms  of N in soils were exchangeable NH4+  and N0-r.  Now it is
known that soils contain fixed NH4+—that is, NH4+ held within  the
lattice structures of silicate minerals.  Present estimates are that 4 to
10% of the N in the surface layer of the soil occurs  as fixed NH4+.
The proportion generally increases with depth, and in some subsoils
as much as 50% of the N may exist in this form.
     The distribution of  fixed NH4+ in representative soils of several
great soil eroups is  shown in Figure 8.4.  With the exception of  the
Pcdzols,  the  A, horizons contained about  60 to 150  ppm  of fixed
NH^-N, equivalent to about 120 to 300 Ib of N per acre plow depth
of soil.  The rooting zone may contain as much as 1,600 Ib of N  per

 FIG. 8.4.  Distribution   of
 fixed NrV-N  in soils rep-
 resentative   of   several
 "great    soil   groups."
 (Adapted  from Stevenson
 and Dhariwal, 1959.)
                                      FIXED NH4-N,PPM

                                         100        200
 acre as fixed NH4+. The fixed NH4+ content is related to clay mineral
 composition; soils rich in micaceous (illitic) types contain the largest
     The  proportion of  the soil  N as fixed NH4+ increases slightly
 when soils are cropped,  indicating that the native fixed NH4+ is less
 available to plants and microorganisms than the humus N.  Increases
 in the content  of fixed NH4+ have been reported through N fertiliza-
 tion (Harmsen  and Kolenbrander, 1965).
     Fixation  of  NH3 by Organic Matter

     It is well known that NH3 can be "fixed" by reaction with lignins
and humic substances (Mortland and Wolcott, 1965; Broadbent and
Stevenson, 1966).  Fixation is associated with oxidation (uptake of
oxygen) and is favored  by an alkaline reaction.  Thus, the applica-
tion of alkaline  fertilizers such as aqueous- or anhydrous NH3 to
soil may result in considerable  fixation.  The NH3 fixed by organic
matter is not immediately usable by plants, although it does become
available eventually through the mineralization process.
     The nature  of the reaction of NH3 with  soil humus is not
known. It is believed, however,  that aromatic compounds containing
two or more hydroxyl groups  are involved.  The initial step involves
the consumption of oxygen and the formation of a quinone, which
subsequently reacts with NH3 to form complex polymers.  Catechol
(I), for example, is readily converted in alkali  to  o-quinone  (II),
which can be hydrated  to form  benzenetriol (III)  (see Mortland and
Wolcott, 1965).  Further oxidation yields  o-hydroxyquinone (IV) and
p-hydroxy-o-quinone (V).

                       CHAPTER 8  / CHEMISTRY OF NITROGEN IN SOILS  / 133
     The incorporation of NH3 into p-hydroxy-o-quinone (V) is postu-
lated to produce structures of the types represented by VI and VII.

     Nitrite  is  not usually  present in  detectable amounts in well-
drained neutral or slightly acidic  soils.  Accumulations  occur, how-
ever, in  calcareous  soils, and recent work  indicates that this  ion
often persists, albeit temporarily, when NH4+- or NH4+-type fertilizers
are applied to soil.  This NO2- accumulation has been attributed to
inhibition of nitrification at the NO2~ stage.  Presumably, NO2~  oxi-
dizing organisms  (Nitrobacter)  are more sensitive  to NH3  and an
adverse soil reaction than NH4+ oxidizers (Nitrosomonas). According
to Hauck and Stephenson (1965), large fertilizer granules, high appli-
cation rates,  and an alkaline pH in the zone of fertilization are par-
ticularly favorable for NO2~ accumulations.
     The possibility that gaseous loss  of fertilizer N  may accompany
temporary N02~ accumulations  has been mentioned in the reviews
of Allison (1965) and  Broadbent  and  Stevenson (1966).  Classical
reactions  involving  NO

 promote the decomposition of NOLr.  One theory is that organic con-
 stituents are involved (Broadbent  and Stevenson,  1966;  Bremner
 and Nelson, 1968).  Another view is that metallic cations  are respon-
 sible (Wullstein, 1967).

      The organic N in soil consists  of  two main groups of com-
 pounds:  (1) nitrogenous biochemicals synthesized enzymatically by
 microorganisms living on plant and animal residues, and (2) products
 formed  by  secondary  synthesis reactions  and  which  bear  no  re-
 semblance  to any of the substances  occurring in plant and  animal
 tissues.  The N in the second group  probably exists as part of the
 structures of the so-called humic and fulvic acids.  The two  groups
 are not easily separated, because some of the biochemicals (e.g., amino
 acids) may be covalently bound to the humic matter.
     Nitrogenous Biochemicals


     The recent application of chromatographic methods to studies
 of soil N have resulted in the isolation of an impressive number of
 amino acids from soil hydrolysates, and these studies have confirmed
 earlier reports indicating that 20 to 50%  of the organic N occurs in
 the form of amino acids (Bremner,  1965,  1967).  In addition to  the
 20 to 22 amino acids generally found in proteins, a variety of other
 compounds have been  identified, including a-amino-n-butyric acid,
 y-aminobutyric acid,  (3-alanine, a,£-diaminopimelic acid, and 3,4-di-
 hydroxypbenylalanine.  The  occurrence of a,f,-diaminopimelic acid
 is of particular interest because this amino acid appears to  be con-
 fined  to certain bacteria, where it occurs  as a structural component
 of the cell wall. The presence  of ornithine, f5-alanine, and y-amino-
 butyric acid in a variety of natural products is now well established.
    Many unidentified  ninhydrin-reacting substances have also been
 detected  in soil hydrolysates.  Thus far,  over 50 compounds have
 been reported; the identity of the majority has not been established.
 Some of the  amino compounds  may be  artifacts produced during
    The persistence of certain microbially synthesized amino acids
 has been reported by Wagner and Mutatkar (1968) in a study of  the
 humification  of 14C glucose.  The  highest specific activities were
 found in those amino compounds known  to be constituents of cell
 walls of microorganisms (alanine, glycine, glutamic acid, and lysine).
 Glucosamine,  an amino  sugar found in  bacterial and fungal cell
 walls, also contained large quantities of 14C.  The cell  walls of cer-
 tain dark pigmented (melanic) fungi appear to be especially resistant
 to deccmposition (Hurst and Wagner, 1969).
    The reviews of  Bremner  (1965, 1967)  show  that  conflicting
results have been obtained concerning the relative distribution of

                       CHAPTER 8 / CHEMISTRY OF NITROGEN IN SOILS / 135

 amino acids in different soil types and between various horizons of
 the same profile.  Data obtained for the Morrow Plots at the University
 of Illinois  indicate that basic amino acids  are selectively preserved
 through long-time cropping; this trend has yet to be confirmed.  Varia-
 tions in amino acid composition may exist between  soils from  dif-
 ferent climatic regions of the earth.
     Considerable controversy exists as to whether proteins as such
 occur in significant amounts in soil organic matter. The well-known
 ligno-protein theory advanced by Waksman has yet to be confirmed;
 many investigators believe that the theory in  its original form is ob-
 solete.  Swaby and Ladd  (1962) failed to detect proteins in  humic
 acids, using  sensitive chemical tests,  and concluded that  neither
 proteins nor ligno-protein complexes accounted for a  significant part
 of the soil N, On the other hand, results obtained using proteolytic
 enzymes, partial  hydrolytic procedures, and infrared  spectrophotom-
 etry suggests that in some humic acids peptide linkages are present.
 Some of the amino acid-N in soil  may occur as mucopeptides. Free
 amino acids have but a transitory  existence, and the amount of N in
 this form rarely exceeds more than a few ppm.
     The relative importance of clay and humus particles in binding
 amino acids,  peptides, and  proteins is unknown.  However,  for  the
 surface layer of  normal agricultural soils, the role played by  humic
 and fulvic  acids  cannot  be overemphasized.  In argillaceous  sub-
 soils, a significant proportion cf the proteinaceous material may be
 held by clay minerals, perhaps on interlamellar surfaces.

     Several studies have indicated that 4 to lOTc of the N in  the
surface layer of the  soil occurs in the form of N-containing carbohy-
drates, namely, the amino sugars. In some soils, the proportion may
increase with depth.  Amino sugars are widely distributed in microbial
tissues; hence, their presence in soil is to be expected.
     Research conducted at  the University of Illinois indicates that a
wide variety of amino sugars are present in soils, including glucosa-
mine, galactosamine, fucosamine, and muramic acid. The latter is a
common constituent cf the cell walls of bacteria. Free amino sugars
have yet to be found in soils.

     A wide  array of naturally  occurring  nitrogenous compounds
other than amino acids and amino sugars have been found in soils.
but in very low amounts.  They include a variety of amines, several
chlorophyll derivatives, amino acid amides  (asparagine  and gluta-
mine), and purine and pyrimidine bases. All of these compounds
combined account for no more than 1 to 2^  of the soil N (Bremner,
1965, 1967).


      Unknown Forms of Organic  N

      The considerations mentioned in the previous section emphasize
 that no more than one-half of the soil organic N can be accounted for
 as amino acids,  amino  sugars, purine  and pyrimidine bases,  and
 other known compounds.  Since practically all of the organic N in
 soil is of microbial origin,  and because  the N of microbial tissues
 occurs almost exclusively  in the above-mentioned  compounds,  the
 conclusion  seems justified  that during  humification, conversion of
 the microbially synthesized products to more stable humus forms has
 occurred.  The N content of humic  and fulvic acids varies widely,
 values between 0.4  and 5.0% having been reported.
     The relative  distribution of the forms of N in acid  hydrolysates
 of humic and fulvic acids  is illustrated in Figure 8.5.  It as note-
 worthy that as much as one-third of the  N in  humic acids cannot be
 solubilized by hydrolysis  with 6 N HC1; as much  as one-half of that
 in fulvic acids is liberated as NH:V  The nature of this N  is uncertain,
 but most of it may occur as part of the structures  of humic sub-
     In considering the properties of humus N, some discussion of  the
 nature of humic and fulvic acids is  desirable. These constituents can
 best be described as  a series of acidic, yellow- to black-colored, moder-
 ately high-molecular-weight polymers which have characteristics un-
 like any organic compounds occurring in  living organisms. The mod-
 ern view is  that they represent a heterogeneous mixture  of molecules
 which range in molecular  weight from  as low as 2,000 to perhaps
 over  300,000.  Interrelationships between such properties as color,
 elemental composition, acidity, degree of polymerization, and molec-
 ular weight are outlined  schematically in Figure 8.6.  No sharp divi-
 sion exists between  the various fractions.
FIG.  8.5.   Relative  distri-
bution of the forms of  N
in  humic  and fulvic acids.
The broken  portion  of the
bars  indicates the range of
values reported.
      r f

                       A-HUMIC ADDS
                       B - FULVIC ACIDS
                                 Amino ocid-N
 Acid     Amino
insolubte-N  sugar-N

                       CHAPTER 8 /  CHEMISTRY OF NITROGEN IN  SOILS / 137
Fulvic acid
Light yellow
Humic acid
Dark brown
 	increase in degree of polymerization	
 2,000?	increase in molecular weight	1
 45%	increase in carbon content	
 48%	decrease in oxygen content	•
 1, 400	decrease in exchange acidity	
     FIG.  8.6.  Chemical properties of humic and  fulvic acids.  The yellow-
     colored fulvic pigments are relatively mobile and  can act as carriers
     of N in streams and  lakes (see text).  (Adapted from a  drawing by
     Scheffer and Ulrich, 1960.)

     The yellow-colored pigments shown in Figure  8.6 correspond to
the crenic and apocrenic acids of Berzelius, and they are the con-
stituents often found in the colored waters of lakes  and streams.  Be-
cause of their low molecular weights, fulvic acids  are highly mobile
and can migrate through the soil profile in percolating waters.
     The  N  of soil  humic  substances may occur  in the following

1. As a free amino  (—NFL) group
2. As an open chain (-NH-, — N-) group
3.  As part of  a heterocyclic ring, such as  an —NH— of  indole  and
   pyrrole or the —N= of pyridine
4.  As a bridge constituent (see structures VI and VII)

     Very little  is known regarding the  manner whereby N is in-
corporated in humic and fulvic acids, but one or more of the processes
illustrated in Figure 8.7 (and discussed below) are probably involved.

     The  interaction between NH3  and  oxidized lignins has  been
suggested as a possible pathway of humus formation.  The autoxida-
tion of both humic acids and lignin  under alkaline conditions in the
presence of aqueous NH3 yields stable N-containing complexes. Re-
actions of the type  discussed earlier are probably involved.   Part of
the fixed N cannot be solubilized by subsequent acid hydrolysis.

     Many scientists  now support the theory that humic constituents
originate through condensation of quinones with N-containing com-
pounds, such as amino acids. According  to this concept, polyphenols,
either derived from the biological breakdown of lignin or synthesized

                        PLANT RESIDUES
                       HUMIC SUBSTANCES
      FIG. 8.7.   Mechanisms of formation of soil humic substances.  Nitrog-
      enous  substances  (e.g., amino  acids) synthesized by  microorganisms
      during  the decomposition of plant and  animal  residues are seen  to
      react with  modified  lignins (reaction 4), quinones (reactions 2 and 3),
      and reducing sugars (reaction 1) to form complex polymers containing
      N as part  of their structures.

 by microorganisms, are oxidized enzymatically by phenoloxidases  to
 quinones, which  then react with amino acids to form  humic  sub-
 stances.  In  the process, cyclic N compounds are formed.
     Flaig's  (1966) concept of humus formation is as follows:

 1.  Lignin, freed  of its linkage with cellulose during decomposition
    of plant  residues,  is subjected to oxidative splitting with the for-
    mation  of primary structural  units  (derivatives of phenylpro-
 2.  The  side chains  of the  lignin-building  units are  oxidized, de-
    methylation occurs,  and the resulting polyphenols  are converted
    to quinones  by polyphenoloxidases.
 3.  Quinones arising from the lignin (as well as from other sources)
    react  with N-containing  compounds to form dark-colored  poly-

     The  importance of microorganisms as a source of polyphenols
 for humus synthesis has recently been emphasized.  Kononova (1966),
 for example, has postulated that  humic  substances  can be formed
 from polyphenols synthesized by cellulose-decomposing myxobacteria
 in soil.  Many fungi are known  to produce humic acidlike substances.
 According to Swaby  and Ladd (1962) humic molecules  are formed
 from free radicals (quinones) produced enzymatically within deceased
 cells while autolytic enzymes are still functioning but before cell walls
 are ruptured by microbes.

     The formation of brown nitrogenous polymers by  condensation
of carbonyl-containing  compounds  (reducing  sugars)  and  amino

                       CHAPTER 8 /  CHEMISTRY OF NITROGEN IN SOILS / 139

derivatives (amino acids) occurs extensively in stored  food  products
and  the reaction  has been postulated to  occur  in soils.   A major
objection to this theory  is the slow rate  at which sugar-amine  con-
densation reactions  occur.  However, drastic  changes in  the  en-
vironment (freezing and thawing, wetting and drying), together with
the intermixing of reactants with mineral material having catalytic
properties, may facilitate condensation.

     It is well known that the N content of most soils declines when
land is cultivated for the first time. Under average farming condi-
tions in the Corn Belt region of the United  States,  about 25%  of
the N is lost the first 20 years, about 10% the second 20 years, and
about 1% the third 20 years. This loss  of N is not spread uniformly
over all of  the  N fractions.  Long-time  cultivation has been found
to result in  increases in the proportion of the  total N as fixed NH4+,
amino sugars, and hydrolyzable N.  The changes are small, however,
and no single component can be considered to be the  major source
of mineral  N for  plant  growth.  Methods of estimating available N
by analysis  of any given fraction would appear to be unsatisfactory.
     Research conducted at the University of  Illinois  indicates that
when soils  are  cropped those compounds intimately bound  to clay
minerals are selectively preserved. Figure 8.8  shows that the  propor-
tion of the organic  N in  the Morrow  Plots  which was  solubilized
through destruction  of  clay with HF  increased with  decreasing N
content.  Thus, it appears  that  loosely bound  substances are  lost
first, followed in order by those held by strong cohesive forces. The
content of  soluble organic N compounds in drainage  waters would
be expected to be  particularly low in soils from intensively cultivated



* 1
: -

, '
11 ORG.-N


s t



: c
G~"0"Cl MI p

^ ^


' B



i ^

                                            FIG. 8.8.  Organic N  and
                                            NHr  extracted  from  the
                                            Morrow   Plot   soils   by
                                            extraction  with  a  2.5N-
                                            HF:0.1N-HCI solution.  The
                                            values in the solid portion
                                            of  the bars represent the
                                            percent  recovery  of   or-
                                            ganic  N.  C — corn,  O =
                                            oats,   Cl = clover,  MLP =
                                            manure,   lime,  and phos-
                                            phate.   (From   Stevenson
                                            et  al., 1967.)
   %N = 0.128 0.158 0.163  0.135 0.212 0.243 0.290


      Figure 8.8 further  shows that the percentage of the total  N
  as fixed NH4+ was highest in those soils where organic matter had
  been  depleted through intensive cultivation (see section on Naturally
  Occurring Fixed NH4+).

      This brief review has served  to emphasize the  complex nature
 of soil N.  Other than gaseous forms,  the  inorganic N consists pri-
 marily as NH4+ and NO;{-.  Part of the  NH4+  is bound to  colloidal
 surfaces and behaves  according to classical reactions  of exchange
 chemistry. Nitrate is free to  move with the  soil water and is the
 form of N  which  is of greatest  concern from the standpoint of
 pollution of water supplies.  Many  soils contain appreciable  amounts
 of NH4+ that cannot be utilized directly by plants  and microorga-
 nisms; this NH4+ is held within the lattice structures of clay minerals.
      Less  than one-half of the organic N in soils can  be accounted
 for in known  compounds (amino  acids, amino sugars,  purine  and
 pyrimidine bases,  etc.).  The  remainder may  occur  as part of the
 structures of humic and fulvic acids.  Part of  the N added to soils
 as fertilizers can be converted to organic forms by chemical reactions
 involving NH,S and NO2-; this combined N is only slowly mineralized
 and may persist in soil for prolonged periods.
      Bacterial denitrification is an  important factor regulating NO;!"
 levels in natural soil and may serve as a means of reducing the NO3-
 content of groundwater when  land  is  used  for the disposal of
 nitrogenous wastes.

 Allison, F. E. 1965. Evaluation of incoming and outgoing processes
     that affect soil nitrogen.  In  Soil nitrogen, ed. W. V. Bartholo-
     mew and F.  E. Clark, pp. 573-606.  Madison, Wis.: Am. Soc.
 Bremner, J.  M. 1965.  Organic nitrogen in soils.  In SoiZ nitrogen,
     ed.  W.  V. Bartholomew and F. E. Clark, pp. 93-149. Madison.
     Wis.: Am. Soc. Agron.
 	.  1967.  Nitrogenous compounds.  In Soil biochemistry,  ed.
     A. D. McLaren and G. H. Peterson, pp. 19-66.  New York: Mar-
     cel Dekker.
 Bremner, J. M., and Nelson, D. W. 1968.  Chemical decomposition
     of nitrite in soils.  Trans.  9th  Intern.  Congr. Soil Sci. Australia
 Broadbent, F. E., and Clark, F. 1965.  Dentrification. In Soil nitro-
     gen,  ed. W.  V. Bartholomew and  F.  E. Clark,  pp.  344-59.
     Madison, Wis.: Am. Soc. Agron.
 Broadbent, F. E., and Stevenson, F. J.  1966.   Organic matter inter-
     actions.  In Agricultural anhydrous ammonia:  technology and
     use, ed.  M. H. McVickar et al., pp.  169-87.  Madison,  Wis.:
     Am. Soc. Agron.
Flaig, W.  1966. The chemistry of humic substances. In The use  of
     isotopes in soil organic matter studies, pp. 103-27.  New York:
     Pergamon Press.

                       CHAPTER 8 / CHEMISTRY OF NITROGEN IN SOILS / 141

 Harmsen, G. W., and Kolenbrander, G. J.  1965. Soil inorganic nitro-
     gen. In Soil nitrogen, ed. W. V. Bartholomew and F. E. Clark,
     pp. 43-92. Madison, Wis.: Am. Soc. Agron.
 Harmsen, G. W., and van Schreven, D. A.  1955. Mineralization of
     organic nitrogen in soil.  Advan. Agron. 10:299-398.
 Hauck, R. D., and Stephenson, H. F.  1965.  Nitrification of nitrogen
     fertilizers.  Effect of nitrogen source, size and pH of the granule,
     and concentration. Agr. Food Chem. 13:486-92.
 Hurst,  H.  M.,  and Wagner, G. H.  1969.  Decomposition  of  14C-
     labeled  cell wall and cytoplasmic fractions from hyaline  and
     melanic fungi. Soil Sci.  Soc. Am. Proc. 33:707-11.  '
 Hutchinson, G.  L.,  and Viets, F.  G., Jr. 1969.  Nitrogen enrichment
     of surface water by  absorption of ammonia volatilized from
     cattle feedlots. Science 166:514-15.
 Kononova, M.  M.  1966.  Soil organic matter, 2nd  ed.  New York:
     Pergamon Press.
 Meek, D. B., Grass, L. B., and MacKenzie, A. J. 1969.  Applied nitro-
     gen loss in relation to oxygen status of soils.  Soil Sci. Soc. Am.
     Proc. 33:575-78.
 Mortland, M. M., and Wolcott, A. R.  1965.  Sorption of inorganic
     nitrogen compounds  by soil minerals.  In Soil nitrogen,  ed.
     W. V.  Bartholomew and F. E.  Clark, pp.  150-97.  Madison,
     Wis.: Am.  Soc. Agron.
 Nommik, H. 1965.  Ammonium  fixation and other reactions involv-
     ing a nonenzymatic immobilization  of mineral N  in soil.   In
     Soil nitrogen,  ed.  W. V. Bartholomew  and F.  E.  Clark,  pp.
     198-258.  Madison, Wis.:  Am. Soc. Agron.
 Scheffer, F., and Ulrich, B.  1960. Humus und Humusdiingung. Bd.
     1.  Stuttgart, Germany: Ferdinand Enke.
 Smith,  G. E.  1968.  Contribution of fertilizers  to water pollution.
     In Water pollution as related to agriculture, pp.  13—28.  Paper
     presented  at joint seminar, Univ. of  Mo., Columbia, and Mo.
     Water Pollution Board, Columbia.
 Stevenson, F. J.  1965.  Origin and distribution of nitrogen in  soil.
     In Soil nitrogen, ed. W. V. Bartholomew and F. E. Clark,  pp.
     1-42.  Madison, Wis.: Am. Soc. Agron.
 Stevenson, F. J., and Dhariwal, A. P.  S.  1959. Distribution of fixed
     ammonium  in soils. Soil Sci.  Soc.  Am.  Proc. 23:121—25.
 Stevenson, F. J., Kidder, G.,  and Tilo. S. N.  1967.  Extraction  of
     organic  nitrogen  and  ammonium from  soil with hydrofluoric
     acid. Soil Sci. Soc. Am. Proc. 31: 71-76.
 Stewart, B. A., Viets, F. G., Jr., Hutchinson, G. L.,  and Kemper, W. D.
     1967.  Nitrate  and other water pollutants under fields and feed-
     lots. Environmental Sci. Tech. 1:736-39.
 Swaby, R. J., and  Ladd, J.  N.  1962.  Chemical nature, microbial
     resistance, and origin of soil humus.  Trans. Intern. Congr. Soil
     Sci. (New Zealand),  Com. IV and V,  pp.  197-202.
Wagner, G. H., and Mutatkar, V. F.  1968. Amino components of soil
     organic matter formed during humification of 14C glucose.  Soil
     Sci. Soc. Am. Proc.  32:683-86.
Wetselaar, R.  1962. Nitrate distribution in tropical soils.  III.  Down-
     ward movement and accumulation of nitrate in the  subsoil.
     Plant Soil  14:  19-31.
Wullstein, L. H.  1967.  Soil nitrogen volatilization.  Agr. Sci. Rev.
     2nd Quart., pp. 8-13.


        HE rapid increase in fertilizer usage has been due largely to
low  fertilizer costs  and the necessity for higher economic yields.
Reliance  on legumes and  the use  of animal manures,  both for
nitrogen and erosion control, have given way to chemical fertilizers
in many areas. This higher fertilizer usage has vastly increased crop
residues which, in themselves,  tend  to protect the soil surface and
improve soil structure for moderating erosion. Crop varieties, with
high yield potentials, have also played a major role in increased crop
production.  In order for these new varieties to attain their maximum
yield  potentials,  increased fertilizer  rates have been necessary.  In
addition to  farm uses,  fertilizers are being  used more on parks,
playgrounds, golf courses, home lawns, roadbanks, forest recreation
areas, and even in forest lands.
     The rapid expansion in fertilizer use has raised many questions
concerning  nutrient  pollution  of our surface and groundwaters.
Since the population of the United  States is rapidly increasing, it
probably will be essential that our land acres produce food  and fiber
at capacity levels in the future. This will necessitate the continued
rise of high  rates of fertilizer. However, management practices  must
be followed such  that the high yields attained are also consistent with
a clean  and safe  environment.

     In the east north-central states of Wisconsin, Michigan, Illinois,
Indiana,  and Ohio, 8.1 million tons of fertilizers were used in 1968,

     W. P. MARTIN is Professor and Head, Department of Soil Science,
     University of Minnesota. W. E. FENSTER is Assistant Professor and
     Extension Specialist in Soils, Department of Soil  Science, University
     of Minnesota.  L. D. HANSON is Associate  Professor and Extension
     Specialist in Soils, Department of Soil Science, University of Minne-
     Miscellaneous Publication Paper No. 1360 of the University of Min-
     nesota Agricultural Experiment Station, St.  Paul.
     See Hargett (1969) for the statistics used in this section.

                               CHAPTER 9 / FERTILIZER MANAGEMENT / 143

or an  average  of about  135 pounds of plant nutrients per acre on
some 56 million harvested acres, 32%  of which was applied in  the
fall. This is approximately four times the usage in 1945. Nitrogen
has shown the most spectacular increase, 46,000 tons in 1945 to 1.3
million tons in 1968, or almost 30 times as much.
     In the west north-central states of North and South Dakota, Min-
nesota, Iowa, Nebraska,  Kansas, and Missouri, 7.8 million tons were
used in 1968, or an average of  about 65 pounds of nutrients per acre
on some 117 million harvested crop acres, and some 34% wras used
in the fall.  This is approximately  17 times the  usage in 1945 and
again nitrogen  has shown the most spectacular gains, increasing from
less  than 6,000 pounds  in 1945 to over 2  million pounds in 1968,
over 300 times as much.
     The north-central states are among the high-use  states, and in
the western and central parts the rapid expansion of irrigation enter-
prises  is accelerating  the use of fertilizer nutrients to maximize pro-
duction. Projection estimates  suggest further expansion, perhaps
even a doubling in fertilizer use in the next 15 years, so that the Mid-
west will account for some 40% of the total used in the United States
(Beaton and Tisdale, 1969).
     Although  the aforementioned  figures are spectacular in terms
of increasing usages of plant nutrients in fertilizers, their utilization
by crops must  be balanced against the pollution aspects  of the soil-
water  system.  Many of  these factors  have been  covered in  great
detail in the preceding chapters, however, it will be necessary to pro-
vide modest documentation in  order to relate principles of soil and
crop management for production to the problem of minimizing  po-
tential pollution of water supplies from use of  fertilizers (Soileau,
1969). The discussion will be  confined to nitrogen and phosphorus,
the two  nutrient elements of  principal concern in water pollution
and eutrophication.

     It should be pointed out that our cropping programs have, in
general,  been exploitive of plant nutrients and that  we are still re-
moving more nutrients than  are being replaced by way of fertilizers,
or from other sources. White (1965) evaluated the situation and es-
timated that major crops in the United States on our 294 million
cropped acres were removing about  8.8 million tons of nitrogen (in-
cluding nitrogen fixed in leguminous plants of approximately 3 mil-
lion  tons) and 2.8  million tons of phosphate. Only in the case of
phosphorus  are the additions  equivalent  to  the withdrawals, and
when it is considered that crop use efficiencies are substantially less
than 50% of that applied, we are still "mining" rather than "enrich-
ing" our soils with plant nutrients.
     Stanford (Wadleigh, 1968) has estimated that in the  past 100
years there has been a loss of organic matter in the top 40 inches of
the cropped agricultural soils of the United States of  some 35 billion
tons, or a loss of 1,750 million tons of organic nitrogen. Nitrogen fer-


 tilizer application, though appearing to be large and now approaching
 annual crop removal levels, is small in terms of "historical losses."
      In Minnesota, for example, on approximately 18 million acres of
 cropland, nitrogen  withdrawals average close to a million tons an-
 nually and phosphorus some  200,000 tons.  Less  than a quarter of
 this amount is being added by way of fertilizers, so even taking into
 account nutrients added through manures and legumes, two to three
 times as much chemical fertilizer could be justified for crop produc-
 tion at current levels.
     It  is evident, however, that we  may  reach  application levels
 where the additions of plant nutrients surpass crop removals and in
 Iccal situations  now, very high  application rates  of nitrogen  par-
 ticularly are sometimes noted (Beaton and Tisdale, 1969).  It is pos-
 sible to enrich local water  supplies,  especially  where  soils are  not
 adequately protected from erosion.  It is necessary,  therefore,  that
 attention be given now to those management factors that can assure
 the crop production needed and at  the same time  minimize  the
 potential for nutrient pollution.

     The conservation movement of the past 30 years has stimulated
 and supported a  major research effort which has documented the
 seriousness of erosion and sedimentation both from the standpoint of
 land destruction and water degradation. It has been estimated  that
 some 4 billion tons of sediment are washed into waterways and reser-
 voirs annually; this is equivalent to about 4 million acres of good top
 soil 6 inches deep (Stallings, 1957; Smith and Wischmeier, 1962;
 Wadleigh, 1968).  Marked abatement of this erosion and sediment de-
 livery  can be accomplished by erosion control structures, crop rota-
 tions, use of minimum tillage, and utilization of crop residues both by
 incorporation to improve structure and by mulching to protect soil
     Smith and Wischmeier (1962)  developed  a "universal rainfall
 erosion equation" by integrating data  from some 35 field research sta-
 tions. This equation aids in management decisions designed to keep
 soil losses in  the field below established "tolerance" limits  of 3 to 4
 tons per acre annually. The  equation identifies key determinants in
 soil loss and sediment delivery and defines them in terms of average
 annual erosion-producing rainfall,  soil erodibility,  topography, crop-
 ping and cultural practices, and erosion  control activities (Ballantyne
 et al.,  1967).  In the future these activities will likely take  on the
 increasingly important role of controlling lake-destroying sediments.
     A further consideration  of interest is the nutrient aspects of the
 land sediments reaching water. Most researchers have felt that  just
 as fertile soils produce more land plants  via higher  equilibrium levels
 of available nutrients, so do fertile  sediments provide more nutrients
for aauatic plants.
     The physical  removal of  nutrient elements by erosion is  non-
 selective in the sense that the elements may be removed in any chemi-
 cal form.  The process, however, tends to be selective in that the

                              CHAPTER 9 / FERTILIZER MANAGEMENT / 145

organic matter  and finer  particles  of soil are more  vulnerable to
erosion than are the coarser soil fractions (Barrows and Kilmer, 1963),
Organic matter is among the first constituents to be removed because
of its low density and high concentration in surface soils. Hays et al.
(1948) reported  951  pounds per acre of organic matter lost annually
from moderately eroded Fayette  silt loam and 668 pounds  from a
severely eroded  phase.  Significant  quantities of nitrogen and phos-
phorus  may be removed in the organic phase.  Massey and Jackson
(1952) calculated regression equations for the  enrichment ratios of
organic matter and plant nutrients from Almena, Fayette, and Miami
soils in Wisconsin, using runoff plot and  small watershed data, and
concluded that they were removed selectively in the following order:
organic matter,  organic and ammoniacal nitrogen, and finally "avail-
able" phosphorus.
     Losses  reported for soluble nitrogen  salts  and unreacted phos-
phatic fertilizer compounds in runoff waters  are exceedingly low
and appear to be  of little  significance (Barrows  and Kilmer, 1963;
Biggar and  Corey, 1968; Wadleigh, 1968). However,  existing data
are insufficient to evaluate the influence  of  such  factors as source,
rate, placement, and time  of application  of fertilizer relative to the
occurrence of runoff. Hauser (1968) recently sampled closed playas
on  the Texas high plains entrapping runoff waters from heavily fer-
tilized adjacent  fields  and found  them  virtually  free  of  nitrates.
Samples taken on five different sampling dates contained  less than
0.5 ppm of nitrate nitrogen, on  an average, and  the  same  values
were recorded for playas whose watersheds were 95% native grasses.
Rogers  (1942) applied triple  superphosphate  at the  rate  of 200
pounds per acre  to Dunmore silt loam in permanent pasture, followed
immediately by  a  series of 1-inch rains from  a rainfall simulator.
The first  rain removed  9.1%  of the  applied  phosphorus  and the
second 4.3% . As much as 22%  of the phosphorus applied to a dry
bare soil was removed  when  rain \vas applied immediately after fer-
tilization.  Phosphorus solution and immobilization by soil fixation
could not occur with sufficient rapidity under  these extreme condi-
tions to prevent some loss in the runoff waters.
     It is evident that erosion can and does in many instances cause
significant losses of soil and organic matter with concomitant removal
of nitrogen  and phosphorus.  Previously mentioned erosion control
measures  can reduce  these losses  by  75% and more (Wadleigh,

     Recently  there  has  been increased  appreciation of  the sig-
nificance of phosphorus in the process of lake eutrophication (Megard,
1969). The nitrogen-fixing capability of blue-green algae often dimin-
ishes the significance of nitrogen as  a nutrient and increases that of
phosphorus.  The phosphorus regimes  in soil versus water vary
markedly in that only a small  portion of  the soil phosphorus is  in an
available form, whereas that in water is almost totally available. The
amount of  phosphorus  in soils is, therefore, much larger than that in


 our natural waters.  Soils may vary from  100 to 4,000 pounds  of
 total phosphorus (1,000 pounds per acre  average) in the plow layer,
 only 5 to 10% of which contributes to the "labile pool" of potentially
 available phosphorus (Bailey, 1968; Black,  1968).  An example  of a
 quantity  of phosphorus contained in  a Minnesota lake to a depth
 where light was sufficient for photosynthesis is 2.8 pounds per acre.
 This amount is sufficient for profuse algae growth (Megard,  1969).
     Lysimeter  and other types of  experiments  have demonstrated
 that phosphorus does not significantly leach downward as a result of
 water percolation; drainage waters thus contain small concentrations.
 The highest amount reported was that in California on well-drained
 soils receiving large amounts of fertilizer and water by w^ay of irriga-
 tion, where Johnston et al.  (1965) recorded a mean concentration of
 0.08  ppm of phosphorus in  the irrigation drainage. Water moving
 into natural waters from underground flow will contain phosphorus
 at levels  consistent with those found generally  in uncontaminated
 waters or from 0.01 to  0.03 ppm (Maderak,  1963).  Agricultural land
 drainage  is  usually in this  same range (MacGregor and  Hanson,
     The magnitude of  runoff and sediment sources of phosphorus is
 under extensive investigation  at  the present time.  Past work has
 tended to emphasize erosion.
     Many experiments on soil loss from erosion have  been carried
 on by the Missouri Agricultural Experiment Station, dating from 1917
 (Duley  and Miller, 1923). In  an  experiment on Shelby loam, with
 plots 90 feet long and having a 3.86%  slope, loss  of phosphorus by
 erosion was 18 pounds per year with continuous corn and 6.2 pounds
 with a good rotation of corn, wheat, and clover.  Under continuous
 bluegrass, only 0.1  pound of phosphorus was lost by erosion, demon-
 strating the effectiveness  of plant  protection against soil loss (Miller
 et al.,  1932).
     Bedell et al. (1946) demonstrated the loss of organic phosphorus
 through erosion. Where corn was grown under prevailing manage-
 ment practices,  over 4.5 tons of solids were  removed per acre, carry-
 ing approximately  20 pounds  of  phosphorus.  Nearly  60%  of the
 phosphorus lost  was in the organic  form. Eroded soil from natural
 runoff-erosion plots on a Barnes loam soil, 7% slope, at Morris, Min-
 nesota, contained 500 to 2,000 ppm total P (Timmons et al., ]968).
These were agricultural soils which had been adequately fertilized,
 and cropping patterns varied from  clean-cultivated fallow,  through
 continuous corn to corn-oats-hay in rotation.
    Recent studies have indicated that lake sediments  are not con-
 tributing to lake water pollution by supplying phosphate and indeed
will be  able  to extract phosphate from  the  waters with which they
come in contact. The phosphate potential for lake bottom sediments
from  several western  Minnesota  lakes  with varying  degrees of
eutrophication were determined and the "index" factors were found
 to vary from 8.22 to only 8.59, indicating a high degree of phosphate
adsorption capacity (White  and Beckett, 1964;  Holt et  al.,  1969;
Latterell et al., 1969).
    The runoff  source of phosphorus appears to  be more important
than was previously suspected.  Of particular interest is Holt's (1969)

                              CHAPTER 9 / FERTILIZER MANAGEMENT / 147

observation that spring snow melt waters carry much higher amounts
of phosphorus  than runoff waters during other times of the year as
measured on natural-rainfall-runoff plots.  This is presumably coming
from plant residues that have been frozen over winter, allowing  the
nutrients to be washed out in  the  spring. Also,  soil contact is pre-
vented by a mat  of  organic material  on the soil surface and  the
frozen soil itself. Amounts, though small, were five to six times higher
in snow  melt runoff than in water percolating through the  soil. This
observation was supported  by  Hanson and Fenster (1969)  in data
comparing  the phosphorus concentration of tile  line waters (soil
percolation)  and adjacent drainage ditch waters  (percolation plus
snow melt) in the spring of 1969 for sampling sites on 20 farms in
southern Minnesota's corn- and soybean-growing  area. Soils at  the
sampling sites were Webster and Glencoe clay loams.  Tile line out-
let waters averaged 0.03 ppm  phosphorus vs. 0.16 ppm for the  ad-
jacent ditches, or five  times as much.
     It will be difficult to intercept  snow melt runoff coming off  the
large areas of natural grasslands. Diversion of surface runoff waters
to seepage areas may in some instances be feasible, but often at con-
siderable expense.  Water  which has  percolated through  a  mineral
soil  is essentially  stripped of its soluble  and particulate phosphate
and  will not be a  significant source of phosphorus in  natural water
supplies  whether or not it comes from fertilized or unfertilized soils.
Land leveling, retention terraces, and other erosion control measures
can  be designed to help on cultivated  soils. Spreading of fertilizer
and  manures on frozen soils on  rolling land adjacent to water supplies
should be avoided (Corey et al.,  1967).

     In many ways the questions raised concerning fertilizer manage-
ment for control of pollution are premature in that water quality
benchmarks for lake eutrophication  or for public and animal health
have not been jointly established or agreed upon. However, regard-
less of interpretation with respect to  these two facets  of the problem,
it  is important that nitrogen  movement and losses  be reviewed as
related to land management activities. Most modern crop production
practices will affect the soil nitrogen regime in some degree and it is
difficult to generalize about them because  of the complex nature of
the soil-climate-plant system.  Interactions among soil, climate, irriga-
tion, drainage, tillage, fertilizer, and crop are exceedingly complicated
and becoming more so with rapidly changing patterns of fertilizer
use and tillage methods.
     Nitrogen in  the  Soil

     Tillage is a significant factor in the release of nitrogen, since it
markedly speeds up oxidation of organic matter with release of, and
subsequent nitrification of,  ammonia.  The high organic matter soils
of the north-central  states containing 0.2  to 0.4%  by  weight of


 nitrogen have  a potential for production of nitrate  which is sub-
 stantial, and yields of 200 to 400 pounds per acre annually without
 supplementary nitrogen were not uncommon during the early years of
 cultivated agriculture in the region (Black, 1968). To  a large extent,
 nitrogen fertilizers as currently used can be considered a replacement
 for lower "yields" of tillage-induced soil-organic-matter-released nitro-
 gen. From  this perspective, chemical nitrogen fertilizer use is not
 essentially different from the older agronomic practices of cultivation
 and the use of legumes and  animal manures to shift the nitrogen
 equilibrium  in  favor of the crop plant.
      Allison (1955,  1965) has extensively reviewed nitrogen balances
 in soils and  concludes that as long as it remains in the organic form
 it is comparatively safe from loss except through erosion.  Normally,
 however, as  noted above, soil organic nitrogen is slowly converted to
 ammonia by heterotrophic soil microorganisms and then into nitrites
 and nitrates by the nitrifying bacteria, and in these forms it is subject
 to the same  losses as nitrogen from fertilizer  sources.  In the absence
 of a crop or leaching, as much as 5 to 10%  of the nitrogen may ac-
 cumulate as nitrate during  a 6-month period in cultivated soils. A
 crop such as corn removes 2 to 3% of the nitrogen in  the plow layer
 during one growing season.  Small grain crops remove half as much
 as corn.  Nitrogen returned to the soil in crop residues may  contribute
 as much as 20%  of the total nitrogen assimilated by the crop.
      Many investigators, using lysimeters, have reported on leaching
 losses of nitrogen in the form of nitrates  (Bizzel, 1944;  Allison,  1955,
 1965; Allison  et al.,  1959;  Black, 1968; Webber,  1969b).  Allison
 (1955), for example, summarized the results of 157 lysimeter experi-
 ments conducted at several locations in the United States.  These in-
 cluded  51 lysimeters kept fallow and 106 that were cropped to  non-
 legumes. The unaccounted-for nitrogen averaged 15%   of that added
 or  which became available in the soil.  He noted that this nitrogen
 loss could not be assigned to denitrification, which under normal soil
 conditions is quite small. Nitrogen recovered in the crop from added
 fertilizer nitrogen, or  from that  which became available in the soil,
 was usually  less  than 50% .  Significantly, however, Pearson et al.
 (1961) showed that equivalent nitrogen recoveries for three successive
 crops ranged from 70  to 77% in the humid southeast if nitrogen was
 applied at a  time when leaching was at a minimum and when crops
 were present to effect  assimilation. Two hundred pounds of nitrogen
 per acre were applied  to corn in the spring with two additional crops
 being grown during the following 16 months.  It was concluded that
 in general, leaching losses of nitrogen are small if an actively growing
 crop is present at all  times, and if the rate of nitrogen  additions
 clearly approximates the needs of the crop. If much nitrogen is added
 or released as nitrate in the  late fall, the losses are likely to be large
 unless a cover crop or permanent sod is present.
     Russell  (1961) documents  that  nitrates are  always  lo\ver in
 cropped land than under fallow, not only because the crop is ex-
 tracting nitrate  for growth but because the crop depresses the rate of
 nitrification in soil.  Nitrogen present in a cropped soil  as nitrate and
in the crop was less  than the nitrate in an adjacent fallow  soil.
 Nitrates accumulated  during the spring  and summer  in the fallow

                               CHAPTER 9 / FERTILIZER MANAGEMENT / 149

soil, but not in the cropped soil.  Fallow soil also appeared to lose
substantial amounts of nitrates by early winter, presumably because
of leaching into the subsoil.
     Smith (1968) in Missouri has also found that nitrates in  small
amounts may reach groundvvater supplies.  Soil  samples in foot in-
crements were taken to a depth of 10 feet in Putnam silt loam which
had been in  continuous corn for 20 years and had  received 120
pounds of nitrogen as  ammonium nitrate per acre annually.  Nine
pounds per acre per year more nitrogen was found in the surface  10
feet of sell than where no nitrogen was applied.  Other  nitrate move-
ment studies  have been underway for 7 years on soils with widely
different characteristics.  Rainfall has ranged from 30  to  50 inches
per year and corn yields from 60 to more than 150 bushels per acre.
Nitrates in small amounts have accumulated  progressively in the 8-
foot profile samples at all levels of application above 100 pounds per
acre except in the sandy soils where leaching has presumably moved
nitrates below the depth of sampling.
     Higher nitrogen losses have been measured by Johnston (1965)
in connection with a study of tile  drainage and  wastewater manage-
ment in the  San Joaquin Valley of California.   He noted that 9 to
70%  of  the nitrogen applied just  prior to the 1962 irrigation season
or with the irrigation water was lost either in the drainage  effluent or
in the tailwater. It was noted that the presence of a continuous  water
table at  or above  the tile systems was necessary to obtain the data
presented.  Nitrogen rates varied from  84 to 260 pounds per acre and
crops were cotton and rice.
     Gardner  (1965) notes  that the downward  movement of  water
through  the "macropore systems" of medium-textured soils is rather
rapid. The larger  the total pore  volume of this system, the  more
readily the water will move. The presence of  a crop, however,  tends
to reduce this downward movement  because of evapotranspiration.
The crop, therefore, greatly minimizes  leaching losses of nitrogen
both directly, by assimilation, and indirectly, by reducing the amount
of leachate.
     A suggested way of evaluating probable leaching losses of nitrc-
gen and  how to minimize them is  to make use of precipitation-evapo-
transpiration curves as noted by Allison (1965).  Using this approach,
for an example, the results for central Minnesota show that there is
little  cr no leaching or movement of  water through  the soil during
the summer  months, and during  the winter months  the soils are
fro/en.  Fall and  spring are the months when drainage  occurs in the
Midwest (Blake et al.,  1960).  In drier regions of the  Great Plains
states, the soil is rarely filled to field capacity beyond the  root zone.
except under  irrigation, and  hence there will be little leaching of
nitrogen.  These data emphasize the importance of a crop for mini-
mizing leaching losses and the importance of avoiding the  accumula-
tion of nitrates in soils in late fall.
     It should perhaps be noted that nitrates reaching the drainage
ditch, lake, or reservoir are  either quickly used by plants or denitrify
in  the anaerobic  high-enerey  environment of  decomposing  plant
materials (Allison, 1965).  Keeney et al.  (1969) were not able to
detect nitrates in Wisconsin  lake  sediments, nor are  nitrates com-


 monly found in more than trace amounts in surface waters (Maderak,
     Sources of Nitrate  in Water Supplies

     Natural  sources of nitrogen,  as  well as those from fertilizers,
 leguminous crops, and  animal waste  disposal operations, must be
 evaluated for perspective in  considering ways in which soil manage-
 ment may minimize the nitrate pollution of water supplies.  These
 quantities are difficult to estimate, so considerable watershed monitor-
 ing will be needed in the future. Unless additions of nitrogen from all
 sources which go into a watershed are balanced by withdrawals such
 as by harvest and removal of crops or  denitrification, nitrate-nitrogen
 will accumulate in surface  water  or  groundwater (Stewart et al.,
     Nitrate contamination of water supplies from barnyards or con-
 centrated livestock feeding operations  has been well documented and
 corrective measures are  being instituted in many states to moderate
 the problem via collecting basins, oxidation trenches or lagoons, and
 land spreading of the animal manures.  Since the animal manures are
 a source of plant nutrients released during decomposition, they are in
 the same category  as fertilizers  when applied to cropland in evaluat-
 ing management procedures which will minimize pollution.
     Smith (1965, 1967) researched sources of nitrogen in some 6,000
 rural water supplies in Missouri and  concluded that animal wastes
 and septic tank drainage coming from poorly constructed  shallow
 wells were the main sources of water contamination. He suggested
 that fertilizer nitrogen was not at this time significant overall, though
 in some instances application rates go beyond efficient crop utilization
     In Minnesota, also, nitrate contamination of rural wells has
 been noted for many years, long before nitrogen fertilizers were used
 to any extent. Shallow wells in  glacial drift and in the Shakopee and
 Oneota dolomites, with recharge directly from the drift, are higher in
 dissolved  solids and often contaminated  with nitrates above Public
 Health  Service standards.  This occurs most notably in communities
 without municipal sewage disposal systems and where  large num-
 bers of livestock are concentrated.  However, recent summary reports
 by  the Minnesota State Department  of Conservation (Maderak, 1965)
 on  the  chemical quality  of groundwaters show that wells from deep
 aquifers such  as the Jordan  and St. Peter sandstones, with recharge
 from  the northern  Minnesota lake and forest areas, are  very low in
 nitrates and dissolved solids. In  general, change in  the quality of
 wrater for  the major aquifers  from 1899 to 1963 has been minor.
     Schmidt (1956) studied  the problem  of anoxemia in  very young
infants as related  to nitrate  contamination of rural wells varying in
depth from 15 to  50 feet in southern Minnesota's prairie soil  area.
Soils obviously containing high  levels  of organic nitrogen from live-
stock  had the  highest nitrifying capacities, and water supplies with
concentrations of nitrate  of  75  to 130 ppm  were located near such
soils.  Normal field  soils were associated with subsoil drainage waters

                               CHAPTER 9 / FERTILIZER MANAGEMENT / 151

 of up to 18 ppm and well waters of up to 35 ppm nitrate. The high
 organic matter  soils of southwestern Minnesota present ideal  condi-
 tions for nitrification and release of nitrogen from the organic nitro-
 gen complex  with  cultivation so that supplementary nitrogen from
 animal manures or other sources can result in high nitrate produc-
 tion which may move into shallow wells which are improperly located
 or constructed.  This illustrates that fertilizer rates  applied must be
 evaluated in terms of the "background" level of fixed nitrogen  in the
 soil water.
     Stout and  Burau (1967) studied nitrate accumulations in the
 groundwaters of a closed 10-square-mile basin near San Luis Obispo,
 California.  This area, containing 2,700  intensively cropped acres, is
 urbanized with  13,500 people and the domestic arid irrigation waters
 are supplied exclusively from wells with a rapid recharge.  These well
 waters ranged from 5 to 130 ppm in nitrate. The nitrogen pool was
 substantial  and mostly associated with the native soil organic matter
 complex supplemented with sewage  waste from area  homes and to a
 lesser  extent  from lawn and farm  fertilization. Cropland manage-
 ment recommendations were developed by Stout and Burau to include
 nitrogen fertilizer rates consistent with the needs of the crop, mostly
 fruits and vegetables, and to include the amounts of  nitrate-nitrogen
 in the well waters which  were used for irrigation. It was recommend-
 ed that domestic waters be taken from prehistoric deep waters  which
 are unaffected by tillage and hydraulic sewage disposal systems.
     One of the more recent and well-documented studies on sources
 of pollution of underground supplies was made in the middle  South
 Platte River valley  in Colorado (Stewart et al., 1968). This valley is
 intensively farmed  and irrigated, has some 600,000 cattle  in feedlots,
 and is surrounded by many cultivated  dryland  fields. Twenty-foot
 cores from the soil surface to water table or bedrock from 10 to 65
 feet deep were taken from 129 sites of differing land use and analyzed
 for nitrates in transit. Average total nitrate nitrogen in  pounds per
 acre for land use types was as follows:  alfalfa, 79; native grassland,
 90; cultivated dryland, 261; irrigated  fields not in alfalfa, 506; and
 cattle corrals, 1,436 pounds.  In general, there was extreme variation
 within classes of land use. Calculations based on core averages and
 rate of water movement through the  profile under irrigation indicated
 that 25 to 30 pounds of nitrate nitrogen per acre  were being lost
 annually to the water table.  Though the losses were small  from ir-
 rigated fields compared to those from feedlots, they contributed much
 more total nitrate because acreages were much more extensive.
     Management to Minimize Nitrogen  Losses

     In summary, it is evident that leaching  of nitrates  below the
rooting zone of plants can and does occur in  soils, depending upon
nitrogen supplies present,  and that it may be larger on sandy soils
under irrigation than on heavier textured soils during summer when
evapotranspiration is greater than  precipitation.  Under fallow or in
late  fall and early  spring  when  soils  are not frozen, movement of
nitrates  downward  within  the soil profile  occurs and some  may


 eventually  reach underground  water supplies.  Erosion losses of
 nitrogen are mostly associated with the selective removal of organic
 nitrogen compounds in the erosion debris.  Thus, to minimize losses
 and moderate potential pollution, soil management should include
 erosion control practices, soil nitrogen should be kept  to a minimum
 during the colder months of the year or in the absence of a crop, and
 fertilizer nitrogen should be added in amounts which allow for, but
 do not greatly exceed, the amounts needed for efficient crop produc-
 tion. It may be necessary to again  emphasize the value of split ap-
 plications  of  nitrogen fertilizer  for efficiency of utilization via the
 irrigation waters, with starter and nitrogen side-dressing of corn and
 summer top-dressing of meadow and  turf—all good crop production
 management  recommendations.   Late season  or  fall  application of
 nitrogen should be in the ammoniacal form where soils are  subject
 to leaching. Temperatures should be low enough  that nitrification is
     Alexander (1965) noted that the optimum temperature for nitri-
 fication falls between 30°  and 35° C. There is no fixed minimum tem-
 perature above freezing,  but rates  are  low.  However, nitrate will
 continue to be formed throughout the autumn in small amounts and
 may be lost by leaching in those situations where there is movement
 of water through the profile.  This nitrate may  be utilized  by soil
 microorganisms if carbonaceous crop residues are incorporated in
 the fall.
     Chemical inhibitors  to delay oxidation of  ammonia  to nitrates
 and nitrites have been suggested (Black,  1968).   Alexander (1965)
 listed many inhibitors and summarized the literature on their use.
 Turner and Goring (1966) examined a number of researchers on the
 use of 2-chloro-6-(trichloro-methyl) pyridine, one of the more effective
 inhibitors,  and concluded that yield and nitrogen content of  several
 crops could be increased by the use of this inhibitor. In general, the
 inhibitors appeared to be more effective at temperatures below 21° C
 and much less effective  at  temperatures up to  32° C.  Studies by
 Janssen and Wiese (1969) in Nebraska and by Huber et al. (1969)
 in Idaho support these conclusions. Further investigation is warrant-
 ed to improve use reliability and for reduction in price.
     Under irrigated soil  conditions,  or farming situations  where
 there is control over the water table, it may be possible to dissipate
 some nitrate entering tile lines with controlled demtrification. Meek
 et al.  (1969),  using  simulated tile  lines  in  soil columns,  reduced
 nitrates in the tile effluent to an average value of 0.5 ppm by sub-
 merging the tile lines, thus creating an anaerobic environment.
     The role of deep-rooted crops, like alfalfa, in a rotation and of
 cover crops in  the fall to remove nitrogen  and  enhance the organic
 nitrogen reserve should  not be  minimized for  selected locations.
 Stewart et al.  (1968) showed that little nitrate was  present  under
 alfalfa fields and grasslands to depths of 20 feet.  Where the water
 table is within this depth, some nitrate may  even be  removed from
 the water table.
     Sod crops and  crop  residues left on soil surfaces  during non-
cropping seasons can also reduce erosion as will  minimum  tillage.
The  adoption  of  minimum tillage practices (Cook, 1962) would ap-

                               CHAPTER 9 / FERTILIZER MANAGEMENT / 153

 pear to be particularly warranted as a management tool for protec-
 tion of the organic nitrogen pool against rapid oxidation and yield of
 nitrates.  This practice also protects against destruction of soil  ag-
 gregates, preserves structure,  and decreases runoff (Burwell  et  al.,
 1968).  Infiltration is increased,  which provides more water for use
 by plants and a more extensive plant cover.
     As noted in an earlier section, nitrogen is a key element in crop
 production because of its transitory nature in soils, and it is becoming
 more economical to add too much fertilizer nitrogen rather than to
 risk not applying enough.  It is evident now that in  selected locations,
 movement of surplus nitrates into water supplies can be serious, not
 only as a potential  pollutant but also as a loss to the efficiency of pro-
 duction.  University and commercial soil and plant tissue testing lab-
 oratories  and procedures are generally available for making fertilizer
 recommendations,  and these are geared for efficiency of production at
 yield potential levels estimated  to be feasible for a given soil  and
 climatic area.  Nitrogen recommendations, particularly, are based on
 the nitrogen requirement of the  crop for maximum efficient produc-
 tion (in the case of crops like sugar beets, potatoes, or malting  barley
 where surplus nitrogen reduces quality, it should be the "minimum"
 requirement for maximum  production of  a product of acceptable
 quality, as suggested by Stanford et al., 1965), efficiency of utilization
 of nitrogen fertilizers used, and  the nitrogen-supplying capability of
 the soil via release of nitrogen from the  organic nitrogen pool (Fenster
 et al., 1969).  As noted earlier, when irrigation waters from surface
 wells containing nitrates  are  used,  cropland  management  recom-
 mendations can be developed to include the amounts of nitrogen
 which will be supplied  with  the  irrigation waters (Stout and Burau.
 1967).  Realistic recommendations can help avoid overapplication.
 The environmental quality factor will have to be  brought into  the
 formulation of responsible recommendations.
     A number of researchers  are attempting  to determine what
 constitutes an  acceptable application rate for nitrogen that will both
 sustain production  and minimize pollution.
     Webber (1967) in  Ontario,  Canada, has postulated an applica-
 tion rate  for farm  manures  which would not release on decomposi-
 tion over 300 pounds per acre of ammonia-nitrogen which could be
 oxidized to nitrate.  This amount  could presumably be utilized by corn
 or by hay-pasture crops and removed with  the crop or be dissipated
 otherwise, such as  by denitrification or tied up by microorganisms in
 the decomposition of crop residue. At this level, it was suggested  that
 there would be little nitrate in surplus which would move into under-
 ground water supplies.  It was further suggested that for small grains,
 or sandy soils under irrigation, a lower figure would have to be used.
 The  figure for "safe" application levels  of nitrogen  for different crop
 management systems is currently being checked out, using 32-inch
 diameter, 42-inch deep  lysimeters on a Guelph loam.
     In Missouri Smith (1968) suggested an application rate no higher
 than is required for optimum  yields,  approximately 100 pounds of
nitrogen per acre, if nitrates are not to reach  groundwater supplies.
As noted earlier, nitrate movement studies have been underway for 7
years on  soils with widely different characteristics and where corn


 yields have varied from about 60 to more than 150 bushels per acre.
 Nitrates accumulated progressively in 8-fcot profiles at  all levels of
 application above  100 pounds per acre annually.
      Cooke (1969) suggests that use efficiency  of nitrogen fertilizers
 which  currently average less than 50%  must be increased  in  the
 future  not only for reduction in  crop production  costs but to avoid
 loss of nitrogen by leaching. Promising researches relate to the con-
 trol of ammonia oxidation and other reactions in soils as noted earlier
 and perhaps also the decomposition of urea, higher analysis and more
 readily available compounds of phosphorus reacted with  ammonia,
 pelleting of fertilizers to control solubility rates and with the seed for
 immediate utilization, and "agronomic control" by plant analysis with
 subsequent and immediate  application of fertilizer if needed by aerial
 topdressings or perhaps in  the irrigation waters. Different soils, cli-
 mates,  and cropping systems would have to be given individual re-
 search  attention.

     Nitrogen and phosphorus, as nutrient  elements, are important
 to both land and  aquatic plants, and normally reach water supplies
 via land runoff in the erosion debris which is selectively enriched in
 organic nutrient materials or  via  the leachate which may contain
 mobile nitrate ions.
     Fertilizer usages in the midcontinent area are rapidly increasing
 to maximize production and increase efficiency, and further increases
 are anticipated.  Current information suggests that phosphatic fer-
 tilizers incorporated in the soil are not contaminating natural waters,
 but nitrogen fertilizers may be contributory in  selected situations.
 For example, where application rates, together  with  soil supplies,
 have exceeded  crop needs  and/or excessive  leaching occurs induced
 by over-irrigation of sandy  soils, nitrates can be contributed to under-
 ground water supplies.
    Fertilizer phosphorus quickly  converts  to unavailable forms  in
 mineral soils and  the  evidence  indicates that one of the ways of re-
 ducing the level of soluble  phosphorus in water would be to effect soil
 contact such as by  filtration through the soil medium.   Some phos-
 phorus is removed from frozen plant materials with snow melt waters
 which  is difficult to  control except perhaps by diversion  terraces into
 seepage areas.
    Nitrogen fertilizer  application rates should  approximate crop
 needs,  which for a given soil type and climatic zone are based on pro-
 duction potential estimates for the crops to be grown.  One hundred
 pounds of nitrogen per acre can apparently be safely  applied  to
 cropped soils without  major contribution of nitrate  to the leachate,
 and up to 300 pounds  per acre in some instances,  though much more
research is needed in this area.
    Management  recommendations  refined  through the years  for
 maximizing production  are not incompatible with the  objective  of
reducing nutrient  contamination of  natural waters.  These involve
 an emphasis on erosion control  measures to include vegetative cover

                               CHAPTER 9 / FERTILIZER MANAGEMENT / 155

 which, in addition to a reduction in runoff and erosion, removes fer-
 tilizer nutrients with the harvest  and effects water transpiration to
 reduce leaching. Other factors include the use of cover crops where
 adapted  and incorporation of crop residues in the fall for protection
 of soil surfaces and utilization of plant nutrients, minimum tillage to
 improve  structure and reduce the  mineralization of organic nitrogen
 reserves, and an emphasis on increasing fertilizer use efficiencies by
 the crop, such as by split applications and  the  use  of ammoniacal
 forms of nitrogen in the fall.
     Further research is needed on nutrient balances and reactions
 in soils to maintain supplies at levels needed for crop production; to
 increase the efficiency of use as a percentage of that supplied, cur-
 rently less than 50% ; and to minimize loss  of nitrate to water sup-
 plies.  This  would include research on nitrification-inhibiting chemi-
 cals so as to retain nitrogen in the ammoniacal form, pelleting,  or
 other to reduce solubility  or application with  the  seed to increase up-
 take, and plant analysis monitoring of nutrients with needed applica-
 tions applied quickly, perhaps by  air or in the irrigation waters.
     Water  quality standards as established by the federal and  state
 water pollution control groups should be compatible with the need for
 maintaining adequate nutrients for efficient crop production  con-
 sistent with management programs designed to minimize  losses  to
 adjacent water supplies.

Alexander,  M.  1965.  Nitrification.  In  Soil nitrogen, ed.  W.  V.
     Bartholomew and F. E. Clark, pp. 307-43.  Madison, Wis.:  Am.
     Soc. Agron.
Allison, F.  E.  1955.  The enigma of soil nitrogen balance  sheets.
     Advan. Agron.  7:213-50.
	.   1965.  Evaluation of incoming  and outgoing processes that
     affect soil nitrogen.  In Soil nitrogen, ed.  W.  V.  Bartholomew
     and F. E. Clark, pp. 573-606.  Madison, Wis.:   Am. Soc.  Agron.
Allison, F. E., Roller, E. M., and  Adams, J. E.  1959.  Soil fertility
     studies in lysimeters containing lake land sand.  USDA Tech.
     Bull. 1199.
Bailey, G. W.  March 1968.  Role of soils  and sediment  in ivater pol-
     lution control. I. Reactions of nitrogenous  and phosphatic  com-
     pounds with soils and geologic strata.  Fed.  Water  Pollution
     Control Adm., Southeast Water Lab. Bull., U.S. Dept. Interior.
Ballantyne, C. R., Schaller,  F. W., and  Phillips, J. A. Dec. 1967.
     Erosion control  factors  and universal soil  loss equation.  Iowa
     State Univ. Coop. Ext. Serv. Bull, p. 410.
Barrows, H. L., and  Kilmer,  V. J.  1963.  Plant  nutrient losses  from
     soils by water erosion. Advan. Agron.  15:303—16.
Beaton, J. D., and Tisdale,  S. L. 1969.  Potential plant  nutrient con-
     sumption in North America. Sulphur Inst.  Tech. Bull.  16.
Bedell, G. D.,  Kohnke, H., and Hickok, R. B.  1946. The effects of
     two farming sytems on erosion from cropland.  Soil  Sci. Soc.
     Am. Proc. 11:522-26.
Biggar, J. W., and Corey, R. B.  1968.  Nitrate and phosphate in lakes
     and streams. Unpublished mimeo.,  Univ.  of Wis., Madison.


 Bizzel, J. A.  1944. Lysimeter experiments. VI.   The effects of crop-
     ping and fertilization on  the losses of nitrogen from the soil.
     Cornell Agr. Exp. Sta. Memo. 256, pp.  1-14.
 Black, C. A.  1968.  Soil-plant relationships.  2nd  ed.  New York:
     John Wiley.
 Blake, G. R., Allred, E. R., Van Bavel, C.  H. M., and Whisler, F. D.
     196C.  Agricultural drought and moisture excesses in Minnesota.
     Minn. Agr. Exp. Sta.  Tech. Bull. 235, pp. 1-36.
 Burwell,  R.  E., Sloneker, L.  L., and Nelson, W. W.  1968. Tillage
     influences water intake. J. Soil Water Conserv.  23:185-87.
 Cook,  R. L.  1962. Soil management for conservation and production.
     New York: John Wiley.
 Cooke, G. W. 1969. Fertilizers in  2000 A.D. Intern. Superphosphate
     and  Compound Manufacturers' Assoc., Bull. 53, pp. 1-13.
 Corey, R. B.,  Hasler, A.  D., Lee,  G.  F.,  Schraufnagel, F. H., and
     Wirth, T. L.  Jan. 1967.  Excessive water  fertilization. Report
     to the  Water Subcommittee,  Natl.  Resources  Com.  of  State
     Agencies, Wis.
 Duley, F. L., and Miller, M. F.  1923.  Erosion  and surface  runoff
     under different soil conditions. Mo. Agr. Exp. Sta.  Bull.  63.
 Fenster, W. E., Overdahl, C.  J., and Grava, J.  1969. Guide to com-
     puter programmed  soil  test  recommendations in Minnesota.
     Minn. Agr. Ext. Serv. Spec. Rept. 1.
 Gardner,  W. R. 1965.  Movement of nitrogen in soil. In Soil nitro-
     gen,  ed.  W. V. Bartholomew and F. E. Clark, pp. 555—72. Madi-
     son,  Wis.: Am. Soc. Agron.
 Carman, W. H. 1969. Nitrogen facts and  fallacies.  Plant Food Rev.
 Haas,  H.  J., Grunes, D. L., and Reichman, G. A. 1961.  Phosphorus
     changes  in Great Plains soils  as  influenced by  cropping  and
     manure applications.  Soil Sci.  Soc. Am. Proc. 25:214—18.
 Hanson, L.  D., and Fenster, W. E.  Oct.  1969.  Phosphorus  and lake
     quality. Crops Soils.
 Hargett, N. L. 1969.  1968 fertilizer summary data Natl. Fertilizer
     Develop. Center, TVA, Muscle Shoals, Ala.
 Hauser, V. L.  1968.  Nitrates in playas.   Agr. Res.  Notes 17:15.
 Hays, O. E., Bay, C. E., and Hull, H. H.  1948.  Increased production
     on a loess-derived soil.  Am.  Soc. Agron. }. 40:1061-69.
 Hemwall, J. B.  1957.  The fixation  of phosphorus by soils.  Advan.
     Agron.  9:95-113.
 Holt, F. G.  1969.  Runoff and sediment as nutrient  sources.  Water
     Resources Res. Center Bull. 13, pp. 35-38,  Univ.  of Minn.
 Holt, R. F., Timmons, D. R., and Latterell, J. J.  1969. Accumulation
     of phosphates in water.  In press.   /.  Food Agr. Chem.
 Huber,  D. M., Murray, G. A., and  Crane, J. M.  1969. Inhibition of
     nitrification—a deterrent to nitrate nitrogen loss  and potential
    water pollution. Soil Sci. Soc. Am. Proc. In press.
 Janssen, K. A., and Wiese, R. A.  1969. The influence of 2-chloro-6-
    (Trichloromethyl) pyridine  with anhydrous  ammonia  on corn
    yield, N-uptake, and  conversion of ammonium to nitrate.  M.S.
    thesis, Univ. of Nebr., Lincoln.
Johnston,  W.  R., Ittihadieh, F., Damn, R. M., and  Pillsbury, A. F.
     1965. Nitrogen and phosphorus in tile drain effluent.  Soil Sci.
    Soc. Am. Proc. 29:287-89.
Keeney, D. R., Konrad, J. G., and Chesters,  G. 1969.  Nitrogen distri-

                               CHAPTER 9 / FERTILIZER MANAGEMENT /  157

     bution in some Wisconsin lake sediments.  J.  Water  Pollution
     Control Federation. In press.
Kilmer, V.  J., Hays,  O.  E.,  and Muckenhirn, R. J.  1944.  Plant
     nutrient and water losses from Fayette  silt loam as measured
     by monolith  lysimeters.  Am. Soc.  Agron. J. 36:249-63.
Latterell,  J.  H., Holt, R. F., and Timmons, D. R.  1969.  Phosphate
     availability in lake sediments.  Personal communications; manu-
     script in press.
MacGregor, J. M., Hanson, L.  D., and Ellis, J.  E.  1969. Unpublished
     research and personal communication.  Univ. of Minn., St. Paul.
Maderak, M. L. 1963.  Quality of waters, Minnesota—a compilation,
     1955-62.  Minn. State Dept. Conserv. Bull. 21.
	.   1965.  Chemical quality of ground ivater in Minneapolis-St.
     Paul area of Minnesota.   Minn. State Dept.  Conserv.  Bull. 23.
Martin, W.  P.  1969.  Controlling nutrients and organic toxicants in
     runoff.  Water pollution by nutrients—sources, effects  and con-
     trol.  Water Resources Res.  Center Bull.  13, pp.  39-48.  Univ.
     of Minn.
Massey, H.  F., and Jackson,  M.  L.  1952.  Selective erosion of soil
     fertility constituents.  Soil Sci.  Soc. Am.  Proc.  16:353-56.
Meek,  B. D., Grass, L. B., Willardson, L. S., and MacKenzie, A. J. Aug.
     18-22,  1969. Nitrate transformation in  a column with a con-
     trolled  water table. Abstr. Western  Soc.  Soil Sci., Wash. State
     Univ., Pullman.
Megard,  R.  O.  1969.  Diagnosing  pollution  in Lake Minnetonka.
     Water  pollution  by  nutrients—sources, effects and  control.
     Water  Resources Res. Center Bull.  13, Univ. of Minn.
Miller, M. F.,  and Krusekoff, H. H.  1932.  The influence of sys-
     tems of cropping and methods  of culture on surface  runoff and
     soil erosion.  Mo. Agri. Exp. Sta. Res. Bull. 177.
Pearson, R.  W.,  Jordan,  H. V.,  Bennett, O.  L., Scarsbrook, C. E..
     Adams, W.  E., and White, A. W.  1961.  Residual effects  of
     fall- and spring-applied nitrogen fertilizers on crop yields in the
     southeastern  United States.  USDA Bull.   1254, pp.   1-19.
Rogers, H. T.  1942.  Losses of surface-applied phosphate and lime-
     stone through runoff from pasture land.  Soil Sci. Soc. Am. Proc.
Russell, E. W.  1961.  Soil conditions and plant growth. 9th ed. New
     York: John Wiley.
Schmidt, E. L.  1956.  Soil nitrification and nitrates in waters. Minn.
     Public Health Dept. Repts. 7:497-503.
Smith, D. D., and  Wischmeier, W.  H.  1962.  Rainfall  erosion.
     Advan. Agron. 14:109-48.
Smith, G.  E.  1965. Water forum:  nitrate problems in water as  re-
     lated to soils, plants and  ivater. Mo. Agr. Exp. Sta.  Spec. Rept.
	.  1967. Fertilizer nutrients as contaminants in water supplies.
     Am. Assoc. Adv. Sci. Publ. 85, pp. 173-86.
     -.  April 9, 1968.  In Water  pollution as  related to agriculture.
     pp.  13-27.  Joint seminar, Univ.  of Mo. and Mo. Water Pollu-
     tion Board, Columbia.
Soileau, J. M.  1969.  Effects of fertilizers on ivater quality—a collec-
     tion of abstracts and references.  Natl. Fertilizer Dev. Center,
     TVA, Muscle Shoals, Ala.
Stallings, J. H.  1957. Soil conservation.   New York: Prentice-Hall.


 Stanford,  G., Ayres, A. S.,  and  Doi, M.  1965. Mineralizable soil
     nitrogen in relation to fertilizer need  of surgarcane in Hawaii.
     SoilSci.  99:132-37.
 Stewart, B. A.,  Viets, F. G., and  Hutchinson, G. L.  1968.  Agricul-
     ture's effect on nitrate pollution of groundwater.  /. Soil  Water
     Conserv. 23:13-15.
 Stout, P. R., and Burau, R. G.  1967. The extent and significance of
     fertilizer buildup in soils as  revealed by vertical distribution of
     nitrogenous matter between soils and underlying ivater reser-
     voirs. Am. Assoc. Adv. Sci. Publ. 85, pp. 283-310.
 Taylor, A.  W.  1967.  Phosphorus  and water pollution.  /. Soil  Water
     Conserv  22'228—31.
 Timmons,  D.  R., Burwell, R. E., and Holt, R. F.  1968.   Loss of crop
     nutrients through runoff. Minn. Sci. 24:16-19.
 Turner, G. O., and Goring, C. A.  I.  1966.  N-serve,  a  status report.
     Down Earth 22:19-25.
 Wadleigh,  C.  H. 1968.  Agriculture and  the quality of our environ-
     ment. USDA Misc. Publ. 1065.
 	. Feb. 4, 1968. Nitrate in soil, water and  food.  Commentator
     response to article, "Pollution hazard  may  curb fertilizer use,"
     appearing in Des Moines (Iowa) Sunday Register.
 Wagner, G. H.,  and Smith, G. E.  1960.  Recovery of fertilizer nitro-
     gen from soils. Mo. Agr. Exp. Sta. Res. Bull. 738.
 Webber, L. R.  1967. The nature  of problem: soil pollution. Ontario
     Pollution Control Conf., Toronto, Can.
 	.  1969a.  Characteristics of  soil percolates folloiving applica-
     tion of liquid manure.  1968 Progress Rept., Dept. of Soil Sci.,
     Univ.  of Guelph, Ontario, Can.
       1969b.  Animal  waste  utilization  using  undisturbed  soil
     lysimeters.   Unpublished  data  and  personal  communication.
     Univ. of Guelph, Ontario, Can.
Webber, L. R., and Elrick, D. E.  1966.  Research needs for control-
     ling soil pollution. Agr. Sci. Rev. 4:10-20.
White,  W. C.   1965.  Plant nutrient  toll  1965.  Plant Food  Rev.
     11 (4):  17-18.
White, R. E., and Beckett, P. H. T.  1964.  Studies on the phosphate
     potentials of soils. Plant Soil 20:1.


J.  T.  PESEK, Leader
R. A. OLSON,  Reporter
         R. PESEK opened the session by summarizing its objectives
as being a forum for questioning the speakers in the formal program,
a second channel  for  bringing into focus the lacking data which
should be filled in by future research, and  a means for all interested
individuals to make statements and discuss any aspect of the role of
fertilizers as water pollutants.
     A lively session among the  40 to 50 participants resulted for
the  prescribed  period.  Procedure  followed  was to read  prepared
statements which had  been submitted, followed by discussion  from
the floor.
     The initial statement by Dr.  L.  B. Baldwin of the University of
Florida Extension  Service concerned Lake Okeechobee and  the St.
John's River Basin Water  Development  Projects  which  constitute
closed water systems.  Herewith,  an attempt is being made to  mea-
sure water nutrient levels from the eutrophication standpoint which,
it is hoped, will provide  information of countrywide interest.  Most
relevant aspects of Dr.  Baldwin's  statement were as follows:

    Florida has several important agricultural  areas adjacent to large lakes
    and reservoirs which are a part of well-developed and closely regulated
    water management projects.  In the case of Lake Okeechobee (740 sq.
    miles) and  the  peat soil  farming  area (1,100 sq.  miles)  around its
    southern perimeter,  water is pumped to  the lake during wet periods,
    and  taken from  the lake  for irrigation.  The  lake itself is  contained
    by levees, and is regulated seasonally for stage  control. During periods
    of below normal rainfall, discharge  may not be necessary, and the lake
    and agricultural area function as a closed system.  This situation may
    contribute substantially to nutrient buildup in the lake.
    A 2-year  study  of the  nutrient condition in  Lake  Okeechobee was
    started in January 1969.  It is the  purpose of the study to  determine
    the level of nutrients in the lake and in  all water  entering the lake.
    This, and subsequent  studies, may show that eutrophication of the
    lake, under present  or proposed future stages, will be  accelerated by
     J. T. PESEK is Professor and Head, Department of Agronomy, Iowa
     State University.  R. A. OLSON is Professor, Department of Agronomy,
     University of Nebraska.



     nutrients from agricultural lands.  All aspects of this situation involve
     important segments of Florida's economy.
     A similar study is  underway in the St.  John's River Basin,  which
     is also part of a controlled system. These studies should produce data
     of interest and use to other areas of the country.  Subsequent studies
     of fertilizer-soil-water management  and water system management
     should also be of importance.

      Discussion on this topic centered particularly around source of
 phosphorus that might be  responsible for its  buildup in lake and
 stream waters in an area such  as Okeechobee  which is surrounded
 by peat and muck soils. Dr. Black  expressed belief from the phos-
 phorus chemistry standpoint that any phosphorus that  did accumu-
 late in this situation would not be from mineralization of the peat
 and muck but rather would come from other sources.
      Dr. George Smith noted the occurrence of  the substantial phos-
 phate deposits a short distance to the north in Florida and  questioned
 the relevance of phosphate rock origin to current considerations in
 the St. John's-Okeechobee projects.  Environmental  conditions were
 entirely different, however, and  presumably there would be no corol-
 lary between the two.
      The next statement was by Dr. Robert D. Harter of the Univer-
 sity of New Hampshire who wrote concerning the perplexing nature
 of phosphorus in surface water and its role in eutrophication.  The
 relevant portion of his statement was as follows :

    Even in  highly eutrophic lakes, the amount of phosphorus in solution
    is  small; much  less,  in fact, than is needed for plant growth. Yet,
    luxurious algal blooms are common.  Where, then, do they obtain the
    needed phosporus?
    Studies of the phosphorus cycle in  lakes are  being conducted, and
    nutrient budgets of lakes  are being worked out.  An increasing amount
    of  this type of study is needed. However, the contribution of the lake
    sediment has frequently been ignored in these deliberations.  Lake
    sediment has been  shown to have  a large adsorption capacity for
    phosphorus.  Further research is  needed on the fate  of  phosphorus
    which is unaccounted for in  nutrient budgets, and is assumed to be
    adsorbed by the sediment.

    Soil scientists have for years attempted to identify the phosphorus com-
    pounds in  soil. Long-term fertility plots have been shown  to contain
    increased amounts of hydroxyapatite, variscite, and other highly in-
    soluble phosphorus  compounds. However, there is little information
    on  the length of  time needed for formation o£ the most insoluble crys-
    tals and  the kinetics of formation.  Before  the eutrophication process
    can be completely understood and any measure of control or reversal
    initiated, we need to know whether  the same insoluble phosphorus
    compounds are formed in lake sediment. If they are not, we need to
    know why. If they are, we need to work out the kinetics of formation,
    with an  eye to increasing the rate of phosphorus fixation in highly
    insoluble forms.

     Discussion  following  this  statement  centered  on  equilibria
established between the solid/liquid  phase, the time required for

                                 CHAPTER 10 / WORKSHOP SESSION / 161

equilibria to be reached, the turnover time involved with algal uptake,
the role of carbon dioxide on algal uptake at low concentrations, and
water stratification implications on equilibria.
     A significant  observation in this  respect  is the lack  of  algal
problems with high sediment levels in the water.  This is responsible
for the fact that the high stem dams of the Missouri River and else-
where are now creating taste  and odor problems in municipal water
supply systems in their vicinities which did not exist before impound-
ment and sedimentation occurred.  Also relevant is water depth, evi-
denced by the lack of stratification in the shallow eastern part of
Lake Okeechobee compared with considerable stratification in the
deeper western part of the lake and a much greater  eutrophication
of the former.  Lake  depth also influences the problem  of bottom
rooted plants, complimented by water clarity.
     There was agreement that studies are needed to establish the
fate of phosphorus in lake  sediments, including the kinetics of for-
mation  of insoluble phosphorus compounds.  Although  some  work
was recognized as being underway in Wisconsin, Oregon, and else-
where, much more is needed in various sections of the country with
a variety of soil sediments, kinds of clay minerals, and environmental
conditions,  especially temperature.  A number  of  questions  were
raised without specific answers, to wit: (1) When and where should
sampling be done of stream and lake waters for expressing nutrient
concentrations—that  is,  a  need  exists  for  sampling  standards.
(2) How  do we best measure phosphorus in stream or lake sediments,
by water extraction?  (3) Do  bottom rooted plants serve as a phos-
phorus pump from these sediments, exuding phosphorus to algae in
the upper waters?
     The next statement by Dr. J. Lunin  of ARS-SWC accepted  that
phosphorus movement into lakes and streams is simplified by reason
of the adsorbed  state of the element  on sediments. Movement of
nitrates,  however, is a much more difficult problem.  The most perti-
nent aspects of his statement follow:

    A nitrogen balance would be highly desirable to determine. But how
    do we quantify deep percolation and denitrification losses?  We can
    study nitrogen transformation processes in the laboratory and green-
    house, with lysimeters,  and on field  plots.  Indeed, we  are studying
    only  segments of a problem. To truly evaluate the contribution of
    nitrogen fertilization  to  the  nitrate  content  of a  stream, lake,  or
    groundwater source, we must integrate  multiple effects found within
    the watershed supplying that water resource.  It is obvious that we
    must  take into consideration all  the hydrologic parameters of that
    watershed because nitrates move with water.

    The  question  is,  How  can  we evaluate agriculture's  contribution
    to the nitrate content of a given  water resource?  Let  us define the
    research required to develop and implement a workable water quality
    model for a watershed that would integrate  all climatic, agronomic,
    animal, etc., effects within that watershed.

     Discussion here recognized that nitrate buildup is usually noted
whenever streams are running high with runoff.  A key question
raised  was, How  often do  geologic sources of nitrate influence re-


  ported stream values, especially with  the high runoff  conditions?
  There was  group concurrence that a  great deal of deep profile in-
  vestigation  is needed for tracing the course of nitrate from the top-
  soil to  the groundwater.  North  Dakota,  for example,  commonly
  finds a pool of nitrate at the  2-foot depth, more or less.  There  may
  well be  similar accumulation zones at  considerable depths in other
  regions that are of rather ancient origin.
      The statement of Dr. James P. Law, Jr., Research Soil Scientist
  of  the FWPCA,  made particular reference  to  nitrate  buildup  in
  irrigated areas.

     The switch to  high-value crops, increased fertilization rates, and in-
     creased irrigation contributes to increased rates of water quality deg-
     radation, especially where  shallow  groundwater exists  as  the  only
     dependable supply for rural  domestic, municipal, and livestock require-
     ments.  These  facts suggest the need  for  serious scrutiny of present
     fertilizer application methods and rates.
     The  time-worn practice of applying fertilizer for entire crop needs as
     one or two slug-feedings during the growing season could very well be
     shown to be both wasteful  and impractical.  In tile-drained  areas it
     has, in fact, been shown that large percentages of the fertilizer nutri-
     ents applied are lost from cropland  in  the  drainage water. Other
     studies have shown increased crop yields by adding fertilizer  require-
     ments in small increments throughout  the growing season—for exam-
     ple, irrigating  grain crops  with  sewage effluents containing limited
     quantities of nutrients (Ref: A.  D. Day and co-workers  in Arizona).
     Fertilizer elements  in excess of immediate crop needs  are subject to
     loss by leaching below the root zone and eventual occurrence as pol-
     lutants in water supplies, both surface and groundwater.
     The following are suggested as  areas worthy of research,  with the ob-
     jectives of correcting some of the present  pollution problems relative
     to fertilizer application methods and rates:
     I.  Subsurface  irrigation lends itself  to  automation and  much more
        efficient water use, which can be beneficial in controlling leaching
        losses of fertilizer elements.  The  control of surface  evaporation
        in subsurface systems also  alleviates the salinity problem associ-
        ated with irrigation return flow.
     2.  Spoon-feeding fertilizer elements  in small increments throughout
        the growing season would greatly lessen the possibility of wasteful
        losses of fertilizer to surface and groundwater supplies. Economic
        benefits  of  fertilizer  applications  would  be increased.   Soluble
        fertilizer fed directly in  the irrigation water is an example.  The
        closely controlled application  of subsurface systems would be a
        beneficial method.
     3.  Further studies into  application of  slew-release fertilizers by  con-
        ventional methods are suggested. The objectives should be to maxi-
        mize fertilizer benefits and minimize environmental pollution.
    4.  Control of excess plant nutrien's arising from fertilizer application
        depends on a better understanding of the movement and ultimate
        fate of these materials.  Studies aimed  at clarification of nutrient
        transport and  deposition  mechanisms may furnish  new leads to
        better control.

     Discussion  following  Dr. Law's paper was concerned especially
with determining what is economic rate of fertilizer application with-

                                  CHAPTER 10 / WORKSHOP SESSION / 163

out building excess residual in the soil.  It is  common opinion  that
some nominal excess in application  rate, as in the order of 50%,
is necessary, due to portions of the soil root zone being dry during
parts of the season.  The pertinent question then is just how much
nutrient exists residually in the entire rooting profile  at  the  begin-
ning of the crop  season for determining what would be the economic
rate of application.
     A primary  question from this area  is, How  do  we  go about
measuring fertilizer influence  on groundwater? Some of the barom-
eter watersheds  as  in  Oregon and the Treynor watershed in Iowa
may be revealing in the near future.
     The statement  of Dr. Ronald G.  Menzel of ARS-USDA  was
similar to that posed by Dr. Lunin, as follows:

    Nitrate  concentrations in groundwater  or  surface water  mean  very
    little by themselves. One must  understand  the dynamics of each
    situation. Where is the nitrate coming from? Where is it going? How
    rapidly?  Only by answering these questions  can we relate fertilizer
    practices to water contamination.  Therefore, it appears that measure-
    ments of groundwater movement,  chemical and biological transforma-
    tions of nitrogen, and gaseous losses of nitrogen are critically needed.
    One major problem is interrelating the different measurements involv-
    ing nitrogen  transformations  and movement. Those  measurements
    that have to be made in the  laboratory must somehow be extrapolated
    to  field conditions. For example, it may be necessary to estimate deni-
    trification losses in the field from laboratory measurements.  Can these
    be made more realistic by  increasing  sample size, controlling com-
    position of the gaseous and aqueous phases, increasing static pressure,
    or by other means?  At the same  time, we  need to  attempt direct
    measurements of denirrification in the field. Possibly an indicator re-
    action, similar to the reduction of acetylene as an  indicator of nitrogen
    fixation, can be found for denitrification. If so,  the difficulty of  dis-
    tinguishing denitrified nitrogen from atmospheric nitrogen  might be

     Discussion in this case brought out  that there has been  an in-
crease of about 15% in recent years of water supplies in  Iowa with
greater than 45 ppm nitrate. An interesting proposal for the immedi-
ate locality was one that would take all of the wastewater from  the
city of Ames, Iowa, which now goes into water courses and tise it year
around for irrigating some 1,000  acres  of  land  in the immediate
vicinity. Thereby, stream pollution would be alleviated at the same
time that many of the  fertility requirements of a  substantial area of
land were taken care of.
     Further discussion centered around  ways of removing nitrate
that has accumulated in a soil zone before it reaches the underlying
groundwater. One under investigation is  the addition of  an energy
source to an anaerobic zone where nitrate has accumulated to pro-
mote denitrification.
     A statement  by Dr. T. R.  Smith of the FWPCA supplied data on
nitrate and tile drains in streams of Illinois as follows :

    Water discharged from tile drains in prairie soils in Vermilion County,
    Illinois, was studied in the spring of 1968.


     In the Middle Fork Vermilion River Basin, two tile drains averaged
     13.5  and  17.3 mg/1 nitrate nitrogen and at the same time the river
     averaged  9.1 mg/1. At baseflow and with no tile  discharge, the river
     contained  0.24 mg/1 nitrate nitrogen.  The  North Fork Vermilion
     River Basin yielded similar data.
     The data  indicate that most of the nitrate was coming from agricul-
     tural land and that it was a widespread condition, otherwise, the river
     would have had  a much  lower  nitrate concentration during  spring
     Nitrate losses in these concentrations pose the possibility of polluting
    reservoirs and groundwater supplies.
     It appears that research may be  needed on this matter  to determine
    whether nitrates could be used more efficiently, with less being  lost in
    drainage water and at the same time maintain high crop yields.
    This problem could occur anywhere in the humid prairie region.

     Complementary  to  this  statement  was  a report  from  Story
 County, Iowa, of 5 to 40 ppm nitrate nitrogen in tile drains. It was
 further contended that nitrate has  been increasing steadily in  rivers
 of Illinois in  the last 10 to 20 years, much more rapidly during the
 last  5 years, and especially in the most productive agricultural  areas
 of the  state.  These increases coincide closely with the pyramidal
 growth in fertilizer  nitrogen consumption  during the  interval in-
     Acknowledged was the need to study again the amount of  nitro-
 gen  received in  precipitation  under modern conditions.  Results
 could be quite different from those obtained early in the century.
     From these discussions the following summary statements and
 questions evolved:

 1. Recognizing  that phosphorus  accumulates  in  water  largely
   through sediments, how do we go about reducing the  phosphorus
   level maintained in the equilibrium solution?
 2. We do not know  with certainty  the source of nitrogen in waters.
   A good deal of research  is  needed for locating  the  source and
   means of abatement.
 3. What quality of  water should the public have reasonable  right
   to expect,  keeping in mind the services demanded and the quality
   levels attainable in relation to economic considerations?
4. It would be  most helpful if agronomists, engineers, and hydrolo-
   gists would work together  closely in solving  the problems in-
5. It should  be made clear  that controls on  the use of fertilizers
   would necessitate some radical  changes in our American eating
   habits, to the very great dissatisfaction of many.  Fertilizers have
   done much toward making this the best fed nation on  earth.



        . s man embarked on global travel during the eighteenth and
nineteenth centuries, a number of events occurred that had immense
consequences in relation to pest control.  The world was searched for
new plants to adorn the greenhouses that were part of every gentle-
man's residence.  These plants brought new pests that flourished in
their new environments. Similarly, other pests were distributed by
shipments of infested food, grain, and other products.  In fact, most
of today's major pest  control problems exist because of man's igno-
rance and indiscretion. Attempts to control  these problems led to the
development of chemical pesticides.  Reviewing the history of some
of these early developments (Ordish,  1968) will prepare  us to  con-
sider a few examples of modern insecticides.
     Until about 1840 most farmers regarded pests as something one
had to accept, as the will of God. By the  late 1840s M. Grison of
Versailles discovered that lime-sulfur was a cure for powdery mildew,
Uncimda necator, a serious pest of grapes that came from America.
Soon after it was discovered  that  the disease could be arrested by
dusting plants with sulfur. This was the first large-scale successful
use of chemicals for pest control.
     The next significant step occurred when pioneers introduced the
potato plant to beetles, Leptinotarsa decemlineata,  feeding on  wild
solanaceous plants  growing  on the  eastern  slopes of  the  Rocky
Mountains from Canada to Texas. This beetle, soon known  as the
Colorado potato beetle, displayed  a  strong preference for its  new
food, the potato. The  beetle began  spreading eastward at an average
rate of  about 85 miles a year, often  destroying entire potato crops
wherever it appeared.  Virtually nothing checked the  multiplication
and spread of the beetle until about 1865 when an arsenic-containing

     PAUL A. DAHM is professor of Entomology, Department of Zoology
     and Entomology, Iowa State University.
     Journal Paper No. J-6509 of the Iowa Agriculture and Home Econom-
     ics Experiment Station, Ames.  Projects No. 1351, 1435, and 1686.
     Preparation of this paper was supported by Public Health Service Re-
     search Grant ES-00205  from the Division of Environmental Health
     Sciences and North Central Regional Project NC-85.


 chemical known as Paris green was used as a spray on potato plants
 to kill  the beetles.  Although Paris green was quite toxic and likely
 to injure plant foliage,  it remained  the leading stomach poison for
 insect control until the  introduction  of lead arsenate in 1892.
      A combination of copper sulfate and lime, subsequently called
 Bordeaux mixture, was  discovered by accident in the  1880s to be an
 effective fungicide for the control of  downy mildew, Plasmopara viti-
      Chemical control of pests was well launched by these discoveries
 during the latter part of the nineteenth  century.  At  the  Columbian
 Exposition in Chicago in 1893 there were  some 42 patented insecti-
 cides offered by several manufacturers.
      Until 1940 insecticides  consisted mostly of arsenicals, fluosili-
 cates, plant-derived chemicals, various petroleum products, synthetic
 thiocyanates,  and several  fumigant chemicals.   Discovery  of the
 broad-spectrum insecticidal properties of p,p'-DDT and y-HCCH (lin-
 dane) in the  1940s stimulated a pesticide bonanza.  The  millenium,
 however, had not  arrived.  When  populations of both harmful and
 useful insects were drastically reduced  by these  modern  chemicals,
 nontarget arthropods occasionally became pests because their preda-
 tors were no longer plentiful enough  to reduce their populations.  In-
 secticide-resistant strains of more than 200 species of arthropod pests
 also developed, owing to chemical selectivity of the new insecticides.
 Although benefits from modern pesticides are manifold, their use has
 been progressively questioned, especially since publication of Silent
 Spring  by Rachael Carson (1962).  We are now at  a stage at which
 people from several disciplines and with different expertise are look-
 ing critically at many facets of pesticide use.  Also, a variety of pest-
 control methods are being examined with the hope of reducing some
 of the problems caused by chemical agents.
     In  1966  over  half  of all U.S. farmers used  weed-,  insect-, or
 disease-control  chemicals on their crops.  In  this same year about
 29% of the farmers used insecticides on one or more crops. But only
 5%  of the crop, pasture, and range acres, or  about 12%  of the crop
 acres excluding pasture and rangeland.  were  treated  for  insect
 control (Fox et al., 1968).  An estimate of the use of insecticides in
 the 48 contiguous states  in the early 1960s showed that less than 5%
 of the  acreage had insecticides applied; about 0.4%  of the total
 area  generally considered favorable to wildlife had insecticides ap-
 plied; and 85% of the acreage planted by U.S. farmers to  crops each
 year was not  treated with insecticides (Hall, 1962). In actual quan-
 tities, about 156 million  pounds of insecticide products were used on
 farms in  the 48  contiguous  states in  1964.   This  amounts to
 about 70 pounds for each commercial farmer in  the  United States.
 Of the total, about 143 million pounds were used on crops (including
 crops, pasture, rangeland, and land in summer fallow) and 13 million
pounds for other purposes (principally livestock and livestock build-
ings).  Although alternative methods of controlling insect pests are
being developed and employed, it has been estimated that conven-
tional insecticides are still needed to control  80 to 90%  of  insect
problems affecting agriculture (Knipling,  1969).


    The abundance and mobility of water and its solvent properties
nave resulted in a variety of relationships between water and insec-
ticides.  Fundamentally, water can transport insecticides, and insec-
ticides can pollute water.  Many insecticides  are applied to  plants
or soil for protection or beautification.  Such applications are made
to fields, lawns, orchards,  forests,  gardens, greenhouses, nurseries,
and shrubs.  Although soil is the principal recipient of insecticidal
chemicals,  water  is  their  principal  distributor  after  application.
Insecticides may pollute water when they are applied to areas harbor-
ing insects and related arthropods and to domestic animals and their
wastes.  Insecticidal pollution of water may also occur when man
accidentally or irresponsibly misuses  these  chemicals.  Back-siphon-
ing of spray materials into  wells when filling spray equipment, dam-
age to containers  of insecticidal chemicals  in transit, improper dis-
posal of insecticides in all forms, excessive applications, and various
misapplications are examples of these misuses.  Occasionally,  indus-
trial wastes containing insecticides may lead to water pollution.  And
there is continuous cycling of small quantities of insecticides by vol-
atilization  from the earth  into  the  atmosphere and  precipitation
back onto soil and water.
     The three major classes of insecticides presently in use are chlo-
rinated  hydrocarbons,  organophosphates, and carbamates.  Of  the
eight insecticides used most in the United States in 1964, four were
chlorinated hydrocarbons (DDT, DDD, aldrin, and toxaphene), three
Yv'ere  organophosphates (methyl parathion, parathion,  and mala-
fhion), and one was a carbamate (carbaryl) (Table 15, Eichers et al.,
1968). These will serve as  examples around which to discuss  the
metabolism of insecticides.

     Both praising and damning declarations have been made about
DDT since its introduction as an insecticide in the 1940s. Campaigns
against this  chemical have  recently been  waged  so vigorously in
communication  media and in legislative and  judicial branches of
our government  that there is considerable doubt that DDT will survive
is an insecticide. Mankind is giving a pragmatic twist to the future
use of DDT by applying the Socratean adage, 'To know is to suffer."
DDT has probably been  studied  more intensively  and extensively
than any  other synthetic chemical. It is one of the cheapest organic
pesticides. Its chemical  stability and  biological effects have  been
praised or criticized, depending upon how one reacts to the need for
and presence of this chemical in the  environment.  The  principal
metabolites of DDT are well known (Fig.  11.1).  The  best  known
metabolic route  involves  dehydrochlorination of DDT  to DDE, 1,1-
dichlorc-2,2-bis(p-chlorophenyl)  ethylene, because   this reaction is
the primary reason for resistance of insects  to DDT (Sternburg et al.,
1953).  Strains of insects resistant to DDT have a  large proportion
of their population possessing an enzyme that can dehydrochlorinate
DDT to less-toxic DDE (Lipke and Kearns, 1960). Susceptible strains

                         Cl-/  "V")   CHCC1

                DICOFOL                       DDD
      FIG. 11.1.  The principal  metabolites of DDT.
 of insects have relatively few individuals with this biochemical pro-
 ficiency;  hence,  they succumb to the insecticide.  DDE is  also the
 most common metabolite of DDT found in avian tissues.  It has be •:;;.•
 suggested that DDE plays  a major role  in causing thinness of e,/-
 shells in  certain species of birds (Heath et al., 1969),  possibly I-
 inducing  hepatic microsomal metabolism of  steroids.  One or n,<
 earliest metabolic discoveries about DDT was its conversion  to DD/';
 bis(p-chlorophenyl)  acetic acid,  in mammals (White  and Swe^ne;,
 1945; Jensen  et' al.,  1957;  Durham et al.,  1963), including x:.i\.
 (Neal et al., 1946;  Durham et al., 1965).  DDA is readily excreted h
 the   urine.    Biological  reductive  dechlorination   of  DDF   to
 DDD(=TDE),  l,l-dichloro-2,2-bis(p-chlorophenyl) ethane, has  tu-K
 proved comparatively recently (Finley  and  Pillmore,  1963; Bar!;;;
 and Morrison, 1964; Walker et al., 1965). This reaction occurs ;!;•;.•
 readily under anaerobic conditions in animal  tissues  and in n;i : v
 organisms.  It is now quite acceptable to report  DDD as a i)H:.ah.a|;h
of DDT but for many years the possibility of  this compound Leaig
formed biologically was scoffed at by some scientists.  DDD is a com-
 mercial insecticide  in its own right. Replacement of hydrogen on u'.e
 tertiary carbon of DDT by a hydroxyl group forms a metabolite of jovr
 toxicity to insects and mammals but cf high toxicity to miles  r^L
moto, .1959; Agosin et al., 1961).  A commercial miticide called die-.';!-!
(Keltharie®, 4,4'-dJchloro-a-[trichloromethyl] benzhydrol)  is id-'in '.(•:•.'


TABLE 11.1   Toxicity of DDT  and metabolites to  adult male rats.
DDT (technical) ....
Dicofol (=Kelthane®)
DDD (-TDE) 	

Acute Oral Toxicity LD.-/>

Relative Toxicity
1 0
4 1

Source:  Gaines (1969).

to this metabolite. Another metabolite o£ DDT is DBF, p,p'-dichloro-
benzophenone (Menzel et al., 1961; Abou-Donia and Menzel, 1968);
this compound has frequently been reported in metabolism studies
with insects. Although several criteria should be used to compare the
toxicity of chemicals,  the most  complete comparison  of the toxicity
of DDT with its principal metabolites can be made on the basis of
acute oral toxicity values (Table 11.1). The toxicity of the chemicals
in this table is inversely related to the numerical values.
     Many metabolites of DDT other than the five already described
have been reported (Abou-Donia and Menzel, 1968).   A  recent dis-
covery about DDT metabolism is the in vivo isomerizations that lead
to the formation of p,p'-DDT from feeding o,p'-DDT to rats (Klein et
al., 1964) and the formation of o,p'-DDT and o,p'-DDD from feeding
p,p'-DDT  and p,p'-DDD to young chickens (Abou-Donia and Menzel,
1968). The approximately 20%  of o,p'-DDT in technical DDT is con-
verted to p,p'-DDT and then to p,p'-DDE  in living avian tissue;  in the
anaerobic conditions after death, o,p'-DDT is metabolized to o,p'-DDD
(French and Jefferies, 1969). The absence of o,p'-DDT and metab-
olites in field specimens is ascribed to the  rapid  rate of breakdown
and a masking of the o,p'-DDD residue during analysis by the relative-
ly large amounts  of p,p'-DDE.  These examples illustrate  the com-
plexity of metabolism  studies and the pitfalls of interpreting analyt-
ical data.
     The  exact biochemical cause of the toxicity of DDT and related
chemicals to certain  organisms  has never been  proved.  Several
theories on how DDT acts have been promulgated.  An extensive study
of feeding DDT in the diet of rats suggested that  the  effects of DDT
depend not only on DDT but also on some unidentified secondary fac-
tor (Ortega et al., 1956).  An example of this hypothesis is the sug-
gestion that DDE is the major factor in  toxicity of DDT and that the
amount of DDE produced from DDT determines the level of toxicity
of  DDT in  different species (Bailey et al.,  1969). An earlier  study,
however,  suggested that residues of DDE were not critical in birds
that died from DDT (Stickel et  al.,  1966).  These  examples are cited
to illustrate the confusion about the toxicity of DDT, its metabolites,
and related compounds. The estrogenic activity of o.n'-DDT (Bitman
et al., 1968) and the  conversion of analogues of  DDT to estrogenic
metabolites (Welch  et al., 1969) are interesting  new developments
that may link DDT metabolism  studies with the  claim that this in-


 sccticide is  the  indirect cause of a reduction of eggshell thickness
 associated with  failing reproduction and population decline  of cer-
 tain predatory birds (Stickel, 1968; Porter and Wiemeyer, 1969).  In
 the past, surveys have not usually distinguished between the presence
 of p.p'- and o,p'-DDT. A change in analytical procedures could clarify
 how widespread  the latter isomer really is.
      Although the use of DDT as an insecticide is declining, the en-
 vironment will continue to be monitored for this  chemical and  its
 metabolites. A review of the voluminous data on DDT, its analogues,
 and its metabolites in the environment is beyond the  scope of this
 presentation.  I   predict,  however,  that  interpreting  these  data
 in terms  of biological effect or no-effect will provide  a continuing de-
 bate for many years. Methoxychlor, 2,2-bis(p-methoxyphenyl)-l,l,l-
 trichloroethane,  is an insecticidal analogue of DDT that has  much
 lower mammalian toxicity than DDT.  For example, the acute oral
 LD-0 of methoxychlor to rats seems to be scmewhere between 5 to 7
 g/kg (Smith et al., 1946; Hodge et al., 1950).  This insecticide shows
 little tendency to be stored in the body fat and other lipids. If there is
 a  general ban on the use of DDT, methoxychlor may  serve as a re-
 placement for DDT in a few pest control situations, but the organo-
 phosphate and carbamate insecticides currently  available will prob-
 ably fill most of the gaps left by withdrawing DDT from pest control

     Aldrin is one of a group of chlorinated hydrocarbon insecticides
that also includes dieldrin, endrin, and heptachlor.  Interrelationships
of structure and activity are known for about 500 of these so-called
cyclodiene compounds (Soloway, 1965).  The following comments
draw upon recent  reviews of  the  metabolism of  these  insecticides
(Brooks* 1966,  1968, 1969; Korte, 1968).  The  1969 review by Brooks
is especially comprehensive in its treatment of the subject. Biological
epoxidation of aldrin (Fig. 11.2), isodrin, and heptachlor produces
dieldrin, endrin,  and heptachlor  epoxide, respectively.  The epoxida-
tion  of these  insecticides is  interesting because  the  metabolites.
dieldrin, endrin, and heptachlor epcxide, are  about as toxic as,  and
more persistent than, their parent compounds (Gaines, 1960, 1969).
 Cl   PI
11.2.  Epoxidation  of  aldrin to dieldrin.


     Many  efforts  have  been  directed  toward finding metabolic
 products of dieldrin, endrin, and heptachlor epoxide (and other mem-
 bers of this group of insecticides).  Until recently these epoxides were
 considered stable in metabolizing systems.  It was sometimes thought
 that the insecticides were stored in fat, as the epoxides in those in-
 stances in which epoxidation could occur, and ultimately excreted in-
 tact in the feces.  It is now known that these compounds are amenable
 to further metabolism, including hydroxylation, hydrolytic (or oxida-
 tive) elimination of chlorine atoms (when present) other than those
 of the intact hexachloronorbornene nucleus, and hydrolysis of epoxide
 rings.  In vivo studies with rats have shown that aldrin is converted
 to polar  metabolites from  either dieldrin formed  from aldrin  or
 dieldrin administered separately (Datta et al., 1965; Korte,  1968).
 Metabolites of dieldrin have  been found  also  in urine  from man
 (Cueto and Hayes,  1962)  and rabbits (Korte, 1968).  A more recent
 study revealed two metabolites  from rats fed a diet containing 100
 ppm of dieldrin (Richardson et al.,  1968).  The mixed-function oxi-
 dases that metabolize so many foreign substances are also involved in
 cyclodiene metabolism in insects and mammals in vitro. The nature
 of the metabolites  so far isolated, the parallel between microsomal
 enzyme induction  and increased  metabolism  in vivo  observed  for
 some mammals, and the action of  synergists in insects provide a link
 between the in-vivo and in-vitro  processes.
     The toxicology and no-effect levels of aldrin and dieldrin have
 been extensively reviewed by a panel selected by the the Secretary of
 Health, Education,  and Welfare from nominations by the National
 Academy  of Sciences (Hodge et  al., 1967).  The following statements
 are from  the  summary of this  review. The acute oral toxicity  for
 either  aldrin or dieldrin ranged from 20 to 70 mg/kg among  12
 species of animals; the estimated lethal dose for man is  approxi-
 mately 5  g. The mortality among several species of animals, after
 either repeated short-term or chronic doses, ranged from 0.5 to 300
 ppm.  No body weight changes occurred among several species  of
 animals at 2 ppm or less in the diet.  Pathological conditions were ob-
 servable at  levels in the  diet ranging from 0.5 to  10 npm among
 several species of animals. And, typical diets in Eneland and in  the
 United States are estimated to contain 1 to 2 ppb of dieldrin; dieldrin
 concentrations in human fat probably average about 0.2 ppm.
    Insecticides of the cyclodiene group have had low residue toler-
 ances imposed upon them from  the beginnings of their use.  Further
residue tolerance restrictions have  been placed on these chemicals in
recent  years. Environmental persistence and unfavorable biological
effects of  some of the cyclodiene insecticides and development of re-
sistance to these insecticides  by some species of insects and other
arthropods suggest  that the  use of these insecticides will decline.

    An anomalous situation exists with respect to our knowledge of
toxaphene, an insecticide used more extensively in the United States

 in 1964 than any other insecticide (Eichers et al., 1968).  The exten-
 sive use of toxaphene, since it became available for commercial use
 about 1947 (Parker and Beacher, 1947), has not been  accompanied
 with published information  about its composition and metabolism.
 Toxaphene is a chlorinated camphene having an approximate empiri-
 cal formula of C]0H]0C1S; it contains 67 to 69% chlorine. Toxaphene
 is a general convulsant that acts  on the central nervous system. In
 this respect it is similar to DDT and the cyclodiene insecticides. In
 contrast to them,  however,  little is known about the metabolism of
 toxaphene. It is probably slowly detoxified in the liver'. This assump-
 tion is based on its close chemical relationship to camphor, which is
 detoxified  in the liver, and the isolation  of ethereal sulfate and glu-
 curonic acid conjugates of toxaphene in  the urine (Conley, 1952).
     Although toxaphene is a highly chlorinated organic compound,
 and hence readily detected  by electron-capture gas chromatography
 (GLC), there is a paucity of residue and metabolism data that distin-
 guish between components  of the  technical product.  Residues of
 toxaphene cannot  be determined quantitatively in environmental
 samples by GLC because toxaphene  is a  mixture of compounds that
 gives a continuum of curves with a wide spread of retention times.
 This results  in mutual  interference  from many common pesticides.
 This difficulty is illustrated in Figure 11.3 by GLC curves of toxa-
 phene,  DDT, and a  combination of the two insecticides (Benevue
 and Beckman, 1966). These GLC curves are especially pertinent be-
 cause one  of the major markets for  toxaphene has been a 2:1 com-
 binat'on of toxaphene and DDT as an insecticide for use on cotton.
     The difficulties of estimating the components  of toxaphene are
 illustrated  in studies of the  persistence of toxaphene in lakes in
 which it has been used as a substitute for rotenone to reduce rough
 fish populations (Johnson et al., 1966; Terriere  et al., 1966). Various
 formulations of toxaphene  showed slightly different gas chromato-
 grams, the components of toxaphene seemed to be degraded  at dif-
 ferent rates, and  the components had  different  toxicities for fish
 (Johnson et al.,  1966).  It is clear that until the chemistry and me-
 tabolism of toxaphene are better known, the fate of this  insecticide
 in natural  waters will be poorly understood.
FIG. 11.3.   GLC  curves of
toxaphene,  DDT, and a
combination of the two in-
secticides.   (Benevue   and
Beckman, 1966.)

                            (R0)2 POX
                                                    • CHOLINE
                        NERVE  SYNAPSE
     FIG. 11.4.  The principal toxic action of organophosphate insecticides.


     Parathion, methyl parathion, and  malathion are members of
a large class  of  organophosphate insecticides.  "Organophosphate"
is often employed as  a generic term  to cover all the toxic organic
compounds  containing  phosphorus.  Organophosphates are  more
specifically designated as phosphates, phosphoiiates, phosphorothio-
nates,  phosphorothiolates,  phosphorodithioates,  phosphoramidates,
etc., depending upon  the atoms attached to  the phosphorus; for ex-
amples, see O'Brien (1960).  There is an extensive literature on these
compounds, including books concerned exclusively with  organophos-
phates (O'Brien, I960; Heath, 1961).
     The most important reaction of Organophosphates is with  acetyl-
cholinesterase, an enzyme involved in the transmission of nerve im-
pulses  (Fig.  11.4). Acetylcholine.  a chemical mediator of nerve im-
pulses  at synapses, is normally hydrolyzed  very  quickly by  acetyl-
cholinesterase.  Any disruption  of this reaction causes acetylcholine
to accumulate.  Acetylcholine is  itself a moderately toxic chemical.  It
acts as a poison and causes well-known  symptoms of poisoning.  Or-
ganophosphates with a P-_=:0 structure irreversibly react  with cholin-
esterases, preventing these  enzymes from accomplishing their hydro-
lytic function.
     Parathion was introduced about 1944 in Germany.  As recently
as 1964, methyl parathion (the  dimethyl analogue of parathion) and
parathion were the most widely  used Organophosphates in the United
States (Eichers et al., 1968). Although methyl parathion and para-
thion are chemically very similar, each seems  needed to control dif-
ferent species of pest insects; therefore, both  exist on the commercial
market. Both are quite toxic chemicals; for example, their acute oral
LD-o values with male rats are 14 and 13 mg/kg for methyl parathion
and parathion, respectively (Gaines, 1969).  Because their properties
and metabolism are so nearly alike, further attention will  be given
only to parathion.
     The toxicity of parathion develops from  a  desulfuration reaction

(C2H50)2  P-0

                      S           0
             (CH30)2 P-S - CMC*	
     FIG.  11.5.   Some points of metabolic attack on parathion  and  mala-
that changes P=S  to P—O (Fig. 11.5). This converts parathion to
paraoxon, a compound with strong cholinesterase-inhibiting proper-
ties.  This  important intoxication  reaction occurs with all  organo-
phosphates having  a P=S  structure and is especially catalyzed by
liver microsomal enzymes.  These enzymes require reduced nicotina-
mide adenine dinucleotide and oxygen  with in vitro reactions.  Two
primary degradation reactions of parathion and paraoxon are shown
in Figure 11.5. One of these is a hydrolytic reaction that yields di-
ethylthiophosphoric acid and p-nitrophenol. This reaction is catalyzed
by liver microsomal oxidases similar, if not identical, to those that ef-
fect conversion of parathion to paraoxon (Nakatsugawa et al., 1969).
The p-nitrophenol from parathion degradation appears in urine and
provides a sensitive indicator of exposure to parathion before any sig-
nificant decline in cholinesterase activity can be  detected (Davies et
al., 1966).  The second degradation reaction illustrated for parathion
in Figure 11.5 involves reduction of the p-nitro  group of parathion
and paraoxon  to  form  aminoparathion  and aminoparaoxon,  re-
spectively.  This  reaction  occurs under  a  variety  of  conditions
(O'Brien, I960; Lykken and Casida,  1969;  Mick and  Dahm, 1970).
     The metabolism of parathion has been investigated with several
species of animals,  plants,  and microorganisms  (O'Brien,  I960; El-
Refai and Hopkins, 1966).  Other reported metabolites of parathion
and paraoxon include desethyl parathion, desethyl paraoxon, diethyl
phosphoric acid, ethyl phosphoric acid,  and phosphoric acid. All
metabolites of parathion, except paraoxon, are less  toxic than the
parent insecticide.
    Malathion was accepted for commercial use  as an insecticide in


 1952.  Since then, it has been used to control many species of insects
 and other arthropods. It is a major insecticide used throughout the
 world, owing partly to its low toxicity to mammals.  For example, its
 acute  oral LD-0 in adult male rats is  1,375 mg/kg (Gaines, 1969).
 The major intoxication route for malathion is desulfuration to mala-
 oxon (Fig. 11.5)  and inhibition of acetylcholinesterase by the mala-
 oxon produced (O'Brien,  1960).  It is the detoxication reactions that
 set malathion apart and are responsible for its remarkably low toxicity
 to mammals. The most  important reactions involve  the  hydrolysis
 by carboxylesterase of one of the two available carboxylic ethyl esters
 of malathion and malaoxon as shown in Figure 11.5 (Dauterman and
 Main,  1966). Malathion monoacid, the  major metabolite of mala-
 thion,  has been identified as O,O-dimethyl-S-(l-carboxy-2-carbethoxy)
 ethyl phosphorodithioate  (Chen et  al.,  1969).  Other detoxication  re-
 actions shown  in Figure 11.5 produce hydrolytic products,  such as
 malathion diacid, malaoxon mono- and diacids, O.O-dimethyl phos-
 phorodithioate, O,O-dimethyl  phosphorothioate, dimethyl phosphate,
 monomethyl phosphate, and phosphoric  acid (O'Brien,  1967).
     Malathion  gained  further  prominence  when Frawley  et  al.
 (1957) showed that simultaneous administrations of malathion and
 EPN, O-ethyl O-p-nitrophenyl phenylphosphonothioate,  to dogs and
 rats caused strong synergistic effects  in  the form of cholinesterase
 inhibition.  Several later  studies showed  that EPN inhibits  the car-
 boxylesterase that hydrolyzes malathion  and malaoxon. A  number
 of other combinations of organophosphates also have been synergistic
 (O'Brien,  1967).  Fears that ingestion  of mixtures  of low levels of
 organophosphates, and possibly other chemicals, as residues on foods
 might  produce  overt  symptoms  of  cholinesterase depression have so
 far proved false.
     Malathion is comparatively more toxic to insects than to mam-
 mals, seemingly because  of less effective hydrolytic detoxication by
 carboxylesterases in  insects.  EPN fails to  synergize the toxicity of
 malathion to insects, and certain strains of insects  resistant  to mal-
 athion have a high carboxylesterase level (O'Brien, 1967).
     The  metabolism of  organophosphate insecticides has received
 special attention in recent reviews by Fukuto and Metcalf (1969) and
 Lykken and Casida (1969). From  these and other sources it can be
 concluded that organophosphate insecticides include chemicals that
 range from high  toxicity  (e.g., parathion) to low toxicity (e.g., mala-
 thion).  These  insecticide molecules usually  possess  several  places
 that are  vulnerable to metabolic attack, and the metabolic products
 are more water soluble than the  parent insecticides.   Organophos-
 phates are physically and chemically less stable than organochlorine
 insecticides (e.g., DDT and dieldrin) and therefore present less of a
 hazard for environmental contamination than organochlorines.

     Carbaryl (—Sevin®) is the most widely used insecticide belong-
ing to a group  of esters of N-methyl and N-dimethyl carbamic acid.
The carbamate insecticides show somewhat erratic patterns of selec-



 FIG. 11.6.  Some points of           HYDROLYSIS -, ^  HYDROXYLAT.ON
 metabolic attack on  car-                -,      0-CNHCH
 baryl.                       EPOXIDATION


 tive toxicity to insects.  These insecticides are fairly potent inhibitors
 of cholinesterases, and the symptoms resulting from this action are
 typically cholinergic. The  inhibitory action of carbamates, however,
 is reversible, in contrast to the action of organophosphates.
     Carbaryl quite readily undergoes  several metabolic reactions
 (Fig.  11.6), including hydroxylation attack on the N-methyl group and
 two locations on the napthol ring and epoxidation followed by epoxide
 cleavage and hydrolysis on the nonphenolic ring. Each of these initial
 oxidation products subsequently  conjugates and is excreted as a sul-
 fate or glucuronide in mammals, but may persist  as a glycoside in
 plants.  Hydrolysis of the carbamyl ester linkage releases 1-napthol,
 which is rapidly  metabolized.  Additional  information  about the
 metabolism of carbaryl and other carbamate insecticides is included
in reviews by Fukuto and  Metcalf (1969)  and Lykken  and  Casida
 (1969).  Carbamates are currently viewed as  competitors of organo-
 phosphates for pest-control purposes.

     Insecticides occur in the environment because of purposive ap-
plications for pest control and because of accidents and carelessness.
The major problems with insecticides arise from the contamination of
the environment and food and the development of resistant arthro-
pod-pest populations.  The persistence  of insecticides  in the atmos-
phere, water, soil, plants, animals, and microorganisms  is being inves-
tigated.  Alterations of insecticides  occur under both metabolic and
nonmetabolic conditions.
     Knowledge of the metabolism  of insecticides is prerequisite to
their development and use for insect control.  Identification and toxi-
cological assessment of metabolic products should precede  establish-
ment of residue and other safety factors.  More basically, metabolism
studies of insecticides reveal intoxication and detoxication processes
and how these relate to physiological effects  and problems of re-
sistance. Some of the ways that organic insecticides are metabolized
in living organisms  are hydrolysis, hydroxylation,  dehalogenation,
dehydrohalogenation, desulfuration (z^cxidation), O-dealkylation, N-
dealkylation, reduction, and conjugation.  Metabolic attack occurs at
one or more sites on  an insecticide molecule.  Plants and animals
often metabolize insecticides by  similar pathways.  With some insecti-
cides, primary metabolic attack  may form compounds whose toxicity


approximately equals or is greater than  the parent insecticide (e.g.,
aldrin^dieldrin;  parathion^paraoxon).   Further  metabolism  pro-
duces compounds of much lower toxicity.  Other insecticides are de-
toxified directly (e.g., DDT-^DDE,  although DDE  may  have subtle
physiological effects on nontarget organisms).  The metabolism of an
insecticide from  administration  to target sites and  in  and  out  of
storage tissues  generally produces compounds of greater water solu-
bility to facilitate excretion of metabolites. Because the persistence of
some of our present organochlorine  insecticides (e.g., DDT, DDD,
cyclodienes)  creates  environmental problems, future insecticide  de-
velopments will probably give  special  attention to effective chemicals
that degrade  to compounds with negligible environmental effects.
     Although this review  is primarily concerned with metabolism,
numerous nonmetabolic factors exert effects on the structure and per-
sistence of insecticides.  Some of these nonmetabolic factors include
light, water,  heat, acidity  and alkalinity,  atmospheric constituents,
metal ions, and soils.  An indication of the nonmetabolic complexities
of the decomposition of insecticides is given  in a review by  Crosby
(1969). Furthermore, the solubilities of insecticides in soil and water
are especially important in relation to their movement and persistence
in the  environment.  An exhaustive search of the literature by Gun-
ther et al. (1968), however, revealed only  Limited useful data on water

Abou-Donia. M. B., and Menzel. D. B.  1968.  The metabolism in vivo
     of   l,l,l-trichloro-2,2-bis(p-chlorophenyl)  ethane  (DDT),  1,1-
     dichloro-2,2-bis(p-chlorophenyl) ethane  (DDD)  and  1,1-dichlo-
     ro-2,2-bis(p-chlorophenyl)  ethylene (DDE) in the chick by  em-
     bryonic injection and dietary ingestion.  Biochem. Pharmacol.
Agosin, M., Michaeli, D., Miskus, R., Nagasawa,  S., and Hoskins,
     W. M.  1961.  A new DDT-metabolizing enzyme in  the  German
     cockroach.  /. Econ. Entomol. 54:340—42.
Bailey, S., Bunyan, P. J., Rennison, B.  D., and Taylor, A.  1969. The
     metabolism  of  l,l-di(p-chlorophenyl)-2,2-dichloroethylene  and
     l,l-di(p-chlorophenyl)-2-chloroethylene in the pigeon.  Toxicol.
     Appl. Pharmacol. 14:23-32.
Barker, P.  S., and Morrison, F. O.  1964.  Breakdown of  DDT in
     mouse tissue.  Can. J. Zoo/. 42:324-25.
Benevue, A., and Beckman, H.  1966. The examination of  toxaphene
     by gas chromatography.  Bull. Exptl.  Contamination  Toxicol.
Bitman,  J., Cecil,  H. C., Harris, S. J.,  and Fries, G. F. 1968.  Estro-
     genic activity of o,p'-DDT in the  mammalian uterus  and avian
     oviduct. Science 162:371-72.
Brooks, G. T.  1966.  Progress in metabolic studies of the  cyclodiene
     insecticides and its relevance to  structure-activity  correlations.
     Wor/d Rev.  Pest Control  5:62-84.
	.  1968.  Perspectives  of cyclodiene metabolism.  Symposium
     on the Science and Technology of Residual Insecticides in Food


      Production with Special Reference to Aldrin and Dieldrin. Spon-
      sored by Shell Chemical Co.
      -.  1969. The metabolism of diene-organochlorine (cyclodiene)
     insecticides. Residue Rev. 27:81-138.
 Carson, R. L.  1962.  Silent spring. Boston:  Houghton Mifflin.
 Chen, P. R., Tucker, W. P., and Dauterman, W. C.  1969.  Structure
     of  biologically produced  malathion monoacid.  J.  Agr.  Food
     Chem. 17:86-90.
 Conley, B. E.  1952. Pharmacologic properties of toxaphene, a chlo-
     rinated hydrocarbon insecticide.  J. Am.  Med. Assoc.  149:1135-
 Crosby, D. G.  1969.  The nonmetabolic decomposition of pesticides.
     Ann. N.Y. Acad. Sci. 160:82-96.
 Cueto,  C., Jr., and Hayes, W. J., Jr.  1962. The detection of dieldrin
     metabolites in human urine.  /. Agr. Food Chem.  10:366—69.
 Datta, P. R., Laug, E. P., Watts, J. O., Klein, A. K., and Nelson, M. J.
     1965. Metabolites in urine of rats on diets containing aldrin  or
     dieldrin. Nature 208:289-90.
 Dauterman,  W. C., and Main, A. R. 1966. Relationship  between
     acute toxicity and in vitro inhibition and hydrolysis of a series
     of carbalkoxy  homologs of malathion.   Toxicol. Appl. Pharma-
     col. 9:408-18.
 Davies, J. E., Davis, J. H., Frazier, D. E., Mann, J. B.,  and Welke,
     J.  O.  1966.  Urinary p-nitrophenol concentrations in acute and
     chronic parathion  exposures.  Advan. Chem. Ser. 60:67—78.
 Durham, W. F., Ortega, P., and Hayes, W. J., Jr. 1963. The effect of
     various  dietary levels of DDT  on liver function, cell morphology,
     and DDT  storage  in the rhesus monkey.  Arch.  Intern. Pharma-
     codijn.  141 (1-2): 111-29.
 Durham, W. F., Armstrong, J. F., and Quimby, G.  E.   1965.  DDA
     excretion  levels.  Arch. Environ. Health 11:76—79.
 Eichers, T., Andrilenas, P., Jenkins, R., and  Fox, A.  1968.  Quanti-
     ties of  pesticides  used by  farmers in 1964.  USDA, Agr. Econ.
     Rept. 131.
 El-Refai, A.,  and Hopkins, T. L. 1966.  Parathion  absorption, trans-
     location, and conversion to  paraoxon in bean  plants.   J. Agr.
     Food Chem. 14:588-92.
 Finlev.  R. B., Jr., and  Pillmore, R. E.  1963.  Conversion of DDT  to
     DDD in  animal tissue.  BioScience 13:41-42.
 Fox, A., Eichers. T.. Andrilenas, P., Jenkins, R., and Blake, H.  1968.
     Extent  of farm pesticide  use on crops in 1966.  USDA, Agr.
     Econ. Rept. 147.
 Frawley, J. P., Fuyat, H. N., Hagan, E. C., Blake, J. R., and Fitzhugh,
     O.  G. 1957.  Marked potentiation in mammalian  toxicity from
     simultaneous  administration  of  two  anticholinesterase  com-
     pounds.  J. Pharmacol. Exptl.  Therap. 121:96-106.
 French, M. C., and Jefferies, D. J.   1969.  Degradation and disappear-
     ance of  ortho,  para isomer of of technical DDT in living and
     dead avian tissues. Science 165:914—16.
Fukuto, T. R., and Metcalf, R.  L.  1969.  Metabolism of  insecticides
     in  plants and  animals.  Ann. N.Y.  Acad. Sci.  160:97-111.
Gaines,  T. B.  1960. The acute toxicity of pesticides to rats.  Toxicol.
     Appl Pharmacol. 2:88-99.
	.  1969. Acute toxicity of pesticides. Toxicol Appl  Pharmacol
Gunther, F. A., Westlake, W. E., and Jaglan, P. S.   1968. Reported


     solubilities of  738  pesticide  chemicals  in water.  Residue Rev.
 Hall, D. G.  1962.  Use of insecticides in the  United States.  Bull.
     Entomol. Soc. Am.  8:90-92.
 Heath, D. F. 1961. Organophosphorus poisons.  New  York: Macmil-
     lan (Pergamon).
 Heath, R. G., Spann, J. W., and Kreitzer, J. F.   1969.  Marked DDE
     impairment of mallard reproduction in controlled studies. Na-
     ture 224:47-48.
 Hodge,  H. C., Maynard, E.  A., Thomas, J.  F.,  Blanchet, H.  J., Jr.,
     Wilt, W. G., Jr., and Mason K. E. 1950. Short-term oral toxicity
     tests of methoxychlor (2,2 di-(p-methoxy phenyl)-l,l,l-trichlor-
     ethane) in rats and dogs.  J. Pharmacol. Exptl. Therap. 99:140-
 Hodge,  H. C., Boyce, A. M., Deichmann, W. B., and Kraybill, H. F.
     1967.  Toxicology  and  no-effect levels of  aldrin  and dieldrin.
     Toxicol. Appl. Pharmacol. 10:613-75.
 Jensen, J. A., Cueto, C., Dale, W. E.,  Rothe, C. F., Pearce, G. W., and
     Mattson, A. M.  1957. DDT metabolites in feces and bile of rats.
     J. Agr. Food Chem. 5:919-25.
 Johnson, W. D., Lee, G. F., and Spyridakis, D.  1966.  Persistence of
     toxaphene in treated lakes.  Intern. J. Air  Water Pollution 10:
 Klein, A. K., Laug,  E. P., Datta, P. R., Watts, J.  O., and Chen, J.  T.
     1964.  Metabolites:  reductive dechlorination of DDT and DDD
     and isomeric transformation of o,p'-DDT to p,p'-DDT in vivo.
     J. Assoc. Official Agr. Chemists 47:1129-45.
 Knipling, E. F.  1969.  Alternative  methods  of controlling  insect
     pests. Food Drug Admin. Papers 3 (1): 16-24.
 Korte, F. 1968.  Metabolism of aldrin, dieldrin, and endrin.  Sympo-
     sium on the Science and Technology of Residual  Insecticides in
     Food Production *vith Special Reference to Aldrin and Dieldrin.
     Sponsored by Shell Chemical  Co.
 Lipke, H., and Kearns, C. W.  1960. DDT-dehydrochlorinase.  Advan.
     Pest Control Res. 3:253-87.
 Lykken,  L., and Casida, J. E.  1969.  Metabolism of organic insecti-
     cide chemicals. Can. Med. Assoc. /.  100:145-54.
 Menzel,  D. B., Smith, S. M., Miskus, R., and Hoskins, W. M.  1961.
     The metabolism  of C14-labeled DDT in the larvae, pupae, and
     adults of Drosophila melanogaster.  J. Econ. Entomol. 54:9—12.
 Mick, D. L., and Dahm, P. A. 1970.  Metabolism  of parathion  by
     two species of Rhizobium.  J.  Econ. Entomol. In press.
 Nakatsugawa, T., Tolman, N. M., and Dahm, P. A.  1969. Degrada-
     tion of parathion in the rat.  Biochem. Pharmacol. 18:1103—14.
 Neal, P.  A., Sweeney, T. R., Spicer, S. S., and von Oettingen,  W. F.
     1946.   The  excretion  of  DDT (2,2-bis-(p-chlorophenyl)-l,l,l-
     trichloroethane) in man,  together  with clinical observations.
     Public Health Rept. 61:403-9.
 O'Brien, R. D. 1960. Toxic phosphorus esters. New York:  Academic
	.   1967.  Insecticides.   Action  and metabolism.  New  York:
     Academic Press.
 Ordish,  G.   1968.  150 years  of crop  pest control. World Rev. Pest
     Control 7:204-13.
 Ortega, P., Hayes, W. J., Jr., Durham, W. F.. and Mattson, A.  1956.
     DDT in the diet of the rat. Public Health Monograph 43.


 Parker, W. L., and Beacher, J. H.  1947.  Toxaphene, a chlorinated
     hydrocarbon  with insecticidal properties.  Univ. of Del. Bull.
     264, Tech. 36. Newark, Del.
 Porter, R. D., and Wiemeyer, S. N. 1969. Dieldrin and DDT:  Effects
     on   sparrow  hawk  eggshells   and  reproduction.    Science
 Richardson, A., Baldwin, M., and Robinson, J.  1968. Identification
     of metabolites of dieldrin (HEOD)  in  the  faeces and urine  of
     rats. /. Sci. Food Agr.  19:524-29.
 Smith, M. L, Bauer, H., Stohlman, E. F.,  and Lillie, R. D. 1946.  The
     pharmacologic action of certain  analogues and  derivatives  of
     DDT. /. Pharmacol. Exptl. Therap.  88:359-65.
 Soloway,  S. B.  1965.  Correlation between biological activity  and
     molecular  structure   of the cyclodiene insecticides.   Advan.
     Pest Control Res.  6:85-126.
 Sternburg, J., Vinson, E.  B., and Kearns,  C.  W.  1953. Enzymatic
     dehydrochlorination of DDT by resistant flies. /. Econ. Entomol.
 Stickel, L. F.  1968.  Organochlorine  pesticides  in the environment.
     Bur. of Sport Fisheries and  Wildlife, Spec. Scientific Rept.,
     Wildlife No. 119, Wash., D.C.
 Stickel, L. F., Stickel, W. H., and Christensen, R.  1966.  Residues  of
     DDT in brains and bodies of birds  that died on dosage and  in
     survivors.  Science 151:1549-51.
 Terriere, L. C., Kugemagi, U.,  Gerlach, A. R., and Borovicka, R.  L.
     1966. The persistence of  toxaphene in lake water and its up-
     take by  aquatic  plants and  animals.  /. Agr. Food  Chem.
 Tsukamoto, M.  1959.  Metabolic  fate  of  DDT in Drosophila melano-
    gaster. I.  Identification of  a non-DDE metabolite. Botyu-Ka-
    gaku 24:141-51.
 Walker, K. C., George, D. A., and Maitlen, J. C. 1965.  Residues  of
    DDT in fattu  tissues of big game animals in the  states of Idaho
    and Washington in 1962. USD A, ARS 33-105.
Welch, R. M., Levin, W., and Conney,  A.  H. 1969. Estrogenic action
    of DDT and its analogs. Toxicol. Appl. Pharmacol.  14:358-67.
White, W. C., and Sweeney, T.  R. 1945.  The metabolism of  2.2-bis-
    (p-chlorophenyl)l,l,l-trichloroethane (DDT).  I.   A metabolite
    from rabbit urine, di(p-chlorophenyl)  acetic acid;  its  isolation,
    identification, and synthesis. Public. Health Rept.  60:66-71.


        ONTAMINATION of the environment by pesticides has been a
subject cf mounting concern for over 20 years.  Within the past year
(1969) we have seen this concern, largely  focused on DDT, reach a
pitch where an aroused public is demanding action.  This is indicated
by the frequency and nature of coverage  in the nation's press, the
appointment  of committees  at the highest levels of government to
consider the problem, and the number of restrictive bills prepared for
presentation to various state legislatures and to the Congress. Ari-
zona, in January  1969, banned the use of  DDT  for agricultural and
commercial purposes for a 1-year trial period.  Michigan has restrict-
ed its employment to control of mice and bats and for emergency
public health purposes on approval of application.  Steps have been
taken in a number of other states to reduce or better control the use
of DDT.
    We in this country are not  alone in  our anxiety.  Sweden has
placed DDT under a 2-year ban. The Soviet Union is considering such
a ban.  Hungary has banned all organochlorine insecticides,  and
Britain is reportedly phasing them out.1  All of this comes at a time
when  world food  production is at an all-time high  and vector-borne
diseases of man and animals are more nearly arrested than ever be-
fore—a condition that must in part be credited to the effectiveness of
DDT and other pesticides.

     Pesticide Production and  Usage

     It is necessary to consider the  amount and nature of pesticide
manufacture and usage to gain perspective  about  the potential for

     H. PAGE NICHOLSON is Chief, Agricultural and Industrial Water Pol-
     lution Control Research Program, Southeast Water Lab., FWPCA,
     USDI, Athens, Georgia.
     1.  Chattanooga ("Tennessee) Times, June 11, 1969.



 pesticide involvement in water pollution.  Although the United States
 production of pesticides exceeded one billion pounds in 1967, not all
 was used domestically; about 40% was exported and an additional 12
 million pounds, primarily herbicides, were imported (Mahan  et al.,
 1968). It is not known how many of the 750 basic pesticidal chemi-
 cals listed by Gunther et al. (1968) were included in these production
 figures.  Some idea can be obtained, however, from knowledge that
 less than half this  number were registered for use and covered by
 legal tolerances or exemptions in the United States in 1966 (Westlake
 and Gunther, 1966). Only 14 insecticides, 8 fungicides, and 5 herbi-
 cides accounted for nearly 54%  of the tonnage manufactured in the
 United States in 1967 (Mahan et al., 1968).
     From these data it may be concluded, at least with respect to the
 quantities of pesticides used within  the United States, that the po-
 tential for widespread water pollution is  currently limited to a rela-
 tively few compounds. These  figures, however, do not preclude the
 possibility of local pollution problems associated with  the manufac-
 ture or processing of any pesticide, nor from accidental spills or care-
 less use.
     Environmental  Contamination and  Significance

     The acute effects of gross pesticide pollution are well known and
 depend upon the toxicity of the compound in question  and its con-
 centration in the environment.  Widespread  and chronic environ-
 mental pollution problems involve only those pesticides and degrada-
 tion products that are not only toxic but also possess the characteristic
 of extended  persistence sufficient to allow their escape from control
 after application, coupled with the  ability to be taken up  and con-
 centrated in  living organisms.  The latter has been called "biological
 magnification."  The pesticides  most frequently involved  are  the
 organochlorine  insecticides DDT, TDE, endrin, heptachlor,  aldrin,
 dieldrin, chlordane,  toxaphene,  Strobane, and BHC  or  its gamma
 isomer, lindane (Nichloson, 1969).  More recently, the organic mer-
 cury compounds have been implicated (Smart, 1968; Novick, 1969).
     Among  these compounds, DDT has been the most objectionable.
 The universal occurrence of traces of DDT is now common knowl-
 edge. It is used throughout much of the world  and its secondary dis-
 persion is  aided by  wind, water, and the movement of  animals  in
 which residues have accumulated.
     We are faced with mounting evidence that traces of DDT are not
 as innocuous as many have believed.  Transovarially conveyed DDT
 was shown  to  be responsible for significant losses of lake trout
 fry in a New York fish hatchery (Burdick et al., 1964), and investiga-
 tions are being  made to determine  if losses of coho  salmon fry  in
 Michigan hatcheries  are similarly caused.
     A drastic decline in populations of  fish-eating  raptorial  birds,
 such as the bald eagle and osprey, has long been associated circum-
 stant;ally with the advent and use of DDT.  Only recently has sup-
porting  evidence been produced to  suggest that DDT and its me-
tabolite,  DDE, can cause an imbalance in calcium  metabolism re-


 suiting in eggshell thinness and loss of eggs in the nest through break-
 age  (Hickey and  Anderson,  1968).  The hypothesis  that  sublethal
 amounts of persistent chlorinated hydrocarbon pesticides are involved
 is further strengthened by experimental studies  with dieldrin  and
 DDT, using captive sparrow hawks (Porter and Wiemeyer, 1969).
     Finally, we have experienced recently (March 1969) the seizure
 by the Food and Drug Administration of 28,000 pounds of coho salm-
 on, caught commercially from Michigan streams, because the  fish
 contained up to 19 ppm of DDT, an amount deemed to be excessively
 high (Congressional Record, 1969).  This seizure  was a severe  blow
 to commercial fishing and recreational interests of the  states adjacent
 to Lake Michigan  as this new fishery was proving to be an  economic
 bonanza (Henkin, 1969).
     It should be  pointed out that in  each  of these instances of in-
 secticide-related loss, the insecticide must first have entered  water.
     Concentrations  in  Water

     A synoptic survey for chlorinated  hydrocarbon insecticides in
waters of 56 of the nation's major drainage basins and 3 of the Great
Lakes was made by the U.S. Public Health Service  on whole-water
samples  collected during the period September 18-29, 1964 (Weaver
et al., 1965).  The samples from 96 sites in 41 states were analyzed
by  thin  layer chromatography and microcoulometric titrarion gas
chromatography. The wide distribution of dieldrin, endrin, and DDT,
with  its  metabolite DDE, is  significant (Table 12.1).  Concentration
values in water  all were less than one ppb.

TABLE 12.1.  Chlorinated  hydrocarbon  insecticides  and related compounds
            in major rivers of the United States.

Heptachlor epoxide

(No. states with
positive or
positive samples)



and Quantified
Range ppb*


Source:  Adapted from data by Weaver et al. (1965).
Note: Except Alaska and Hawaii.
* Minimum detectable  concentrations of dieldrin, endrin, DDT,  DDE, al-
drin, and heptachlor ranged from 0.002 to 0.010 ppb.  Comparable values
for TDE, heptachlor epoxide, and BHC were 0.075, 0.075, and 0.025 ppb,


 TABLE 12.2.  Pesticides in ten selected western streams,  1965-66.

                         No. Samples Positive             Range
    Compound         of 114 from 11 Stations      of Concentration
Heptachlor epoxide . . .
2 4-D 	

... 75 positive for one
or more pesticides
0.005-0 020
0.005-0 015
0 005-0 090
0.005-0 020
0.025-0 110
0.005-0 015
0.005-0 015
0.005-0 015

 Source:  Adapted from data by Brown and Nishioka (1967).

     Similarly, the U.S. Geological Survey, in October 1965, began the
 collection and analysis  of water  samples with associated suspended
 sediment from selected  streams west of the Mississippi River (Brown
 and Nishioka,  1967; Manigold  and  Schulze,  1969).  The  samples,
 taken  monthly,  were examined  by  electron  capture gas  chroma-
 tography for the  common chlorinated hydrocarbon insecticides  and
 the herbicides 2,4-D, silvex,  and  2,4,5-T.  Results  of the first year's
 work on  10  rivers are summarized in Table 12.2,  and those for the
 subsequent 2 years from 19 rivers are given in Table 12.3.
     All 12 pesticides or derivatives were recovered at one time or
 another.  Sixty-six percent of  114 water samples taken during the first
 year were positive for 1 or more pesticides. Forty-nine  percent of 333

 TABLE  12.3.  Pesticides in nineteen selected western streams, 1966-68.

                        No.  Samples  Positive             Range
   Compound          of 333 from 20 Stations      of Concentration
2 4-D 	
2 4 5-T 	
Silvex • 	
Heptachlor epoxide . . .

, . . . 82
, . . . 41
. . . 164 positive for one
or more pesticides
0.01-0 04
0.02-0 04

Source:  Adapted from data by Manigold and Schulze (1969).


samples were positive in the subsequent 2-year period.  Concentration
values all were less than 1 ppb.
     Lindane,  dieldrin, and heptachlor epoxide were recovered most
frequently during the period October 1965 through September 1966.
During the following 2 years DDT, DDE, and the herbicide 2,4-D were
most commonly found. No explanation was offered for this difference
in frequency of recovery.
     The  concentration of chlorinated  hydrocarbon insecticides  in
water alone, however, does not necessarily correctly reflect the availa-
bility of  these compounds to  living components of the hydrosphere.
Studies in the major agricultural river basins of California have in-
dicated that an average pesticide concentration of 0.10 ppb to 0.20
ppb in water may mean that bottom sediments contain 20 ppb to 500
ppb of the compounds (Bailey and Hannum, 1967).
     A pond near Denver was treated with 0.02 ppm of DDT, and the
insecticide was quantified in water, mud, vegetation, fish, and cray-
fish for 16 months (Bridges et al., 1963).  As the DDT residues in the
water decreased to none detectable at 4 weeks, residues in the mud in-
creased to 8.3 ppm at 24 hours and disappeared in 12 months; vege-
tation residues were 30.7 ppm within 30 minutes and declined to 0.6
ppm at 12 months; rainbow trout, black bullhead, and crayfish still
contained DDT and the metabolites TDE and DDE at 16 months.
     Quite clearly  DDT  is remarkably hydrophobic  and does not re-
main in  water in very large quantities. It  does tend  to concentrate
and  persist  in other compartments of the hydrosphere.  Such data
give  reason to pause and reconsider whenever the urge strikes to set
permissible  limits  for this and  similar compounds in water  alone
where the objective is to maintain a. suitable overall environment for
aquatic life.

     Pesticides may enter water in a variety of ways. These include
runoff from the land, industrial waste discharges,  carelessness and
accidents, and by direct application to control unwanted plant and
animal pests (Nicholson, 1969).  Other sources may be airborne resi-
dues, products of home use and garbage disposals (sewers), dumped
products containing residues higher than  tolerances, dead animals
and  animal excreta, and decaying plant tissues (Westlake  and Gun-
ther, 1966).  The significance of these additional  sources remains to
be more  fully documented, but their validity as sources is  not ques-
tioned. Contamination may be more or less continuous, generally at
very low levels (less than 1 ppb), or in brief episodes that may reach
concentrations sufficient to kill fish and other aquatic life.

     Runoff from  the land  is probably the most widespread single
source of low level surface water contamination by pesticides and
has been demonstrated repeatedly (Nicholson et al.,  1962; Hindin et


 al., 1964; Nicholson et al., 1964; Lauer et al., 1966; Nicholson et al.,
 1966; Bailey and Hannum, 1967). Runoff may be more  or less con-
 tinuous throughout the year at levels generally less than 1 ppb or may
 occur sporadically (Nicholson,  1969). Transport from land to water
 may occur while the pesticide is adsorbed on eroded particulate mat-
 ter, while in solution in runoff water, or by both means.  It has been
 shown that  sodium humate,  a common  soil constituent, solubil'.zes
 DDT in water (Wershaw,  1969).  This phenomenon would be ex-
 pected to facilitate the transport of DDT.
      Factors that control the runoff of pesticides are the nature of the
 pesticide and the extent to which it is used, edaphic considerations,
 climatic factors, topography,  and land usage and management prac-
 tices. Pesticides having short half-lives do not possess the runoff po-
 tential of persistent types.  High-humus-type soils  will yield less in-
 secticide than will sandy soils  (Lichtenstein, 1958).  Heavy rainfall
 immediately following application of chlorinated hydrocarbon-type in-
 secticides is a classic cause of runoff that sometimes causes fish kills
 (Young and  Nicholson,  1951).
     Industrial Waste

     Perhaps the second most significant source of pesticides in water
 is industry.  The types of industries involved include producers of
 basic pesticides, pesticide formulators,  cooperage firms that reclaim
 used pesticide drums, textile plants that moth-proof woolen yarns and
 fabrics with dieldrin, and paper manufacturing industries that use
 phenylmercury  acetate (PMA) as a fungicide.
     Releases from industrial sources may be continuous in manu-
 facturing or process effluents, or occasional in high concentration slug
 discharges following  in-plant mishaps or breakdowns.  In the latter
 case, biological catastrophies may result in receiving streams.
     As an example of a plant breakdown, an instance that occurred
 in Alabama may be cited (Alabama Water Improvement Commission,
 1961). A plant which manufactures parathion and methyl parathion
 normally treats  its wastes very effectively by  neutralization  with
 strong alkali followed by double activated sludge treatment through
 its own plant and that of a nearby city.  During a breakdown in  1961
 neither treatment plant could handle the load of toxic materials and
 60%  of  the  combined industrial effluent  containing parathion  and
 city  sewage was discharged untreated to a creek until corrective ac-
 tion  could be taken. Fish, turtles, and snakes died along 28 miles of
 the stream whose average discharge at the time was 211 million gal-
 lons  a day at a  velocity of 3A mile per  hour.  The creek entered the
 Coosa  River which then had  an average  discharge about  28 times
 greater than that of the creek.  Even with that dilution, traces of para-
 thion residues were recovered  90 miles down the Coosa  and some
lesser fish kills  occurred in it.  Unfortunately,  water samples  were
not taken for analysis until the third day after fish were first observed
dying.  At that time a maximum 0.21 ppm  of parathion was found at
a point on the creek 22 miles from the  city.  On that same day, 667


ppm of parathion was reported in thickened sludge at the city sewage
treatment plant.
     A second example involved a  manufacturer of DDT who pro-
vided minimal waste treatment.  The DDT content of water in a ditch
receiving these wastes was 0.1  ppm  to  7.8  ppm.  Bottom  samples
analyzed from 0.6% to 2%  DDT along the half-mile length of this
ditch after it had been carrying away wastes from  this industr-al
plant for IVfe years.2
     Barthel et al. (1969), in their study  of pesticide residues  in the
sediments of the lower Mississippi River  and its tributaries,  found 5
pesticide-formulating  companies that  dumped  waste materials  in
city  sewers, in  channels and sloughs near  their plants, and on city
and privately owned dumps where they could be washed away by rain-
fall.  Residues found included dieldrin, aldrin, endrin, isodrin, chlor-
dane, lindane, and DDT analogs and metabolities.  Many of  the resi-
dues in river bottom sediments were in concentrations less than 0.05
ppm, but some ranged in the thousands of ppm in the vicinity of in-
dustrial plants.
     Accidents end Carelessness

     Although strenuous efforts have been made within agricultural
and related industries to minimize accidents and  carelessness with
pesticides,  some instances of water pollution from these causes still
are reported. An instance of carelessness having potentially serious
human health implications occurred in Florida in 1964 (Florida State
Board of Health, 1964).
     A rancher instructed his hired hand to dispose of approximately
50 four-pound bags of over-age 15% parathion  dust. This was done
without the rancher's knowledge, by dumping them from a highway
bridge into the Peace River 1 mile upstream from the municipal water
intake of Arcadia, a town of about 6,000 people.  The  act was dis-
covered when boys fishing near the bridge hooked a bag and reported
     The town fortunately had an auxiliary well and immediately re-
verted to it.  The citizens were instructed not to use the water, and
flushing of the mains was begun. Subsequent analysis of water sam-
ples showed that parathion  concentration in the distribution system
was generally less than  1 ppb.  However, a series  of samples taken
from a tap at the local bus station ranged from 10 ppb to 380 ppb.
     Investigation  revealed  that  the  bags  of  parathion  had  been
dumped in the river about 10 days before their discovery. They were
polyethylene lined and resisted rapid disintegration.  All but 8  to 12
bags were  eventually recovered. Those unrecovered  bags that disin-
tegrated apparently did so over a period of several weeks. This may
have been the  reason that  residue levels sufficiently high  to be a
threat  to human health or the fish in the river  did not occur.  Para-
thion residue occurred in the river water for about 2 weeks after dis-
     2.  Charles Kaplan, FWPCA, Southeast Region, Atlanta, Georgia, per-
     sonal communication.


 covery at concentrations generally less than 1 ppb.
     Accidents  also have caused real or potential water  pollution
 problems.  In March 1965, during the night, 2,500 to 3,000 pounds of
 5% chlordane wettable powder spilled from a truck passing through
 Orlando, Florida.  After  recovering what could be salvaged,  about
 1,300 to 1,700 pounds  was hosed into the street's storm drainage
 system from which it passed into a dry creek bed not far from  one
 of the city's lakes. When the potential for damage to  the lake was
 realized, the contaminated water and soil were removed for safe  dis-
 posal elsewhere.
     Other  Sources

     The chemical control of aquatic weeds, rough fish, and aquatic
 insect pests often results in some pesticide residue in water. These
 activities are generally managed by professionals so that undesirable
 consequences are minimized. Toxaphene, however, that was first used
 for control  of rough fish in lakes  in  the early 1950s has sometimes
 caused trouble.  Although toxaphene-treated lakes may generally be
 restocked within 6 months to a year later, occasionally a lake may re-
 main toxic to restocked fish for 5 years (Kallman et al., 1962; Terriere
 et al., 1966).  Such was  the case at Miller Lake in Oregon that was
 treated in 1958 at an estimated rate  of 40 ppb. The initial residues
 declined sharply to less  than 2 ppb and remained near the level for
 approximately 5 years (Terriere et al., 1966).
     Airborne dust containing pesticides may also contribute to pesti-
 cide levels in water either by  direct  deposition or by deposition  on
 land with subsequent runoff.  Dust deposited on Cincinnati, Ohio, in
 January 1965, originating from the southern high plains of Texas and
 adjacent states, was shown to contain 0.6 ppm DDT, 0.5 ppm chlor-
 dane, 0.2 ppm DDE, 0.2 ppm Ronnel, 0.04 ppm heptachlor epoxide,
 0.04 ppm 2,4,5-T, and  0.003  ppm dieldrin (Cohen  and Pinkerton,
     Local drift of dusts and sprays from areas of pesticide applica-
 tion is well known and can also be  a source of water contamination.
     Urban  storm  water has also been suggested as a  carrier of pesti-
 cides (Weibel et al., 1966). The significance of urban  sites and activi-
 ties as sources of  water pollution by  pesticides is now being investi-
 gated at Michigan State University.

     Can  anything  practical be done  to control water pollution by
pesticides?  The answer is most definitely yes.
     Point sources can be controlled most easily. These are industrial
sources where waste  effluents enter watercourses through single or
adjacent outfalls. A variety of effective waste treatment systems are
now employed. Research  and demonstration grant funds are availa-
ble through the Federal Water Pollution Control Administration for


studies leading to the development of improved waste treatment mea-
sures.  Research on chemical degradation of pesticides is expected to
result  in knowledge that can be engineered into effective new treat-
ment technology.
     Since the more persistent chlorinated hydrocarbon insecticides
are recognized as being more troublesome than less refractory pesti-
cides,  their usage can be minimized or reserved  for those purposes
where substitutes will not do.
     Advanced concepts of pest control  are being developed  that
range  from male sterilization techniques to use of micro quantities of
pesticides that are  effective, yet not as wasteful  of toxicants as  are
present techniques.  The acceptance of levels of pest  control some-
what less than eradication also is being emphasized.  The latter in-
volves reduction  of pest population levels  only to that point  where
economic loss will not result.
     Scientists at the Southeast Water  Laboratory are working to de-
velop a means to predict runoff pollution and to  prevent it from oc-
curring (Nicholson, 1969).  The Universal Soil Loss  Equation,  de-
veloped by soil conservationists to guide conservation farm planning
throughout the United States, is being considered as the basis for a
new formula  for predicting pesticide  loss from  the  soil.  Knowing
what to anticipate, control  would be  accomplished by using recom-
mended chemicals and practices.

Alabama Water Improvement Commission. 1961.  A report on fish
     kills occurring on Choccolocco Creek and the Coosa River during
     May 1961.
Bailey, T. E., and Hannum, J. R.  1967. Distribution of pesticides in
     California.  /. Sanit. Eng. Div.  Am. Soc. Civil Engrs.  93 (SA5):
Barthel, W. F.,  Hawthorne,  J. C.,  Ford,  J. H.,  Bolton, G. C., Mc-
     Dowell,  L.  L.,  Grissinger,  E.  H., and  Parsons,  D. A.   1969.
     Pesticides in water.  Pesticide Monitoring J.  3:8-66.
Bridges. W. R., Kallman, B. J., and Andrews, A. K. 1963. Persistence
     of DDT  and its metabolites in a farm pond. Trans. Am. Fish-
     eries Soc. 92:421-27.
Brown, E.,  and Nishioka, Y. A. 1967.  Pesticides in selected western
     streams—a contribution to the  national  program.  Pesticide
     Monitoring  ].  1:38-46.
Burdick, G. E., Harris, E. J., Dean, H. J., Walker, J. M., Skea, J., and
     Colby, D.  1964. The accumulation  of DDT in lake trout and
     the  effect  on  reproduction.  Trans.  Am.  Fisheries Soc.  93:
Cohen, J. M.,  and  Pinkerton, C. 1966.  Widespread translocation of
     pesticides by air transport and rain-out.  In Organic pesticides
     in the environment, Ad-van. Chem. Ser. GO.  Wash., D.C.:  Am.
     Chem. Soc.
Congressional Record—Senate (S9417) Aug. 8, 1969.
Florida State Board of Health.  1964. Report of Peace River para-
     thion  incident  Dec. 23, 1964.  Jacksonville:  Bur.  of Sanit. Eng.


 Gunther, F. A., Westlake, W. E., and Jaglan, P. S.  1968. Reported
      solubilities of 738 pesticide chemicals in water.  In Residue Re-
      views, ed. F. A.  Gunther, pp.  1-148.  New York:  Springer-
 Henkin, H. 1969.  Problems in PPM. Environment 11:25, 32-33,37.
 Hickey, J. J., and Anderson, D. W.  1968.  Chlorinated hydrocarbons
      and eggshell changes in raptorial and fish-eating birds.  Science
 Hindin, E., May, D. S.,  and  Dustan, G. H.  1964.  Collection and
      analysis of synthetic organic pesticides from surface and ground
      water. In Residue  Revieivs,  ed. F.  A.  Gunther,  pp. 130-56.
      New York: Springer-Verlag.
 Kallman, B. J., Cope, O.  B., and Navarre, R. J.  1962.   Distribution
      and detoxification of toxaphene in  Clayton  Lake, New Mexico.
      Trans. Am. Fisheries Soc. 91:14-22.
 Lauer,  G. J., Nicholson, H. P., Cox, W.  S., and Teasley, J. I.  1966.
      Pesticide contamination of surface  waters by sugar cane farm-
      ing in Louisiana.  Trans. Am. Fisheries Soc. 95:310-16.
 Lichtenstein, E. P.  1958. Movement of insecticides in soils under
      leaching conditions.  /. Econ.  Entomol. 51:380-83.
 Mahan, J. N.,  Fowler, D. L., and  Shepard, H. H. 1968.  The Pesti-
      cide Review 1968. Wash., B.C.:  USDA, Agr. Stabilization and
      Conserv. Serv.
 Manigold,  D.  B.,  and Schulze,  J.  A.  1969.  Pesticides in selected
     western streams—a progress  report.  Pesticide Monitoring  J.
 Nicholson,  H.  P.   1967.   Pesticide  pollution  control.  Science
 	.  1969.  Occurrence and significance of pesticide residues in
     water.  J. Wash. Acad. Sci. 59:77-85.
 Nicholson, H. P., Webb, H. J., Lauer, G. J., O'Brien, R.  E., Grzenda,
     A.  R.,  and Shanklin, D.  W.   1962.  Insecticide contamination
     in  a farm  pond. I.  Origin  and duration.  Trans. Am. Fisheries
     Soc. 91:213-17.
 Nicholson, H.  P.,  Grzenda, A.  R., Lauer,  G. J., Cox, W. S.,  and
     Teasley, J. I.   1964.  Water pollution by insecticides in an agri-
     cultural river basin.  I. Occurrence of insecticides  in river and
     treated municipal water.  Limnol. Oceanog.  9:310—17.
 Nicholson, H. P.,  Grzenda, A. R.,  and Teasley,  J. I.  1966.  Water
     pollution by insecticides:  a six and  one-half year study of a
     watershed.   Proc.  Symp. Agr. Waste Waters,  Rept. 10, pp.
     132-41. Davis:  Univ. of Calif.
 Novick, S.   1969.  A  new pollution problem.  Environment 11:3—9.
 Porter,  R.  D.,  and Wiemeyer,  S.  N.  1969.  Dieldrin  and  DDT:
     effects  on  sparrow hawk eggshells  and  reproduction. Science
 Smart, N. A. 1968. Use and residues of mercury  compounds in agri-
     culture.  In Residue Reviews, ed. F. A. Gunther, pp.  1—36.  New
     York:  Springer-Verlag.
Terriere, L.  C., Kiigemagi, U., Gerlach, A.  R., and Borovicka, R. L.
     1966.  The persistence of toxaphene in lake water  and  its up-
     take by aquatic plants  and  animals.  /.  Agr.  Food  Chem.
Weaver, L., Gunnerson, C. G.,  Breidenbach, A. W., and Lichtenberg,
     J. J. 1965.  Chlorinated hydrocarbon  pesticides in  major U.S.
     river basins.  Public Health Rept.  80:481-93.


Weibel, S. R., Weidner, R. B.,  Christiansen, A. G.,  and Anderson,
     R. J. 1966.  Characterization,  treatment, and disposal of urban
     stormwater.  Third Intern. Conf. Water Pollution Res., Munich,
     Germany.  Section I, Paper 15, pp. 1-15. Wash., D.C.:  Water
     Pollution Control Federation.
Wershaw, R. L., Burcar, P. J., and Goldberg, M. C.  1969. Interac-
     tion of pesticides  with organic material.  Environ. Sci. Teclmol.
Westlake, W. E., and  Gunther,  F.  A.  1966.  Occurrence and  mode
     of introduction of pesticides in the environment.  In Organic
     pesticides  in  the environment,  Advan.  Chem. Ser.  60,  pp.
     110-21. Wash., B.C.: Am. Chem. Soc.
Young, L. A., and Nicholson, H. P.  1951. Stream pollution resulting
     from the use of organic insecticides.  Progressive Fish-Culturist



        I ERBICIDES are essential and widely used tools in our modern
agriculture. In 1964  approximately 120 million  acres  of cultivated
fields, pastures, grazing lands, and forested areas were treated with
200 million pounds of herbicides for weed control (U.S. Department
of Agriculture Census, 1964).  In 1967 the total  sales  of herbicides
had increased to 348  million pounds.  A small but relatively signifi-
cant proportion of those herbicides were  used to control weeds in
irrigation  and drainage canals, on ditchbanks, in farm ponds,  and
in irrigation reservoirs.
    Thus far, monitoring studies have shown few significant herbi-
cide residues in our streams, ponds, and lakes resulting from runoff
from treated fields, rangelands, and forests.  There is some  concern
about the effects on water quality of herbicides appied directly  into
or over the water  or  on adjacent banks  from which  drift, overlap
spray, or runoff may get into surface water supplies.
    This  chapter reports the extent of herbicide use for control of
aquatic and bank weeds, the levels of residues found in water after
such applications, the  rate of dissipation  of such residues,  and
whether and to what extent herbicides in irrigation water are found
in irrigated crops used  for food or  feed.   Only limited information
is presented on herbicide residues in water from other sources.

     No statistics are available on the total amount of herbicides used
annually in the United States for control of aquatic and bank weeds.
However,  several examples  of  the  extent of  aquatic  areas where
weeds are serious problems  and the amount of herbicides used in

     F.  L. TIMMONS is Research Agronomist,  Crops Research Division,
     ARS, USDA, Laramie, Wyo. P. A. FRANK is Plant Physiologist, Crops
     Research Division, ARS,  USDA, Denver. R. J.  DEMINT is Research
     Chemist, Crops Research  Division, ARS, USDA, Denver.


                                  CHAPTER 13 / HERBICIDE RESIDUES / 195

certain  areas provide reliable indications of the total  amounts  used
in and adjacent to water.
     About 150 species of aquatic and  semiaquatic  marginal plants
create weed problems in one or more aquatic situations in the United
States (Timmons, 1967).  According to  the  latest available  statistics
(U.S. Department of  Agriculture  Census,  1959, 1964), there are
more than 2 million ponds and reservoirs, 189,000 miles of  drainage
ditches, and  173,000 miles of irrigation canals.  Most of the ponds
and  drainage  ditches are in the north-central and  southern states.
Three-fourths  of the irrigation canals  and most of  the reservoirs
which  supply  irrigation  water  are in  the  western states.  The
numerous reservoirs in the southern and north-central states are used
primarily for recreation and municipal purposes. All of these aquatic
areas are infested or susceptible to infestation by aquatic and bank
     In  1957 a careful survey was conducted by the  Agricultural Re-
search Service  and the Bureau of Reclamation (Timmons,  1960) to
determine the  extent of weed infestation,  annual losses caused by
weeds, and the cost of weed  control on irrigation systems of the 17
western states.  The survey revealed that 63% of the  144,000 miles of
canals were infested with aquatic weeds.  More  than 759c  of the
530,000 acres of ditchbanks were infested with 1 or  more of 4 kinds
of bank weeds. In that year, 54% of the weed-infested canals and
80%  of the weed-infested ditchbanks were  treated for weed control,
mostly with herbicides.
     A questionnaire survey made in 1961  (Timmons, 1963) among
agencies and aquatic weed specialists revealed that aquatic and mar-
ginal weeds  were  serious problems in most  ponds  and  drainage
ditches  in the  north-central,  southern,  and western states.   The ex-
tensive  annual losses  from lack of drainage and water utilization
caused by those weeds were  reported in Agriculture Handbook 291,
Losses in Agriculture, 1965.
     An extensive weed control program has been  continued since
1957 on western irrigation systems.  The herbicides  used  most ex-
tensively in the control programs are xylene; (2,4-dichlorophenoxy)
acetic acid  (2,4-D);  (2,4,5-trichlorophenoxy)  acetic acid (2,4,5-T);
copper sulfate; 2,2-dichloropropionic  acid (dalapon); and 3-amino-s-
triazole (amitrole). Aromatic weed oils are used  extensively in the
southwestern states. The  amounts of herbicides ordered by irrigation
districts in  Oregon, Washington, and  Idaho for use on irrigation
systems during 1969 were xylene, 800,000  gal; acrolein, 22.000 gal;
copper  sulfate, 216,000  lb;   2,4-D, 187,000 lb; 2,4,5-T,  9,200 lb;
dalapon, 3.500 lb; and amitrole + ammonium thiocyanate (amitrole-
T), 5,000 lb.1   This is for only 3 of the 17 western states.   Most of
the other western states do not use herbicides as extensively as do
the 3 northwestern states.
     General information indicates that weed problems  in  drainage
ditches  of eastern states  are  as critical and probably more so than
those in western irrigation systems.  However, the use of herbicides
does not seem to have been as extensive for control of  weeds in those
drainage ditches except possibly in Florida  and Louisiana.  Mechani-


 cal methods have not proved to be adequate substitutes for herbicides
 in drainage ditches and ponds in those states.
      Irrigation and drainage of agricultural land are important fac-
 tors  in  the conservation and  use of water resources in the south-
 eastern  states.  The aquatic weed problems in lakes, streams, and
 water-control canals in that region are much  more extensive and
 serious  than in the north-central states.  A survey was conducted in
 1963 in 8 Gulf and South Atlantic Coast states  (U.S. Department of
 the Army, 1965).  The survey  showed total infestations  of  162,000
 acres of water hyacinth (Eichhornia  crassipes  [Mart.] Solms). 99,000
 acres of alligator weed (Alternanthera philoxeroides [Mart.] Griseb.),
 and 207,000 acres of submersed weeds.  The survey did not include
 farm ponds  and  tidal  marsh  areas, most  of  which  are  heavily in-
 fested by aquatic and marginal weeds in those states.
      Herbicides, chiefly 2,4-D, have been  used extensively  since about
 1950 for  the control  of water hyacinth and  certain other  floating
 and emersed weeds in Florida  and Louisiana.  During the extensive
 unrestricted  use of 2,4-D prior to 1967, no serious problems of injury
 to  fish, livestock,  or man from use of treated water were apparent.
 During  the 4 years 1959-62, approximately 188,000 acres of water
 hyacinth and  alligator weed were sprayed with 2,4-D in the  U.S.
 Army Corps  of Engineers Expanded  Aquatic Plant  Control Program.
 That did not account for the herbicide usage by other agencies and
 private individuals during those years.
     Aquatic herbicides such as  2,4-D; 6,7-dihydrodipyrido[l,2-a:2',
 l'-c]pyrazinediium salts (diquat); 7-oxabicyclo[2.2.1]heptane-2,3-di-
 carboxylic acid (endothall);  dalapon; and  2-(2.4,5-trichlorophenoxy)
 propionic acid (silvex) are used to a considerable extent in the south-
 east,  especially in Florida.  The highly successful aquatic and mar-
 ginal weed program of the Central and Southern Flood Control Dis-
 trict  is  an excellent example  of what  can be   accomplished by
 extensive  and  careful use of all available registered herbicides for
 control of aquatic and marginal weeds.
     At present 6 herbicides are registered  by  the Pesticides  Regula-
 tion Division of  the Agricultural Research Service  for  control  of
 algae, 4  for control of floating weeds, 6 for control of emersed weeds,
 and 12 for control of submersed weeds.  In addition, 17 herbicides
 are registered for control of  ditchbank weeds.  That is a  total of 35
 different herbicides registered  for the control  of  aquatic or  bank

     The principal means by which herbicides enter water  are  (1)
from surface-runoff water during irrigation or rainfall;  (2) by appli-
cation of herbicides to soil or water for control of submersed weeds in
canals, ponds, or  lakes; (3) by herbicide treatment  of  floating  and
emersed  weeds,  and  (4) from treatment of banks of streams  and
canals for control of bank and marginal weeds.

     1.  W. D. Boyle, Bureau of Reclamation, Boise, Idaho, 1969, personal

                                 CHAPTER 13 / HERBICIDE RESIDUES / 197

     Residues  in Surface Runoff

     During a  3-year program of monitoring agricultural pesticide
residues, herbicides were found only rarely,  and usually in concen-
trations less than 10 ppb  (U.S. Department of Agriculture,  1969).
Monitoring  studies in  an area of  irrigated agriculture showed that
before use, irrigation water contained very small quantities (<1 ppb)
of pesticides of any kind. Of a number of herbicides used in the area
monitored, no detectable residues  of these were found in waste irri-
gation water.
     In other monitoring studies (Marston et  al., 1968) small quanti-
ties  of amitrole were detected in runoff water for  5  days  following
aerial spraying of a 100-acre watershed for control of salmonberry.
A maximum concentration of 155 ppb  amitrole was  found 30 min-
utes after spraying began, but was reduced to 26 ppb after 2  hours.
On the other hand, amitrole  was found in runoff water for only 35
hours from  a  similar but larger watershed treatment (Norris et  al.,
1967). When  2.1 acres of  a 46.5-acre rangeland watershed were
treated with 9.3 Ib per acre of 4-amino-3,5,6-trichloropicolinic acid
(picloram), the runoff water from this watershed contained picloram
in concentrations of 0.37 to 0.046 ppm for 11 months (Davis et  al.,
1968). No picloram was found after this period. Concentrations of
1.5 to 2.0 ppm of 2,4-D were detected in runoff water for a period of
7 days following treatment of 150  acres of forest with 40 Ib per acre
of the nonyl ester of 2,4-D (Aldhous, 1967).  In a  subsequent sam-
pling of  the runoff water  28 days  after treatment,  the residue of
2,4-D was below the detectable level of 0.005 ppm.
     Experimental data showing the extent of herbicidal residues in
runoff water are limited.  Where residues were shown to  occur, in
most cases the total volume of water affected was not large.
     Residues  in  Water from  Control of Submersed Weeds

     Recommendations  for control  of  submersed weeds  usually
specify herbicide-usage  levels in terms of ppm of the herbicide  in
water. Therefore, the initial residue level most often represents the
recommended or predetermined concentration  of herbicide found to
be effective for  control  of the weed  species present.  A number  of
herbicides  and recommended application  rates for control  of sub-
mersed weeds are given  in Table 13.1.
     Submersed  weed control in waterways such as irrigation canals
is accomplished primarily by the use of acrolein,  aromatic solvents,
and  copper sulfate.  Diquat, endothall,  and the ester  of 2,4-D are
used less frequently.  Copper  sulfate is  commonly used for control
of algae; however, low concentrations applied over extended periods
have been reported recently to provide good control of vascular weeds
(Bartley, 1969).  Where very little water movement occurs in water-
ways, good weed control is obtained with diquat and the amine salts
of endothall.
     When maximum possible herbicide residues were found in water
from the applications recommended in Table 13.1, all of the applied


 TABLE 13.1.  Herbicides and application rates recommended for control  of
             submersed aquatic weeds.
Aromatic solvents ....
Copper sulfate . . .
2,4-D ester 	

Disodium salt
Amine salt
Potassium salt

Application Rates*
0.1-0.6 ppmt
4-7 ppm|
600-740 ppm§
0 1-2.0 ppm
10-15 Ib/a
0.9-1.4 ppm [|
0.25-1.5 ppm
0.5—4 ppm
0.05-2.5 ppm
15-20 Ib/a
1.4-1.8 ppmll
1.5-2 ppm
20-40 Ib/a
1.8-3.6 ppm|[
 * From USDA (1969) Suggested Guide  for Weed Control.  Agr. Handbook
 332. Application rates are in terms of acid equivalent or active ingredient.
 t For extended application time in flowing water.
 $ For treatment of weeds in quiescent water.
 § Emulsifier added at concentrations of 1.5 to 2.0%.
 || Ppm  concentration arbitrarily expressed in terms of 4 ft of water.

 herbicide  recovered  in water usually dissipated  rapidly.  Volatile
 herbicides such as  acrolein and aromatic  solvents (mostly xylene)
 are lost from water at relatively rapid rates.  Diquat concentrations
 are rapidly reduced by weed growth,  organic matter, and sediment
 (Coats et al., 1966).  Granular formulations may prevent occurrence
 of high concentrations of certain herbicides  in  water by confining
 portions of the herbicides  at the hydrosoil surface. Granular formu-
 lations of  2,6-dichlorobenzonitrile  (dichlobenil)  and  the  ester of
 2,4-D  are  notable in this  respect. Following treatment of 2 ponds
 with 0.58  and 0.40 ppm of granular dichlobenil, only 0.32 and 0.23
 ppm, respectively, were recovered (Frank and Comes, 1967). Like-
 wise, in a pond treated with 1.33 ppm of granular  butoxyethanol
 ester of 2,4-D,  the  maximum residue level of 2,4-D observed  was
 0.067 ppm.  At the same time, relatively high concentrations of both
 herbicides were found in the  upper  1 inch of hydrosoil.  On the
 other hand (2,3,6-trichlorophenyl) acetic acid (fenac) was rapidly lost
 from granules and nearly  all of the  herbicide applied was found in
 water above the granules which  remained at the  bottom of the pond
 or lake. During 1966 the Tennessee Valley Authority used large-scale
 applications of granular butoxyethanol ester of 2,4-D  at rates of 40
 to 100  Ib per acre for control of Eurasian watermilfoil (Myriophyllum
 spicatum  L.).  The highest concentration of 2,4-D recorded  at  any
of the water-treatment plants where water was monitored was 2  ppb
(Smith and Isom, 1967).

                                 CHAPTER 13  HERBICIDE RESIDUES /  199

     Residues in  Water from Control of Floating Weeds

     Herbicide residues resulting from treatment for control of float-
ing weeds are dependent not only on the application rate and water
depth  but also on  the type  of floating weeds and  the  amount  of
exposed water surface. Very few residue data on these  applications
are available.
     In one series of experiments, pools  10 feet in diameter contain-
ing growths of alligator weed  were sprayed with propylene  glycol
butyl ether  (PGBE) ester of silvex at 8 Ib per acre TCochrane et al..
1967). Highest possible concentrations of silvex residues would have
ranged from 2.70 to 3.04  ppm if all of the herbicide applied was
found  in  the water.  However, the greatest recovery of  silvex  in
water  at  any time was  approximately  1.6  ppm. In  this  study  no
estimates  of uncovered water surface were made. In similar  studies
involving  applications of the dimethyl  amine salt of  2.4-D and the
PGBE  esters of 2.4-D  and  silvex applied at  4 Ib per  acre on water
hyacinth  or alligator  weed, almost all  of  the maximum residue
levels  were  between 1 ppm and  650 ppb (Averitt.  1967).  In  both
of the above studies,  the highest  concentrations of herbicides  did
not appear in the water  until approximately 1 to 2 weeks  after the
treatments.  The authors concluded that the herbicides were  absorbed
by the plants and later  released  into the water through roots and
other submersed plant tissues.
     Residues  from Ditchbank Weed Control

     Spreading weed infestations have caused irrigation system man-
agers and maintenance workers to become more conscious of weed
control  on  banks of waterways.  Where  periodic treatment  with
2.4-D was once  considered  adequate  for  ditchbank maintenance.
extensive and varied weed control programs involving  other herbi-
cides or  mixtures are now common.  Among the most serious ditch-
bank weeds are several species such as sedges—for example. Carex
aquatilis Wahl and  reed canary grass (Phalaris  arundinacea  L.")—
which grow at the water margin. The proximity of weeds to water
almost invariably results in some herbicide entering the water during
herbicide application.   Principal  factors  affecting the amount  of
herbicide found in the water are treatment rate, water volume, na-
ture of the weed growth, and spray  overlap  at the water's edge.
     A number of ditchbanks  were  sprayed with various  herbicides
and the  water sampled and analyzed to determine the  quantities
of residues  present  (Trank  and Demint.  1967. 1968).  Herbicides.
treatment rates, and water volumes,  along  with the highest concen-
trations of herbicides found in the water of a  number of irrigation
waterways, are shown in Table 13.2.  With  one exception, all treat-
ments were made on 1 bank, with a vehicle-mounted boom traveling
in an upstream direction.  Both banks of the Boulder Feeder Canal
were  treated prior to the  entry of water.  The 98 ppb of amitrole


 TABLE 13.2.  Highest concentrations of residues found in  irrigation  water
             following ditchbank treatment with  several herbicides.
 Herbicide and Irrigation      Treatment    Volume of
       Waterway              Rate      Water Flow
            of Residue
Boulder Feeder Canal* . .
Farmer's ditch 	
Manard lateral 	
Yolo lateral 	
Five-mile lateral 	
Lateral no. 4 	
Manard lateral 	
Yolo lateral 	
Lateral no. 4 	
Manard lateral 	
Yolo lateral 	
Lateral no. 4 	
Manard lateral 	
Yolo lateral 	


, 6.7
, 10.5










Source: Unpublished data from P. A. Frank and  R. J. Demint, Annual
Report of Weed Investigations. USDA, ARS, Denver,  Colo.
* Both banks treated for distance of 0.7 mile.
t N-oleyl 1,3-propylenediamine salt.

represent the  herbicide picked up by the initial water filling the canal
and were  of  very short duration. Minimum and  average residue
values for all treatments were considerably less than the maximum
levels shown in the table.  It will be shown later that residues in the
concentrations listed in Table 13.2 would be most unlikely to injure
crops or produce significant residues in crop plants.

     Dissipation is an extremely important factor in  the use of herbi-
cides for control  of aquatic and bank weeds.  Most of the herbicides
registered for use in  aquatic  situations have water-use  restrictions
which require at least  partial  dissipation of the herbicide before
normal water use is  resumed.  The pathways leading to dissipation
are almost as varied as the chemicals  themselves.  Volatilization  is
the most important factor in the dissipation of aromatic solvents and
acrolein. Sorption processes predominate in the disappearance from
water of herbicides such as diquat, paraquat, and possibly endothall.
Biological and chemical degradation account for  much of the loss  of
2,4-D, silvex, dichlobenil, and other herb cides.
     The dissipation  of  herbicides in water has been studied  most
extensively in small ponds, pools, and reservoirs.  Data from some  of

                                  CHAPTER 13 / HERBICIDE RESIDUES / 201
TABLE 13.3.   Residue dissipation in  ponded water  following application of
                                         Concentration Detected
                             (ppm)    (ppm)  (days)
                               Liquid  applications
Amitrole*  	   1.0      1.34      1.0
Fenac*  	   4.0      5.2       1.0
Diquat*  	   2.5      3.27      2.0
Paraquat*  	   2.1      1.05      1.0
2,4-D methylamine saltt  ....   1.5      0.139     1.0
Silvex, PGBE estert§  	   2.9      1.6       7.0
Diquatjl  	   0.62     0.49      1.0
Paraquatil  	   1.14     0.55      1.0
Endothall||  	   1.0      0.18      2.0
Copper^  '	   0.50     0.42      0.1
Endothall**   	   1.2      0.79      4.0

                              Granular applications
                          (ppm)    (days)
2,4-D butoxyethanol ester|| . . .
* Grzenda, Nicholson, and Cox (1966).
+ Averitt (1967).
i Cochrane  et al. (1967).
•? Average of three treatments.
j Frank and Comes (1937).
1 Toth and  Riemer (1968).
** Yeo (1969).

the more typical studies were compiled and are shown in Table 13.3.
Some  of the most effective aquatic herbicides, such as  dichlobenil,
fenac, and silvex, were found to be among the more persistent com-
pounds. The excellent and often complete control of weeds by these
herbicides  may be attributed in part to their persistence.  Diquat,
paraquat, 2,4-D, and  endothall disappeared from ponded water at
rapid to moderate rates.  While rapid dissipation  from water is desir-
able from  the standpoint of residues, it  may also result  in the total
ineffectiveness  of  diquat and paraquat  in waters containing sus-
pended sediment  or  organic matter  (Coats et al.,  1966).  In some
cases dissipation of the herbicides from  water was found to be ac-
companied by accumulation of high concentrations  of the herbicides
in the hydrosoil (Frank and Comes, 1967).
     Dissipation of herbicide residues in  the flowing water of canals,
ditches, and streams has been studied less extensively than in ponds
and  very few data are published.  Most  of  the studies  reported here
were carried out recently by personnel of the Agricultural Research
Service and cooperators.


      While aromatic solvents have been used many years for control
 of submerged weeds in irrigation canals, it was not until 1967 that
 the  dissipation  of this herbicide was studied in some detail (Frank
 and  Demint, 1967). Two canals carrying 11 and 13 cubic feet per
 second (cfs) of water were treated with 575 and 550 ppm emulsified
 xylene, respectively. Loss of xylene, largely by way of volatilization,
 was  rapid.  After traveling 9 miles, the concentration  of xylene in
 the  canal  treated at the  rate of 550 ppm was  reduced to  17 ppm.
 The  concentration of xylene in the second  canal was  reduced to
 6 ppm after a downstream flow of 8 miles.
      Acrolein  is another highly  volatile  herbicide and  is used for
 control of submersed weeds in large irrigation canals. The loss rate
 of acrolein from concentrations of 0.6 and 0.7 ppm was  determined
 for  a canal carrying  132 to 135 cfs  of  water (Battelle-Northwest
 Laboratories, 1966, 1968). In 2 tests, the loss of acrolein  was shown
 to be temperature dependent. In water of 64° F the original  concen-
 tration of 0.7 ppm of  acrolein was reduced by 98%  while the water
 traveled a distance of 19 miles.  At the lower and less typical  tem-
 perature of 48° F, the loss  was only 62%  at  a  distance 27 miles
 downstream from the  point  of application. The dissipation  data of
 both aromatic solvent  and the acrolein showed  a linear relationship
 between the log of the herbicide concentration and distance of water
 flow downstream.
     Copper sulfate  is frequently used  to control algae in irrigation
 canals. The commonly used  slug treatment of 1 Ib of copper sulfate
 pentahydrate per cfs of water flow, when applied to a 411-cfs canal
 in Washington,  gave concentrations of 1.6, 0.36, 0.23, and 0.04  ppm
 at 0.5, 6, 12, and 23 miles downstream, respectively (Nelson et al.,
 1969).  A 3-year study was made to determine  the efficacy of daily
 application of copper sulfate for control of submersed weeds in irri-
 gation canals  (Bartley, 1969).  Five pounds  of  copper  sulfate were
 applied per hour to a flow of 26 cfs of water.  An average maximum
 concentration  of 0.21  ppm copper ion was  found 0.25 mile below
 the treatment site.  The copper ion concentration was reduced 86%
 to 0.03 ppm 9 miles downstream.
     In one study a  single bank of each of 2 irrigation laterals was
 sprayed with amitrole  in an upstream direction for a distance of 0.5
 mile. Treatment rates  were 3 and 4 Ib per acre.  Overlap of the spray
 pattern at the  water's edge  was estimated to vary from 12 to 24
 inches.  Water samples taken at varying distances downstream from
 the area treated with 4 Ib per acre of amitrole showed a reduction in
 residue levels from 31 to 24 ppb  over  a  4.5-mile distance of water
flow.  Reduction of  amitrole residue  from the bank  treated  at  the
 rate  of 3 Ib per  acre was 43 to 26 ppb over a distance of  water  flow
of 3  miles.
     Frequently  it is necessary to treat canal banks for weed control
prior to filling with water for the growing season. One such canal
was  treated with a 4-foot  swath  on both banks for a distance of 0.7
mile. On turning 50 cfs of water into  the canal, an initial concen-
 tration of 98 ppb of  amitrole occurred in the water front. This  resi-
 due level was reduced  to  46  ppb  at 1.3 miles downstream and after

                                 CHAPTER 13 / HERBICIDE RESIDUES  / 203

TABLE 13.4.  Dissipation of herbicides in irrigation wafer.

Miles Downstream            Dalapon          TCA          2,4-D

                              (ppb)           (ppb)          (ppb)
                                         Manard  lateral
      0.5  	     66              31            25
      4.25 	     40              20            14

                                          Yolo lateral
      0.5  	    289             ...            55
      3.0  	    182             ...            36
flowing 9 miles, the residue  level in the water amounted  to  only
23 ppb.
     Two irrigation laterals (Yolo  and Manard) were treated with  a
commonly  used mixture  of herbicides.  A  study was made  of the
resulting residue levels in the water and the extent of dissipation of
these levels as the water traveled downstream. Water volume and
the treatment rates of dalapon, trichloroacetic acid (TCA), and 2,4-D
used are shown in Table 13.2.  Residue levels in the irrigation water
0.5 mile below the treatment sites  and at the ends of the laterals are
shown in Table 13.4. The input of herbicide during bank treatments
such  as these was quite variable.  At any instant  it may vary  as
much  as ± 100%  of the  average  or calculated input.  Also,  as the
water  traveled downstream, water containing the maximum residue
level became a smaller fraction of the total volume of residue-bearing
water.  For  this reason values based  on  the average residue levels
may reflect more accurately the dissipation of herbicides in flowing
     In other studies, dalapon, amitrole-T, and the  isooctyl ester of
2,4-D were applied directly to irrigation water at constant rates, and
reduction in residue levels was determined as the water flowed down-
stream.  A canal which carried  16 cfs of water was sprayed for  75
minutes to provide  a mile  of water containing 400 ppb of 2,4-D.
The dissipation of residues of 2,4-D  was nonlinear. The  400 ppb
were reduced to maximum residue levels of 383, 285, 210, 206, and
190 ppb at distances of 0.1, 1, 3, 5, and 8 miles downstream, respec-
tively. These data show an initial rapid loss during the first 3 miles
of water flow, followed by a slow but constant decrease up to 9 miles.
     Another canal, which carried 19 cfs of water, was sprayed with
a solution of the sodium salt of dalapon for 51 minutes to provide a
mile of water containing 100 ppb  of  dalapon.  A plot of maximum
concentration in ppb against mileage gave a  straight  line with a
slope of 5.6.  Another canal which carried 49 cfs of water was  simi-
larly treated  for 18  minutes with amitrole-T to provide  a half-mile
length of water containing 50 ppb of amitrole  (Demint et al., 1969).
A similar plot, for the 5.25 miles sampled, gave a straight line with
a slope of 6.  Using these rates of dissipation, downstream mileage
at which total dissipation might occur was calculated as 18 miles for


  dalapon and slightly under 9 miles for amitrole. The conformance
  to linearity was  an indication that only 1 factor was involved for
  these 2 water-soluble herbicides.  Dissipation was achieved  through
  elongation of the herbicide  cloud.  The  magnitude  of the  dilution
  was so  great as to obscure possible losses from sorption or degrada-
  tion. Caution should be exercised in attempting to use these dissipa-
  tion rates to predict the complete disappearance of these herbicides
  to other canals. Among the  complicating factors are time required
  for complete dispersion, canal capacity changes attendant with flow
  rate changes, the retarding effect of bank treatments compared with
  idealized applications to the center of the canal,  and length  of bank

      Nearly all of the herbicides used for weed control in irrigation
 canals or on canal banks have been tested on most of the important
 field crops at  1 to 4 of our Agricultural Research Sendee  research
 stations in the western states (Arle, 1950; Bruns, 1954; Bruns et al.,
 1955, 1958, 1964; Arle and McRae,  1959).  The treated water was
 applied by flood or furrow irrigation methods in 1 to 3 acre-inches
 of water.
      In  general,  xylene-type aromatic  solvents, acrolein, amitrole,
 and dalapon were found to cause no injurious effects on crop growth
 or yields  at concentrations or rates used for weed  control.  Even
 2,4-D at rates up to 1  Ib, and usually 2 Ib,  per acre did not affect
 growth or yields of such sensitive crops as cotton,  grapes, and sugar
     The results of this research have verified the extensive experi-
 ence and observation  in  connection  with the widespread  use  of
 irrigation water on crops from canals treated  with aromatic solvents,
 acrolein, or copper sulfate and  on  which bank weeds were  treated
 with 2,4-D, dalapon,  amitrole, 2,4,5-T, or silvex. No known substan-
 tiated instances of damage to crops by any  of the  extensive uses
 during 5 to 20 years have  been reported. This extensive use  and
 experience have been documented in annual weed and pest  control
 reports of the 7 Bureau of Reclamation regional offices.
     In 1966 equipment was developed at  Prosser, Washington,  for
 field application to crops by sprinkler irrigation  of water containing
 herbicides.  This provided  an opportunity to  compare the effects  of
 herbicides in water  on  irrigated crops when  applied by overhead
 sprinkler and furrow methods.  It also provided an opportunity  to
 compare the amounts of herbicide residues assimilated by the crops
 when treated water was applied by  each of  the 2 methods.
     In 1967. 2,4-D and  silvex were applied to crops at rates of 0.1.
 0.5,  and 2.5 Ib per acre by furrow irrigation. These  rates provided
 concentrations  of 0.22,  1.11, and 5.55 ppm, respectively,  in 2 acre-
inches of water.  Only at the  highest rate of silvex did  significant
yield reductions occur in beet tops and bean  seed.  Small  but statis-
tically nonsignificant reductions were measured in beet tops  and
roots for  the 2  highest rates of 2,4-D.  There was no reduction in

                                  CHAPTER  13 / HERBICIDE RESIDUES / 205

 yield of corn fodder or grain by either herbicide at any of the 3 rates.
     Both 2,4-D and silvex were applied at rates of 0.01, 0.1, and  1.0
 Ib per acre by sprinkler irrigation.  These rates provided concentra-
 tions of 0.022, 0.22, and 2.22 ppm in  2  acre-inches  of irrigation
 water.  Surprisingly, both 2,4-D  and silvex  produced significant  in-
 creases in  the yields of sugar beet tops and roots at all  3 rates.
 Neither 2,4-D nor silvex affected the yield  of corn. The 2 highest
 rates of silvex reduced the yield of soybean seed, but the lowest rate
 of silvex and none of the rates of 2,4-D reduced the yield of soybeans.
 The rates for sprinkler irrigation were lower than those used for fur-
 row irrigation.
     In samples  taken  7 days after furrow irrigation, the highest
 rate of 2,4-D resulted in a residue of 0.11  ppm in beet roots, fresh
 weight (Bruns  and Comes,  1968).  No residues were found in other
 crop tissues. Samples of crop tissues taken at maturity showed  no
 residues of either  herbicide  after irrigations containing  0,22  or
 1.11 ppm.
     Low concentrations of 2,4-D were found in most crop tissues in
 samples taken 2 days after sprinkler irrigation at all rates. However,
 the highest concentrations from the 2 lower rates ranged up to 3.94
 ppm dry weight  basis  in  beet  roots.  These concentrations  were
 lower than  the tolerance of 5  ppm already  established  for 2,4-D in
 some food  and feed crops.  Also,  sugar  beet roots would never be
 used for feed or sugar production at that stage of growth. It is pos-
 sible that  sweet corn roasting ears or soybeans  as hay might  be
 harvested at that  immature  stage of growth. At maturity, when  all
 of these crops are usually  harvested, none of the crops contained
 any 2,4-D from the 2 lower rates and only beet roots contained 0.06
 ppm from  the highest rate,  1 Ib  per acre (2.22 ppm). This is 40 to
 50 times the highest concentration of 2,4-D found in water thus far,
 following applications for control of aquatic or bank weeds.
     No silvex residues were found in any  crop tissue receiving  the
 lowest rate of 0.1  Ib per  acre (0.22 ppm).  By normal harvest time at
 crop maturity,  most of the silvex residues had disappeared, even in
 crops irrigated with the highest concentration.
     Silvex  residues in  crop tissues  following sprinkler irrigation
 were found in  all crop  tissues from the 2 highest rates in  samples
 taken 2 days after harvest.  Also, soybean and corn foliage and beet
roots contained measurable residues from the lowest rate. However,
by normal harvest date at crop maturity, no residues were present in
 any crop tissues from the 2  lower  concentrations  of 0.022 and 0.22
of silvex in  irrigation water.
     Additional data on residues of 6 different herbicides in 6 differ-
ent irrigated crops are being obtained in  our contract research with
Stanford Research Institute. In this contract the crops were grown
in 2-gallon  greenhouse  crocks.  Each  crop was  irrigated at early
growth  and late growth  stages with 2 concentrations of each herbi-
cide.  The  treated water was applied in 1  acre-inch by both  flood
or soil and  overhead sprinkler irrigation  methods.  Results are now
available on 5 of the herbicides in all 6 crops (Stanford Research
Institute, 1968, 1969).
       No 2,4-D was found  in onions or soybeans from the highest


 rates of 0.22 and 1.11  ppm and the residues were negligible in car-
 rots, milo,  or potatoes.  Even in leaf lettuce the  residues were less
 than one-tenth the tolerance established for 2,4-D on some food crops.
     No silvex was found in milo, carrots, or lettuce from the highest
 concentrations of 0.22  and 1.11 ppm, and residues were very low in
 potatoes, soybeans, and onions.
     No amitrole residues were found in any of the tissues of green-
 house-grown  and treated crops or in field-grown  beans, corn, and
 wheat at Bozeman, Montana, which were furrow irrigated with water
 containing  4 Ib per acre of amitrole. In another  experiment at
 Prosser, Washington, no  amitrole residues were found in crops fur-
 row irrigated with water containing up  to 2.5 Ib per acre of amitrole
 (Bruns and Comes, 1966).
     The dalapon residues were determined  in  greenhouse-grown
 crops treated  and analyzed by Stanford Research Institute. The high-
 est rate used  was 0.5 Ib per acre (2.22 ppm) except on potatoes, car-
 rots, and onions.  For the latter 3 crops, the rates were increased 5-
 fold. Despite the heavy rates of  treatment on carrots and onions, the
 dalapon residues  were  very  low.  The highest  concentrations of
 dalapon were in soybeans, 1.18 to 2.79 ppm.  These  concentrations
 were less than one-tenth the tolerance of 30 ppm of  dalapon estab-
 lished by the  U.S. Food and Drug Administration  for asparagus.
     No diquat was found in any of the 6 crops which were irrigated
 by soil-flooding or overhead sprinkling at 0.09 or 0.45 ppm.  Because
 of the rapid dissipation of diquat  in water, irrigation water would
 seldom, if ever, contain a residue  of 0.45 ppm following a normal
 application for weed control.
     The  same  equipment that was  used  at  Prosser, Washington,
 for comparing effects and residues from furrow and sprinkler irriga-
 tion of  2,4-D and silvex in 1967  was used for comparing furrow and
 sprinkler irrigation of acrolein in 1966, and again in 1968 (Bruns
 and Comes, 1966, 1968).  The concentrations used were 0.1, 15, and
 60 ppm in 1966 and 0.1, 0.6, and 15 ppm in 1968.  Only the highest
 concentration, 60 ppm by furrow irrigation,  caused injury to soybean
 and sugar beet  foliage. The  injury  from  sprinkler  irrigation  was
 greater than that from furrow irrigation but no  injury occurred from
 concentrations used for weed contol.  None of  the  furrow irrigation
 treatments reduced corn yields.  Analyses of water samples showed
 that only 5  to 10% of  the acrolein was lost from the water during
 furrow irrigations. However, 60 to 90%  of the acrolein was lost from
 the water during sprinkler irrigation before the water fell on  the crop
 plants.  That  probably explains  why no damage to crops was ever
 reported by farmers who applied acrolein-treated water directly  from
 canals by sprinkler irrigation. Battelle Laboratories found no acrolein
in any of the crop samples.

     The effectiveness  of herbicides  and the economics involved in
agricultural production have caused their  extensive use for weed
control  in  and adjacent to aquatic  areas,  especially  on irrigation

                                  CHAPTER 13 / HERBICIDE RESIDUES / 207

systems.  As additional  data concerning residues and  toxicity are
developed, and as adequate  tolerances are  established for residues,
greater use of herbicides in and around agricultural waters  may be
     Maximum residues  of herbicides used for weed control in farm
ponds and reservoirs are low, ranging from a fraction of 1  ppm to
several ppm. In  most cases  these levels are of short duration. With
the exception of aromatic solvents and copper sulfate, most herbi-
cides occur in irrigation water at concentrations under 100  ppbJ
Only under the most adverse conditions in  small irrigation  laterals
are significantly  greater  residues found. The transport of herbicide
residues in irrigation water prevents extensive exposure of any given
irrigated area.  However, the flowing water  may  at times carry resi-
dues  to  areas where their  presence  may be  objectionable.  While
reduction in residue levels varies with the canal and herbicide, many
residues are dissipated after a water flow of  10 to 15 miles. In most
cases, the dissipation can be attributed to dilution in water or absorp-
tion by bottom mud.
     The concentrations  of herbicides  found in irrigation water are
unlikely to cause  injury in crops. Crop tolerance  studies showed that
crops can tolerate greater quantities  of herbicides  than would be
found in the water after applications for weed control.  Where resi-
dues were found  in  crops following irrigation with water containing
herbicides,  the levels  were  generally  much lower  than tolerances
already established  for  the  same or  similar crops.

Aldhous, J. R.  1967. 2,4-D residues in water following aerial spray-
     ing in a Scottish forest. Weed Res. 7:239-41.
Arle, H. F. 1950.  The effect of aromatic solvents and other aquatic
     herbicides on  crop  plants  and animals.  Proc.  Western Weed
     Control Conf.  12:58-60.
Arle, H. F., and McRae, G. N.  1959. Cotton tolerance to applications
     of  acrolein in irrigation water.  Western Weed  Control Conf.
     Res. Progr. Rept., p. 72.
Averitt, W. K.  1967. Report on the persistence of 2,4-dichlorophen-
     oxyacetic acid and its derivatives in surface waters when used to
     control aquatic vegetation.  Univ. of Southwestern Louisiana,
     Lafayette. Unpublished.
Bartley, T. R.  1969. Copper residue on irrigation canal.  Paper 98
     presented  at meeting of Weed Sci. Soc. Am., Feb. 11-13, Las
     Vegas, Nev.
Battelle-Northwest Laboratories.  1966,  1967, 1968.  Progress reports
     on herbicide residues in irrigated crops.  Unpublished.
Bruns, V. F.  1954. The response of certain crops to 2,4-dichloro-
     phenoxyacetic acid  in irrigation water.   I. Red Mexican beans.
     Weeds 3:359-76.
Bruns, V. F., and Clore,  W. J.  1958. The response of certain crops
     to  2,4-dichlorophenoxyacetic acid in irrigation water.  II. Con-
     cord grapes. Weeds 6:187—93.
Bruns, V. F., and Comes, R. D.  1966,  1967,  1968. Annual report of
     weed investigations in aquatic and noncrop areas. USDA, ARS,


      Crops Res. Div.  Unpublished.
 Brims, V. F., Hodgson, J. M., Arle, H. F., and Timmons, F. L.  1955.
      The use of aromatic solvents for control of submersed aquatic
      weeds in irrigation channels.  USDA Circular 971.
 Bruns, V. F., Yeo, R.R., and Arle, H. F. 1964.  Tolerance of certain
      crops to several aquatic herbicides in irrigation  water.  USDA
      Tech. Bull.  1299.
 Coats,  G. E., Funderburk, H. H., Lawrence, J. M.,  and Davis,  D.  E.
      1966.  Factors  affecting persistence  and inactivation of diquat
      and paraquat. Weed Res. 6:58-66.
 Cochrane, D. R., Pope, J. D., Jr., Nicholson, H. P., and Bailey, G. W.
      1967.  The persistence of silvex in water and  hydrosoil. Water
      Resources Res.  3:517-23.
 Davis,  E. A.,  Ingebo, P. A., and Pase, C. P.  1968.  Effect  of a water-
      shed treatment  with picloram on water quality.  Forest  Serv.
      Res. Note RM-100.  Fort Collins, Colo.: USDA.
 Demint, R. J., Frank, P. A.,  and Comes, R. D.  1969.  Amitrole resi-
      dues and dissipation rate in irrigation  water.  Submitted for
 Frank,  P. A., and Comes,  R.  D.  1967. Herbicidal  residues in  pond
     water and hydrosoil. Weeds 15:210—13.
 Frank,  P. A.,  and Demint, R.  J.  1967, 1968.  Annual report of  weed
     investigations. USDA, ARS. Unpublished.
 Grzenda, A. R., Nicholson, H. P., and Cox, W. S. 1966.  Persistence
     of four herbicides in pond water.  J. Am. Waterworks Assoc.
 Marston, R.  B., Schults, D.  W.,  Shiroyama, T., and  Snyder, L. V.
     1968.  Amitrole concentrations in creek  waters downstream
     from an aerially  sprayed  watershed  sub-basin.    Pesticides
     Monitoring]. 2:123-28.
 Nelson, J. L., Bruns, V. F., Coutant, C. C., and Carlile, B. L.  1969.
     Behavior and reactions of copper sulfate  in an irrigation canal.
     Pesticides Monitoring J.  In press.
 Norris,  L. A.,  Newton, M., and Zavitkovski, J.  1967.  Stream contam-
     ination with amitrole from forest  spray operations. Western
     Weed Control Conf. Res. Progr. Rept. pp. 33-35.
 Smith,  G. E., and Isom, B.  G.  1967. Investigations  of effects of
     large-scale  applications  of 2,4-D  on aquatic fauna  and water
     quality.  Pesticides Monitoring J.  1:16—21.
 Stanford Research Institute.  1968, 1969.  Progress  reports on herbi-
     cide residues in irrigated crops. Unpublished.
 Timmons, F. L.  1960. Weed control in western irrigation  and drain-
     age systems. USDA, ARS 34-14.
 	.  1963.  Herbicides in aquatic weed control. Proc. 16th South-
    ern Weed Conf.,  pp. 5—14.
    -. 1967.  The waterweed nuisance. In U.S. Dept. of Agriculture
    yearbook of agriculture, pp. 158-61.
Toth, S. J., and Riemer, D. N. 1968.  Algae control in inland water.
    Weeds Trees Turf 7:14-18.
U.S. Dept. of Agriculture. 1959, 1964. Agriculture census.
U.S. Dept.  of Agriculture.  1969.  Monitoring agricultural pesticide
    residues 1965-1967. ARS Rept. 81-32.
U.S. Dept. of the Army.  1965. Expanded  project for aquatic plant
    control.  House Document 251, 89th Congress, 1st Session.
Yeo, R. R.  1969.  Dissipation of  endothall in water  and effects  on
    aquatic weeds and fish.  Weed Science.  In press.


       HE late Paul Errington, an ecologist in our department, once
said that the human mind craves constants but in biology deals with
variables. The words maximum and minimum both connote such rel-
ative value judgments. Furthermore, pesticide usage has  been  ac-
companied  by certain ironies—controlling disease-carrying insects
has contributed to our population  crisis, and while  crop protection
has been a major factor in increased  production, this has often been
followed by reduced prices.  In an era when  science  and technology
are playing a major role in shaping  our society, it is altogether  too
easy for the individual scientist to lose his  objectivity and assume
that his particular insights entitle him to become a demagogue.  The
subject of pesticides has certainly lead to  such polarity  (Carson,
1962; Rudd,  1964;  Egler,  1964a,  1964b;  Whitten,  1966;  McLean,
1967).  The challenge today is for an enumeration of alternatives in
environmental management and an admission that with any strategy
there will be a certain amount of compromise. Pesticide usage con-
tinues to be confronted with the need for  compromise.  We may be
near the  end  of the golden  age of agricultural pesticide technology
since there seems to be a geometric increase in regulations regarding
the chemical inputs for pest control.  Wellman (1969) estimated that
the cost of developing a typical pesticide is now $4.1 million, up from
$2.5 million in 1964.
    In an  effort  to facilitate your  understanding  of this area,  I
would like to outline the pest management  strategies  available, relate
them  to  specific production commitments,  and  then  consider  the
ramifications to the role of agriculture in clean water. Consideration
needs to be given to both the quantity and quality aspects of this sub-
ject so that we can propose a rational compromise between pests and
    In entomology we often refer to pest population reductions which
occur without the influence of man as being natural controls. (The
     DON C. PETERS is Professor, Department of Zoology and Entomology,
     Iowa State University.



 term control has been overworked  to include the agent, the action,
 and the results.)  Natural control can  be subdivided  into climatic,
 edaphic, and biotic aspects. The reason for mentioning natural con-
 trol is that we hope we can understand it and capitalize upon it as we
 try to improve direct or applied pest control.
      Modern agriculture is still largely dependent on proper climatic
 conditions.  The same sunlight, moisture, and nutrients are utilized
 by weeds as well as planted crops.  Crop adaptation is a matter of
 growing the crop in an area where it has  at least some competitive
 advantages and applying  additional controls as  needed.  Since each
 organism has specific moisture, light, and temperature requisites, it
 follows  that pest species are not uniformly distributed  and man can
 capitalize on this knowledge. However, some diseases and insects may
 be carried great distances by winds. The most dramatic illustration
 of wind distribution would probably be the cereal rusts which have
 been referred to as continental pathogens because they spread from
 the subtropical regions to the north to cover the entire cereal acreages
 in North America.
     In nature,  diversity appears to be a solution to catastrophic out-
 breaks  and  destructive  changes.  Dasmann (1968)  said  that  com-
 plexity  appears to  be accompanied  by stability  and man seeks to
 simplify the complex so that he can manage it.  If a  great variety of
 plants are growing  in an area, the chance of spread for a host-spe-
 cific disease is greatly reduced. For this reason the chances for insect
 and disease  outbreaks are  much greater in cultivated  monocultures
 than in  natural areas. Under the conditions in northern forests, age
 may act as diversity.  As man has tried to manage forests and pre-
 vent fires, he has occasionally allowed large areas with trees of the
 same age to grow up. These may be attacked by insects or pathogens
 which normally attack only a specific age  category.  When such  at-
 tacks occur the losses are more severe than would be true of a forest
 with diverse age groups  or species of trees.
     I feel that a better understanding  of  the balance of nature is
 needed for a meaningful communication of the  science of pest con-
 trol. "Key factor analysis" is a recent concept used by insect ecologists
 such as Clark et al. (1967) in trying to characterize the major factors
 contributing to  population levels of insect  groups. An extension of
 this approach may  be the reason why in  each crop  we have  a few
 major persistent pests, several species which become pests  during
 sporadic outbreak periods,  and an additional group of potential pests
 associated with  a larger number of  species which cause no  damage
 but occur in  the area as scavengers  or parasites  and  predators. The
 interactions between these groups are frequently drawn in a web
 configuration, but this may communicate a concept of peaceful coop-
 eration whereas intense competition for resources is more in line with
 the "key factor" approach.   Work summarized in the  National Acad-
 emy of Science  (NAS) publication  on insects (1969) indicates that
 food may be a key factor in regulating a pest, but that parasites and
 predators,  disease,  weather, and migration have been found  to  be
key factors with as great a frequency.  As an illustration, the Colorado
potato beetle in  Canada was found to be limited by food.  However,

                     CHAPTER 14 / PESTICIDES AMD PEST MANAGEMENT /  211

I doubt that any potato farmer would consider it reasonable to allow
this vegetative feeder to completely devour the above-ground parts of
the plant before some direct means of control was sought.
     It is my impression that similar relationships between crops and
pests exist in the realm of plant pathogens and to a different degree
in weed competition.  I feel that the work of Kooper (1927) and Holm
(1969) relating to the competition between plants by  growth inhibi-
tion of one species by another encourages speculation that if we knew
what inhibits some seed growth in the presence  of other plants a
more effective weed control could  be  achieved. Species competition
should be managed for our good.
     One other point I would like to emphasize before discussing  ap-
plied controls is that when man put his hand to the plow and began
to modify plant diversity, he began a high-risk enterprise.  There  are
still no absolute measures of what is progress as far as manipulating
the disturbed cultivated environment.  Many of us have gone along
with  Swift's adage of the man "who can make two ears of corn or
blades of grass to grow where  but one grew before," but I feel that
most thinking biologists today have conceded that  man is not capable
of continuing to feed himself and his progeny unless he devises effec-
tive means of regulating his population.  The question of the quality
of our environment is another thing that merits more attention.

     Applied controls are those biological, cultural, legal, or chemical
practices which man utilizes in an effort to reduce losses caused by
pests.  Each of these has its disadvantages and advantages and the
cost/benefit ratio needs to  be continually investigated in a dynamic
agriculture  and civilization. Paul Sears,  in  a recent  visit to  Iowa
State, warned that another danger in this scientific age was doing
things simply because they became technically feasible. For example
an insect-free cornfield may not be the most desirable condition. Let
us first consider biological  control which may be either natural or
applied. Biological control  probably has its main desirable aspects in
that it usually produces no  side effects and frequently is a one-time
operation.  Once it is set in motion there need not be an annual cost
for crop production. Biological control works best where some  dam-
age can be incurred to the crop without serious economic loss and
where the soil is not disturbed. This means that we should look for
the most frequent  successes in forest lands and in orchards, and the
least successes for biological control in the intensive cultivation  prac-
tices of truck farming.
     The three  main aspects involved in the utilization of parasites
and predators  are introduction,  conservation, and  augmentation.
While the introduction of an insect species for control of another in-
sect or weed is a complicated matter (NAS, 1968b), there have been
sufficient successes in this area, particularly in those instances where
the pest was not native, that continued work is certainly justified.  It
has been estimated that if the program is effective, 80% of the intro-


 ductions are effective within three generations.  By the conservation
 of parasites and predators, I have reference to such situations as strip
 mowing  of alfalfa so that the shelter for predators is not completely
 eliminated at any one time during the production period. The aug-
 mentation  of field populations by  laboratory-reared  parasites and
 predators has met with varying success.  It seems to show more
 promise  where the target insect infests  a localized  area and where
 the  parasites and predators are  limited  to the immediate  area.  I
 know of no successful program of augmentation in the Upper Missis-
 sippi Valley.
     Insect pathologists have been working on diseases of insects for
 over a hundred years, and a recent report (NAS, 1969) indicated that
 there  were 1,165 microorganisms which attacked insects. In this re-
 gion disease agents have been used against the European corn borer
 and Japanese beetles.  However, there is  a possibility of the insects
 developing resistance to these diseases. A recent paper by Hoage and
 Peters (1969) demonstrated the ability of honeybees to develop larval
 resistance to American foulbrood disease. Similar disease resistance
 probably occurs in nature as part of the  overall web of competition
 and survival of the fittest.
     I have chosen only to mention and not discuss  some other con-
 cepts in biological control such as the areas of competitive displace-
 ment, antimetabolites, feeding deterrents, or genetic regulation of
 pests since these are still largely in the investigative stage and lack
 working field programs to confirm their potential.
     Host-plant resistance is frequently considered as a part of the
 biological control approach and certainly it is a modification of the
 host organism in an effort to reduce losses from pest infestations.  It
 is doubtful if we would be able to continue cultivation of any of the
 cereal crops without disease-resistant cultivars. And yet in the case of
 the cereal rusts, we are probably witnessing evolution working at an
 extremely rapid and efficient rate but not toward our varietal improve-
 ment goals.  Van der Plank (1968) is  optimistic and  states that crop
 breeders  should continue their work on developing disease-resistant
     By  contrast to the great  number  of disease races that have
 cropped  up in relation  to  varietal  resistance,  the  story on insect-
 resistant  crops is not nearly so complex. Three  exceptions  are the
 corn leaf aphid and pea aphid races or "biotypes" reported by Carrier
 and  Painter (1956) and  Carrier et al.  (1965)  and  the  Hessian fly
 where there  are currently at least four races (Gallun et al.,  1961).
 Host-plant resistance is probably the most ideal means of controlling
 the major disease and insect pests of  the major crops. Development
 of resistant varieties  does entail a considerable expenditure of time
 and  the  cooperative effort of a team of investigators  and therefore
 will probably be limited to  only the major pests of the various cul-
 tivated crops. The potential for breeding insect- and disease-resistant
 animals is certainly not great.  I have no  idea of how one would go
 about breeding a corn plant for resistance  to foxtail competition. All
 of the  biological control approaches require a lot of specific research
before they can be utilized.

                     CHAPTER 14 / PESTICIDES AND PEST MANAGEMENT / 213

     Cultural pest control is among the oldest of man's practices in
trying to come to grips with his pests. Sanitation as illustrated by
crop residue destruction and animal waste removal is an important
means in reducing the breeding potential of a number of pests.  Till-
age practices can have an impact on any of the three pest groups that
we have been considering. By way of illustration, it is hard  to say
whether the Iowa farmer should plow his cornstalks under to control
European corn borers, weeds, or the  yellow leaf blight disease. Re-
duced tillage may encourage some pest species, but increased  tillage
will also destroy many of the organisms that would tend to  afford a
competitive balance between the organisms in the field.
     The economics of current production practices in the Corn Belt
leave little  leeway for pest management in timing the planting opera-
tion  or  the intensity of fertilization.  Since both of these need to be
maximized from  an  agronomic standpoint, workers in pest  manage-
ment are confronted with the need to devise some means of  compen-
sating for agronomic practices which may be at odds with optimum
pest  control.  Early harvesting can certainly help to avert some of the
potential losses that  might otherwise  be attributed to stalk-attacking
insects  or diseases.
     Physical or mechanical controls are seldom of importance in the
large acreages of cultivated crops common in modern agriculture, but
such  things as the flaming of  alfalfa fields may reduce the  alfalfa
weevil threat and give some reduction in  the chickweed problem as
     Another illustration of mechanical means is the light  trap.  As
far as reducing crop pests, light traps have been of limited value, with
the most favorable data coming from the  tobacco-growing area in
North Carolina (Lawson et al.,  1966). There is also  a report of re-
duction in  Heliothis  spp. as cotton pests in Texas, following the use
of artificial light  (Nemec, 1969).
     Insect sterilization has received a lot  of popular publicity in the
past  10 years because of the success of the  screw worm program in
southeastern United  States (Bushland et al., 1958) and more recently
in the Texas area. However, there are several drawbacks to this ap-
proach. It is extremely expensive in comparison to other programs
with which entomologists have been  associated.  Sterilization would
not appear to be  practical where a pest overwinters in an extensive
area or where numbers are not severely reduced in the spring. Chemo-
sterilants have been and are being investigated but in the past decade
they have not proved to be commercially acceptable in even a single
field program in the United States.
     The potential use of attractants and repellents still must be con-
sidered nebulous, although there have been some excellent  results
where attractants and insecticides were combined on island situations
in eradication programs (Beroza, 1966).  Personally, I have serious
reservations about man's ability to totally eradicate any insect pest
species  from the  continents. Eradication  of weeds and diseases is
even less likly (NAS, 1968a).
     Chemical control of pests is an old practice.  It probably began
when the Arabians discovered the benefits  of sulfur for louse control


 for their horses or with the early observations of the herbicidal effects
 of salt water.  It has only been during the last 40 years that man has
 begun to synthesize chemicals rather  than to depend  upon  those
 which he could obtain by mining or refining. The intensity to which
 he  has used  these synthesized products has had a considerable in-
 fluence in the gains in production potential on  a number of crops as
 illustrated in a paper by Decker (1964).  I have tried to update these
 production figures in Table 14.1.  Insecticide use on  oats, hay, and
 soybeans has been low.  For oats and hay the returns have been low
 but the per acre net return from soybeans has been almost as good as
 corn.  The relative increases in per acre yields  for corn,  cotton, and
 potatoes since the advent of DDT and other organic insecticides has
 been much greater than for oats, hay, or soybeans. I certainly do not
 believe that the yield increases are entirely caused by insect control,
 but the insecticides must obviously be aiding a total production pro-
 gram.  There may also have been a profit differential that justified the
 decision to use the chemicals at the time the first synthetic pesticides
 were applied.  The economics of production would appear to continue
 to dictate similar pesticide use patterns.
     The hope held  out for growth-regulating hormones  as  "third
 generation insecticides"  by Williams (1967) may be only a hope, and
 certainly many of us will need  to change our attitude about taxes if
 these hormonal mimics are to be used.  I doubt whether industry will
 be willing to expend the resources necessary to develop these specific
 means of control. I would expect the financial returns to be consider-
 ably less favorable than with  the conventional multi-use pesticides
 available today.  Persing (1965) wrote that if DDT were  specific for
 houseflies, its profits would not have equaled research and develop-
 ment costs.  I have heard  a lot of talk  to the contrary, but specific
 pesticides have not been forthcoming in the past  decade.  A  good
 demonstration of the problem is a product by the name of Manazon
 which is excellent for aphid control but apparently the  company own-
 ing this product does not feel that it would be a profitable product to
 develop at this time.  By contrast, the top ten pesticides in 1967 sales
 were all broad-spectrum  materials (Mahan et al., 1968).
     There are certainly many  pitfalls  that can arise  from  over-
 dependence on use of chemicals in crop production.   Smith (1967,
 1969a,  1969b) has done an excellent job of describing  some  such
 problems in cotton production and he also tells of  the potential in-
 tegrated control has as a means of maximizing the effectiveness of
 chemical applications. As Mills (1968) indicated, we need  to continue
 to sharpen our entomological knowledge  of space and time in insecti-
 cide applications. Knipling (1966) has  calculated that 1 Ib  of the
most effective boll weevil insecticide would be enough to kill  all the
boll  weevils in the United States if applied topically  in the  spring
when weevil levels are lowest.
     These are the pest control alternatives available.  The next ques-
tion is, How and to what extent are these being used in various pro-
duction programs?

TABLE  14.1.  Average yields for selected crops in 48 states.

Soybeans . . .
in 1966



Yield per
Acre for Yea
rs Indicated


49.2 bu.
1.94 tons
25.5 bu.
76.5 bu.
479.0 lb.
211.0 cwt.
 * Ratio of production in each period to production from  1901 to 1940, except in soybeans where ratio base used  was



      There are a number of functional considerations or variants in-
 volved in considering the pest problems as related to the food, shelter,
 and  clothing  areas  of production. There  is also a need to keep in
 mind the "aesthetic needs" of man.  How many "Madison Avenue"-
 gendered needs can we afford on our crowded planet?
      In the area of food production, the cereal crops and potatoes are
 the major carbohydrate sources.  In the United States, the percentage
 of cereal crops receiving insecticide  treatment in  1964 was around
 3%  for small grains, but about 33%  for corn, of which a large per-
 centage  was for soil insect control.  According to  Fox et al. (1968)
 chemicals for  disease control were used on less than 0.5% of the
 acres on all of these crops, while herbicides were used on 57% of the
 corn acres and about 30% of the acres planted to the other grain
     The amount of money spent on livestock pest control, according
 to Gale et al. (1968), was less  than 5%  of the total farm use, and the
 estimated pounds of insecticide were an even smaller ratio.  In spite
 of this relatively  small  volume, the  point source  principle used in
 identifying and detecting pollution may present problems  for pesti-
 cide usage on livestock and poultry.  However, the major problem is
 from misuse or contamination, since there is little likelihood of water
 contamination from the materials used in fly control today.
     Production of fruits and vegetables is the most intensive high-
 value crop production program, but as pointed out previously there is
 a considerably greater potential for biological strategies to effectively
 control orchard pests as  compared to truck  farming operations.  In
 1966 Fox et al. (1968) found that 28% of the vegetable acres in the
 48 contiguous states were treated with a herbicide,  whereas the per-
 centage of apple acres treated  was 16% , and  that for other deciduous
 fruits was only 13% . Similar figures on insecticides indicate that use
 on vegetables was 56% , on  apples  92% ,  and  on other  deciduous
 fruits 72% . Apparently, the consumer insistence on perfect fruit has
 encouraged a lot of spraying. While Irish potatoes would normally be
 considered a carbohydrate source, the percentage  of crop acres  on
 which insecticides were used was 89%  over the contiguous states, but
 reached  100%  in the southeastern and southern plains states.  The
 demand for fruits and vegetables free of insect damage has certainly
 been met with an intensive use of pesticides.  With current harvesting
 and processing it is doubtful if this usage can be changed.
     Turning to clothing, I have  already indicated  that a very small
 percentage of the total amount of pesticides is used on  livestock and
 as one would expect  the amount of pesticides used in wool production
 would be minimal.  By contrast the proportion of herbicide and in-
 secticide usage in cotton  production would be far  greater than that
 for other field  crops  with  the exception of tobacco and potatoes.  Ac-
cording to  Gale et al. (1968), in 1964  the United States average  per
 acre  pesticide  expenditure was $11.27 for cotton compared to only
 Si.87 per acre of corn.   This illustrates the deceptive potential of
figures since there were 66 million acres of corn as compared to  14

                     CHAPTER 14 / PESTICIDES AND PEST MANAGEMENT / 217

million acres of cotton. Therefore the total expenditure for cotton
was only 20%  greater than  for corn pesticides.  In another sense
pesticide use on cotton is much more intensive and does not allow for
as great a dilution as it enters the environment.
     The  current  trend is  toward more  synthetic fibers.  There  are
some indications that these too may possibly have some harmful side
effects. Determining the long-term influences of these products on
experimental animals should be pursued with the same rigor as has
the toxicology of chemical control agents.
     Man's need for shelter is influenced by diseases and insects only
to the extent that he utilizes wood in providing these shelters.  The
critical times for timber seem to be  during the establishment of  the
young trees, as the standing crop nears harvest, during the processing
period, and after the structure has been completed, when termite and
decay problems may arise.  Economics of lumber production are such
that it has not been feasible to treat large acreages repeatedly. Conse-
quently current estimates are that less than 5%  of our forest lands
have ever been sprayed with any insecticides.  The  hazard  is that
when forest areas are sprayed it is usually done in large  contiguous
blocks treated as part of a federally coordinated program.  Such mas-
sive programs usually involve large aircraft in terrain where it is not
expedient to avoid spraying of streams and other areas where fish and
wildlife are concentrated.  Therefore, while the  direct problems of
chemical treatment to the wood are nil, the ramifications  to  the fish
and wildlife populations may be considerable, since  the only logical
places for significant  wildlife populations  are in the forests  and
ranges of the United States.
     I would like to consider the aesthetic ramifications  associated
with agricultural production.  One concern is with  the farm fence
row.  To many people the  uniform growth of grasses which can be
achieved by annual 2,4-D spraying is desirable. Others like  a diver-
sity in plants and do not find this uniform grass population appealing.
It is certainly true that this is a more costly practice than allowing the
plants to grow as they wish. It is difficult to extrapolate Scott's (1938)
data to modern times with increased miles traveled on our primary
and secondary roads, but it  is time we determine if wildlife would
increase if ditches and roadsides  were left to grow up in a  natural
vegetation as game cover.
     Most of us  became too embroiled in the Dutch elm disease con-
trol  program to consider it with  much objectivity.  While  this has
basically been an urban rather than  a rural problem, I believe  that
there are lessons  to be learned from  the successes  and  failures of
various strategies and tactics tried in controlling this pest. The Iowa
Cooperative Extension  Service (1961) outlined a 4-step program of
(1)  evaluation and education, (2) sanitation,  (3) maintenance,  and
(4) spraying. Steps 1 through 3 were seldom executed effectively,  but
the extension entomologists were certainly blamed for any  robins that
died after DDT spraying. All phases of agriculture must get involved
in public  information and  communication.
     Home gardens, lawns, and flowers are not usually considered an
agricultural problem and yet in the area of pesticide pollution they


 should not be overlooked. I submit that there is as great a probability
 that the suburbanite will dump the leftover spray into the sewer as
 there is that the agriculturalist will  dispose of his leftover pesticides
 in such a manner as to directly contaminate water sources.  When one
 considers the population ratio between rural and suburban peoples in
 the  United States  today,  the magnitude  of  this problem becomes

     What then are the ramifications of pest management to agricul-
 tural waters? If we assume that changes will be brought  about by
 the due process of legislation  and education I think we can  make
 some fairly good assumptions and suggestions as to what can be done
 to educate ourselves about the proper use of pesticides in farm pro-
 duction programs.  Figures  14.1  and 14.2 on land uses  and relative
 intensity  of pesticide use on crops in the United States should put
 the problem in perspective.  First let us consider small grain produc-
 tion.  With the  present net return from  these crops, it is  doubtful
 whether additional chemicals will be used in an effort to achieve more
 efficient production.  These crops  are  grown on  relatively  large
 acreages and there is little likelihood that yield can be increased with-
 out the addition of irrigation.  Cultural practices  and host  plant re-
 sistance should continue to  be mainstays of pest control  on  small
     After visiting with  agronomists, agricultural  engineers, and
 others interested in corn production, I believe that future corn pro-
 duction will see increased emphasis on narrow row  spacing with  a
 moderating trend  in the immediate future to  30-inch row  spacing.
 Even the shift to 30-inch row  spacing means essentially a  30%  in-
 crease in insecticide usage to achieve the same amount of rootworm
 protection as compared to 40-inch row spacing. To partially  offset
 this we have been working (Peters, 1965; Munson et al.,  1970)  to try
 to combine one chemical treatment for both corn rootworms and Eu-
ropean corn borer control. Some of the current insecticides under in-
vestigation might control com leaf aphids also, and thereby  get three
birds with one stone.  The emphasis on minimum tillage will need to
be watched and possibly modified in the future in line -with efficient

        FIG. 14.1.  Primary  land
        use, 1964.                 i °r*e£^\i    PASTURE
                                 L	—^^    & RAN,GE

                      CHAPTER 14 / PESTICIDES AND PEST MANAGEMENT / 219
                                           FIG. 14.2.  Insecticide  use
         33^     \   /                       on  U.S. crops, 1964.  Insec-
    CCR\Y         \   /                       ticides were  applied  on
                             „, ' *LCHE' BEETS 52%  of  the  small  grain
                             /o —OTHER              _   .,   .   =
                               J FIELD CROPS  acreage. Small grains were
                                           grown on more acres than
                                           any other crop.
 use of herbicides and  insecticides.  Disease problems may become
 greater if crop residue is left on the soil surface. The ideal system for
 corn would be to maintain the intertillage  system so  that the space
 between the row need not be  treated with pesticide.
     We have been searching  for a replacement for aldrin and hepta-
 chlor, the two major chlorinated hydrocarbons still used in corn in-
 sect control. To date we have not found a product that will control
 white grubs, wireworms, or cutworms at economically feasible rates.
 The carryover problem  of herbicides is one reason given for planting
 corn after corn.  This means that the use of persistent herbicides has
 intensified the need for corn  rootworm control measures since corn
 rootworms do economic damage mainly  in fields of corn  following
 corn.  If this situation can be  alleviated we would hope that the acres
 treated for corn rootworms can be stabilized or even reduced.  How-
 ever, it is possible that with a major change in crop sequence patterns
 other insect pest problems may increase in importance; Cole (1966)
 has pointed  out  that all  communities harbor  opportunistic species.
 The disease problem seems to be increasing under higher plant popu-
 lations.  Some promising disease  control chemicals are being  eval-
 uated but the potential economics of such pesticide usage has not
 been worked out.
     At the risk of sounding provincial, I would like to emphasize  that
 soybeans should  be grown in  the northern part of the United States
 rather than as a  replacement  for cotton. In the South  chemicals will
 have to be used to control corn earworms,  stink bugs, and other insect
 pests of soybeans, whereas these insects have not occurred in damag-
 ing numbers in soybean culture in the Upper Mississippi Valley. I feel
 that this is  a place where timely legislative action could reduce the
 overall pesticide burden in the United States.
   Cotton culture is an  enigma for it is profitable under the present
 allotment system to use relatively large  quantities of pesticides to
 maximize yield on a per acre basis.  If allotment were on  a basis of
 pounds of lint cotton per farm unit it might  be  possible to reduce
 pesticides.  Another consideration might be the emphasis of produc-
 tion on those areas where  cotton pests are less of a problem.  Addi-
 tional irony in this  situation is that while Smith (1969b) and his
 associates in California have  done  an excellent job of describing in-
tegrated control to the scientific community, the amount of pesticide


 money spent per acre of cotton production in California was higher in
 1966 than any other area of the country.  It is hoped that the Califor-
 nia entomologists can devise as efficient a means of communicating to
 the growers in the state as they have to the scientific community.
      There is still a demand for tobacco products.  This high intensity
 crop will probably continue to utilize large amounts of pesticides in
 localized areas. If hopes are realized, continued work on mechanical
 or biological controls can reduce the insecticide usage for this crop.
 While the total tonnage of pesticide used may be small, the likelihood
 of local stream or pond contamination is still very real.  Potatoes and
 sugar beets also present a. problem since the net return to the grower
 is very small unless a high yield per acre can be achieved. Therefore,
 we again have the potential of a point source contamination situa-
 tion. Since these  crops are grown by relatively few farmers it should
 be possible to amplify the educational effort toward  the reduction of
 unnecessary treatments or low return treatment.
      The juxtaposition  of  metropolitan areas and intensive  truck
 farming will continue to be the cause of friction in pesticide usage.
 The  high value commodities would seem  to justify the use  of pesti-
 cides in order to insure a favorable return to the grower; however, the
 chance  of water contamination is a continuing problem when such
 usage is adjacent  to streams, ponds, and lakes with high recreational
      Every effort  should be made to encourage fruit growers to in-
 vestigate the feasibility of enhanced biological control in their or-
 chards.  An important corollary for  such a program to be completely
 successful is that we will also need to educate the consumer to accept
 less than perfect produce.
     There is currently a boom  in large feedlot operations in  the
 southwestern states of Texas, Oklahoma, Kansas, and Colorado.  This
 has been brought  about, in part,  by increased irrigation of sorghum
 and corn as  a grain source for livestock feed.  These lots tend to be
 large and the livestock  are confined.   If we continue to emphasize
 that "the solution to pollution  is  dilution" then this  is a step in  the
 wrong direction.  It is true  that these are areas where  the moisture
 problems in feedlots are reduced,  but if pesticides are needed to con-
 trol flies and other insect problems under  the crowded livestock con-
 dition, it is only conjectural as  to what would be the concentration of
 insecticide in the  small  amount of runoff that  might occur from
 these areas.
    If we really feel that the  chlorinated hydrocarbons are a detri-
 ment  to our  society,  we should encourage  an  immediate ban on all
 small-package registrations of  these products. This is an  area in
 which misuse is most likely to occur  since the homeowner-gardener is
 not confronted with the cost/return ratio to the extent that people in
 crop production are.
    I feel that I must take a stand in opposition to many of the large
 federal  insecticide pest control programs.   These  have been subject
to considerable criticism, for even though  they achieve a major cost
reduction per treated unit, the  tendency, as pointed out by Cope  and
Springer (1958), is that by achieving these efficiencies in distribution

                      CHAPTER 14 / PESTICIDES AND PEST MANAGEMENT / 221

 there is a comparative loss in target precision and effectiveness.  I
 believe the Plant Pest Control Division would do well to continue its
 emphasis  on  the  distribution  of biological control  agents or at-
 tractants.  The substitution of Mirex bait in the fire-ant program was
 certainly a progressive step in the right direction.
      In summary, I would say that the various  agencies in pest con-
 trol have been and  continue to be concerned about the use of pesti-
 cides in relation to  the total environment. We  are working and will
 continue to work to the limit of personnel and  funds available. The
 need for increased funds  in the future  is great  since the newer
 strategies  are  of a nature that will require  public support for their
 application.  Recently a national joint task force  on pollution proposed
 less than a 1% increase in effort for the pesticide area for the next
 decade!  If the public demand for sophistication in pest control is to
 be achieved, more imaginative research support will have to be found.
 Man is the dominant species on earth today and the question is not if
 he will  modify the environment, but  the question is how can he
 modify the environment in such a way as to achieve a stability which
 will allow his long-term existence.
      Starr's (1969)  article on social benefits versus technological risk
 merits careful consideration.  In order to maximize pest management
 for maximum  production, crop protectionists should expect a reason-
 able  risk ratio along with other agricultural, industrial,  and urban
 sources of water pollution.  The alternative loss of 10 to 30% of our
 basic agricultural production to pests and diseases needs to be held
 before the  consuming public.

Beroza, Morton.  1966.  The  future role of natural and synthetic at-
     tractants for pest control in pest control by chemical, biological,
     genetic, and physical means. USDA, ARS.
Bushland, R. C., Knipling, E. F., and Lindquist, A. W. 1958.  Eradi-
     cation of the screw-worm fly by releasing gamma-ray  sterilized
     males among the natural population.  Proc. Intern. Conf. Peace-
     ful Use Atomic Energy Geneva 12:216-20.
Carson, Rachel. 1962. Silent spring. Boston:  Houghton Mifflin.
Cartier, J. J., and Painter, R. H.  1956. Differential reactions of two
     biotypes  of the corn leaf aphid to  resistant and  susceptible
     varieties, hybrids and selections of sorghums.  /.  Econ.  Entomol.
Cartier, J. J., Isaak, A.,  Painter, R. H.,  and Sorensen, E. L.  1965.
     Biotypes of pea aphid Acythosiphon pisum  (Harris) in relation
     to alfalfa clones. Can. Entomol. 97:754-60.
Clark, L.  R., Geier, P. W., Hughes, R. D., and Morris, R. F.  1967.
     The ecology of insect populations in theory and practice.  Lon-
     don:  Methuen.
Cole, Lament C. 1966. The complexity of pest control in the environ-
     ment.  In Scientific aspects  of pest  control, pp. 13-25.  Nat.
     Acad. Sci., Nat. Res. Council Publ. 1402, Wash., D.C.
Cooperative  Extension Service. 1961. Diseases and insects attacking
     Iowa elms. Iowa State Univ. Pamphlet 250 (Rev.)


 Cope, O. B., and Springer, P. F.  1958.  Mass control of insects: the
      effects on fish and wildlife. Entomol. Soc. Am. Bull. 4:52-56.
 Dasmann, R. F.  1968.  Environmental  conservation.  2nd ed. New
      York:  John Wiley.
 Decker, George C. 1964.  The past is prologue.  Entomol. Soc. Am.
      Bull.  10:8-15.
 Egler,  F.  E.   1964a,  Pesticides in  our ecosystem.  Am. Scientist
      52(1): 110-36.
 	.   1964b.  Pesticides in our  ecosystem:   communication.  II.
      BioScience 14 (11): 29-36.
 Fox, Austin, Eichers, T., Andrilenas, P., Jenkins, R., and Blake, H.
      1968.  Extent of farm pesticide use on crops in 1966.  USDA,
      Agr. Econ. Rept. 147.
 Gale, J. F., Andrilenas,  P., and Fox, A.  1968. Farmers' pesticide ex-
      penditures for crops, livestock, and other selected uses in  1964.
      USDA, Agr. Econ. Rept. 147.
 Gallun, R. L., Deay, H.  O., and Cartwright,  W. B. 1961.  Four races
      of Hessian fly selected and developed from  an Indiana popula-
      tion. Purdue Univ. Res. Bull. 732.
 Hoage, T.  R., and Peters, D. C.  1969.  Selection for American foul-
      brood resistance in larval honey bees.  /. Econ. Entomol. 62:
 Holm, LeRoy.  1969.  Chemical interactions between plants on agri-
      cultural lands. Doivn Earth 25:16-22.
 Knipling, E. F.  1966. New horizons and the outlook for pest control.
      In Scientific aspects of pest control, pp.  455-70. Nat. Acad. Sci.,
      Nat. Res. Council Publ. 1402. Wash., D.C.
 	.  1968.  The role  of chemicals in the  general insect control
     picture. Entomol. Soc. Am. Bull.  14:102-7.
 Kooper, W. J. C.  1927.  Sociological and ecological studies on weed
     vegetation of Pasurian. Rec. Trav. Bot. Neerl. 24:1-255.
 Lawson, F. R., Gentry, C. R., and Stanley, J. M.   1966.  Experiments
     on  the control of  insect populations with light traps in pest
     control by chemical,  biological, genetic, and physical means.
     USDA, ARS.
 McLean, L. A.  1967.  Pesticides and the environment.  BioScience
 Mahan, J.  N., Fowler, D. L., and Shepard,  H. H.  1968.  The  pesti-
     cide revieiu 1968.  USDA, Agr. Stabilization  and  Conserv. Serv.
 Mills, H. B.  1968.  Summary  and  conclusions  in  Symp. on the
     Science  and Technology of Residual Insecticides in Food Pro-
     duction  with Special Reference  to  Aldrin and Dieldrin.   Shell
     Oil Co.
 Munson, R. E., Brindley, T. A., Peters, D. C.,  and Lovely, W. G.   1970.
     Control of both the European corn borer and corn  rootworms
     with one  application of  insecticide.  Submitted  to J.  Econ.
 National Academy of Sciences.  1968a. Plant-disease  development
     and control. Principles of plant and animal pest control. Vol. 1.
 	.  1968b.  Weed control.  Principles of plant  and  animal pest
     control.  Vol. 2.
        1969.  Insect-pest management  and control.   Principles  of
     plant and animal pest control. Vol. 3.
Nemec, S. J.  1969. Use of artificial lighting to reduce Heliothis spp.
     populations in cotton fields.  /. Econ. Entomol. 62:1138-40.
Persing,  C. O.  1965.  Problems in the development of tailor-made

                     CHAPTER 14 / PiSTICiDSS AND PEST MANAGEMENT / 223

    insecticides,  specific insecticides.  Entomol.  Soc. Am.  Bull.
Peters, D. C.  1965.  Chemical  control of  resistant corn rootworms
    in Iowa. Entomol Soc. Am. Bull. 20:58-61.
Rudd, R. 1964. Pesticides and the living landscape. Madison:  Univ.
    of Wis. Press.
Scott, T. G.  1938. Wildlife  mortality on Iowa highways. Am. Mid-
    land Naturalist 20:527-39.
Smith, R. F. 1967. Principles of measurements of crop losses caused
    by insects.  FAO Symp. on Crop Losses, Rome, 2-6 Oct. 1967,
    pp. 205-24.
	.  1969a. The importance  of economic injury levels in the  de-
    velopment of integrated pest control programs. Qualitas  Plant.
    Mater. Vegetables 17:81-92.
	.   1969b.  Patterns of crop protection  in  cotton ecosystems.
    Mimeo  of talk given at Cotton Symp. on Insect and Mite Control
    Problems and Res. in Calif., 12-13 March  1969, Hotel  Clare-
    mont, Berkeley, Calif.
Starr, Chauncey.  1969.  Social benefits versus  technological risk.
    Science 165:1232-38.
Tukey, John W.,  chairman.  1965. Restoring the quality of our  en-
    vironment.   Report  of  the  Environmental  Pollution  Panel of
    President's Science Advisory Committee.
U.S. Dept. of  Agriculture.  1966a. Field crops by states,  1959-64.
    Statistical Bull.  384.
	.  1966b. A century of agriculture in charts and  tables.  Agri-
    culture Handbook 318.
Van der Plank, J. E.  1968.  Disease resistance in plants.  New York
    and London:  Academic Press.
Wellman, Richard H.   1969.  Ag  chemicals industry  faces  big
    changes.  Chem. Eng. News, pp. 22-23.
Whitten, J.  L.  1966.  That we may live. Princeton, N. J.:  D. Van
Williams, Carroll M.   1967.  Third-generation pesticides.  Scientific
    Am. 217:13-17.


 DON C. PETERS, Leader
 H.  B.  PETTY, Reporter
       I HE general public  at  present  suffers from  lack  of  factual
 and realistic information pertaining to (1) the real role and impor-
 tance  of pesticides in food production,  (2) the present restrictive
 regulations which  govern labeling, sale, and use of pesticides,  (3)
 the accident-safety record of pesticide use,  and (4) the food-monitor-
 ing work of the HEW,  FDA,  which so carefully protects our food
 supply from any contamination that could be considered deleterious.
 The general public is unaware of extensive use-education programs
 and they have the opportunity to read only the overpublicized alarm
 stories, many of which are not realistic.  There is a need, therefore,
 for the Cooperative Extension  Services to expand their programs to
 include more than just agriculture.

     Those interested in clean water  and pesticides must realize that
as long  as chemical tests in parts per trillion can  be made,  trace
amounts of the pesticides or their  metabolites will at some stage be
found in water. Thus,  permissible levels in water  must be  estab-
lished if we are to continue use of  any pesticides. Such levels  might
be established only for drinking water, or they might include marine
waters, irrigation waters, or waters for swimming, boating, and fish-
ing.  It would be possible to consider overall environmental permis-
sible levels or  international  permissible levels. It would be  impos-
sible to set levels defined by (1) the lowest level of testing accuracy
and (2) the level at which it could be guaranteed that there was not,
nor ever would be, any deleterious effects of any kind.
    Aquatic  herbicides, to be effective, are  applied to water or im-
mediately adjacent to water. There are at present  very few  toler-
ances established for these aquatic herbicides in water.  Some  older
ones such as arsenic  do have  established levels. Tolerances  have
been established for food crops for the "newer"  herbicides but not

     DON C. PETERS is Professor, Department of Zoology  and Entomology,
     Iowa State University. H. B. PETTY is Professor and Extension Ento-
     mologist, University of Illinois.

                                 CHAPTER 15 / WORKSHOP SESSION / 225

for water. It is imperative that permissible levels be established soon.
     Chlorinated hydrocarbon insecticides present a different prob-
lem. Some people believe that permissible levels in water can be set,
others do not.  Although the hydrocarbons are occasionally applied
directly to water,  their appearance in water  usually results from a
nonwater use.  This can be runoff from actual use,  but it can  also
be manufacturing waste.  These insecticides are not water soluble
and escape  from it at every opportunity; thus they accumulate on
the  aquatic  plants and bottom  sediment where so many of  our
aquatic organisms live. The organisms concentrate these chemicals
in their bodies, sometimes to  thousands of times  the amount in
the water.
     All  chlorinated hydrocarbons  accumulate  in living  organisms
in varying  degrees.  Of the many, only a few present problems.
Dieldrin probably  persists  in the environment longer  than others,
although  DDT approaches it in  persistence.  Dieldrin  content in
fish is in relation to water  content, not food content. When fish are
fed excessive amounts of dieldrin there is a quick uptake, but the
body content returns to  the water-dieldrin eouilibrium within  a
month.  This is the opposite to DDT. Although DDT  and its metabo-
lites stay in the environment, DDT is apparently responsible for the
upset in the calcium metabolism in  some birds.  Endrin, with the
highest acute toxicity, does not persist as long as the other two  and
organisms cleanse themselves of endrin readily.  Endrin is, at the
moment, suspected of svstemic absorption.
     Although it is difficult to set permissible levels  for chlorinated
hydrocarbons for  all  situations,  levels  have  been set for drinking
water.  As a result of a committee of about 50  experts pooling their
knowledge, in Mav 1968 the Water Pollution Control Administration
published "Water Quality Criteria."
     It is possible  to set permissible levels for the organophosphates
since the amount which will produce cholinesterase inhibition can be
defined with some accuracy.  As much  as 10%  inhibition may be
permissible.  Furthermore,  the  residues of organic phosphates in
water are short-lived  and runoff from agricultural use, if found at
all, is present for  a very short time.  Manufacturing wastes rmy be
more important than use runoff.  However, one point to be considered
is the speed of reversibility of phosphate effect (comparativelv low)
and carbamate effect (comparatively  high) on  cholinesterase levels.
     Manufacture  of  these products provides  another  avenue for
contamination. There are examples where waste from pesticide man-
ufactrring plants  seriously contaminated miles of streams.  A  per-
missible level for these products in factory effluent must also be set.

     Knowledge of the metabolism of barbiturates 15 years ago has
been greatly enlarged upon and  even changed. The same thing is
happening with pesticides.  DDT is  not  a single compound but a
mixture  of  many materials.   Recently  in vivo isomerizations have
been authentically reported in feeding experiments in which o,p'-DDT


 was fed and p,p'-DDT was formed in the animal. Similarly, p,p'-DDT
 was fed and o,p'-DDT was formed in the animal. It has also been
 published in Science that o,p'-DDT has pronounced estrogenic effects.
     Reference  has been made to the absence or the quick disap-
 pearance  of organophosphates.  Do  we know  the  metabolites  of
 phorate,  Dyfonate,  carbofuran, etc.?  How  about amino parathion
 and its effects?  In short, we need to know more about the biological
 effects of pesticide metabolites than we do at present. A small amount
 of a toxic chemical can be tolerated by most organisms. A person can
 ingest a very low level of parathion and be unable to detect a reac-
 tion, except through very rapid sampling, as parathion is detoxified
 rather quickly.  But with an increase in this level of intake, one soon
 starts detecting it or some of its  metabolites.  A little more  and
 symptoms of intoxication are evident.
     Is it possible to determine the no-effect level for pesticides and
 their metabolites for the most important sensitive species of animals?
 It might be man, the peregrine falcon, or others, but we could estab-
 lish a base line. We will have to settle on the most toxic form of a
 given chemical  as well as the most  sensitive species.  In the case of
 DDT is it the p,p'-isomer, the o,p'-isomer, DDE,  or  possibly  DDD?
 What is the important sensitive species for which we can determine
 an environmental level for any given pesticide?  Is  this level con-
 sistent with its  use in agriculture  and public  health? Is this kind
 of approach a practical one? What is best for our human society?
 These were some of the unanswered questions  concerning pesticides
 and the quality of our environment.

     Pesticides protect plants and  animals from pest losses  to the
benefit of mankind.  This should be done so as not to harm man now
or in the future. However, no scientist ever could or would positively
guarantee that no harm could ever occur from the use of a certain
chemical.  To  answer every  question  that  could  be posed  would
require 30 years of search for answers,  and such  detail, if not scien-
tifically impossible, is financially impossible.
     Without pesticides, food  production  would be  reduced  some
40 to 50%, and the quality would  be greatly reduced. Bread as we
know it would still be present but would  contain  insect fragments
and  some rodent excrement.  Today food processors are on  the horns
of a dilemma—foods are inspected for both insect fragments and
pesticide,  and an entire day's pack  can be  confiscated if contamina-
tion  (by either) is found in any one  can  or case of processed food.
     In the past,  tolerances were established for  chemicals on food
crops. The safety factor was considered to be 100 to 1. That concept
is no longer valid.  We are now searching for minor or hidden effects.
We have searched for flaws in DDT for some 25  years and can still
find  a few weaknesses.  The Russians are  interested  in  carbamates
and  are diligently searching for hazards. We constantly search for

                                 CHAPTER 15 / WORKSHOP SESSION  / 227

metabolites, side effects, etc.  In the meantime,  we have  constant
pleas for help to control pests in order to enhance food  supplies.

     Little  can be learned about pesticide contamination until  the
materials are used in our environment.  Mock environments can be
assembled, but based only on  our past  experience with DDT and
other chlorinated  hydrocarbons.  Had it  not been for widespread
use of these materials it is doubtful that we would have been able to
foresee  and  prevent any  of  our present-day problems.   We can
theorize, but until we use a chemical and find it in streams,  for
example, we do not know the  actual environmental problems  in-
     As greater chemical detection finesse is attained we change  our
views about residues. The one  philosophy that might be acceptable
is the one  used by the USDA in clearing labels—if you can use an
insecticide  in  such a way as to avoid having a residue, then there
should be no permissible level.
     Coho salmon survival from eggs from Lake Michigan was lower
than from  eggs from Lake Superior.  It seems that more  information
is needed on this entire situation.

     We too often view insecticides as though a  single one will be
with us for a  lifetime.  Actually, DDT lasted about 10 to 15 years,
others a shorter  time.  The commercial  life of an insecticide  is  a
matter of years, not decades, so the time to find  the  answer to the
questions is  limited.  Resistance of insects to an insecticide can
develop rapidly and a product can be on the market and gone before
problems even arise.  Insecticides of the future  will have  a short
commercial life, not a long one.
     We have  alarmed people who now want to do something about
pollution, including pesticides, and we are in no  position to answer
all the questions  and supply the guidance needed.  We need much
more research which will cost taxpayers large sums of money if they
want answers  first.
     It was the hope of the group that public pressure was not dictat-
ing programs  and answers.  Science must be cold-blooded and give
answers based on fact, not emotion.  However, science does  dictate
its needs and we do respond to this.
     Overcaution  so far as our environment is concerned should be
the goal for pest control specialists, and we should not use pesticides
unless their use can be completely justified. On this basis, DDT and
other insecticides should not  be banned  from use but should be
usable at least on a permit basis.  With proper discretion in  use, it is
possible that no permit, ban, or other restrictive measures would be



        ISCHARGE of livestock and poultry manure into the environ-
ment is a practice as old as the animal. Historically, animal manure
was randomly  deposited on  the land surface where  the nutrients
were utilized by growing vegetation and the organic matter was in-
corporated into  the soil humus.   Current livestock manure produc-
tion, in excess of 1.5 billion tons  per year (Wadleigh,  1968), results
from a combination of the historical range  or pasture production
and some  degree  of confinement in  which traditional on-site soil
incorporation may not be applicable as a manure disposal system.
As much as 50% of the current manure production is  from confine-
ment production (Law and Bernard, 1969).

     The major water pollutants arising from animal manures are
oxygen-demanding  matter  (principally  organic matter),  plant nu-
trients, and infectious agents.  Color and odor are potential polluting
constituents of secondary importance.  Organic matter from livestock
wastes, like that from other sources, serves as a substrate for aerobic
bacteria when it enters a receiving stream. Associated with bacterial
metabolism is the utilization of dissolved oxygen.  When the rate of
oxygen utilization exceeds the reaeration rate of the stream, oxygen
depletion occurs.  Whenever sufficient organic matter enters, oxygen
concentrations will be reduced below the level necessary for fish sur-
vival, and in more severe cases, complete oxygen depletion will occur
and cause the development of anaerobic conditions.

     J. R.  MINER is  Assistant Professor, Department of Agricultural En-
     gineering, Iowa State University.  T.  L. WILLRICH is Professor, De-
     partment of Agricultural Engineering and Extension  Agricultural
     Engineer, Iowa State University.
     Journal Paper No. J-6378 of the Iowa Agriculture and Home Econo-
     mics Experiment Station, Ames.  Project No. 1730.  Prepared for pre-
     sentation to A  Conference Concerning the Role of  Agriculture in
     Clean Water, Ames, Iowa, November  18-20, 1969.


 TABLE 16.1.  Pollutional  characteristics of  untreated  animal wastes,  sum-
             mary of values.
Beef cow . . .
Dairy cow . .
Poultry ....
0 06
(Ib/day P2O5)
     Organic matter in wastewater has historically been measured as
 biochemical oxygen demand (BOD), This measurement evaluates the
 concentration of oxidizable organic material that can be utilized by
 aerobic bacteria in terms of how much oxygen they will require to
 metabolize  this material during  a specified time, generally 5  days,
 and at a specific temperature, generally 20° C. Having determined
 the BOD and knowing the quantity of waste produced, it is possible to
 determine a daily BOD production for  various animal species. The
 BOD of animal wastes has been  evaluated by numerous researchers
 (Jeffrey et al.,  1964;  Taiganides  et al.,  1964; Dornbush and Ander-
 son, 1965; Hart and Turner, 1965; Witzel et al., 1966; Dale and Day,
 1967; Jone^t al.,   1968).  From  these  data, representative  BOD
 quantities from various animals can be determined.  These values are
 summarized in Table 16.1.
     Chemical  oxygen demand (COD) is another measure of organic
 and other oxygen-demanding water based on  chemical rather than
 biological oxidation.  The COD exceeds the BOD of a waste due to the
 inability of aerobic bacteria to  completely oxidize the more resistant
 constituents under the conditions of the BOD test.  Table 16.2 com-
 pares the BOD and and  COD of  various wastes by using untreated
 municipal sewage as a reference.
     In addition to oxygen depletion and resulting changes in aquatic
 life, decomposing organic matter  contributes to color,  taste, and odor
 problems in public water  systems utilizing  surface  sources.   Such
 problems are often difficult to solve, yet  are of great significance. Re-
 duced inorganic substances, such as ammoniacal nitrogen, exert an
 oxygen demand in addition to organic matter.  Ammoniacal nitrogen
 exert an oxygen demand in addition to organic matter.  Ammoniacal
 nitrogen concentrations ranging from 1 to 139 mg/1 were foimd in
 feedlot runoff (Miner et al., 1966)  and from 197 to 332 mg/1 in swine
 manure lagoon effluent (Koelliker, 1969).
fAELg 16.2.  BOD and COD concentrations in various wastes.

domestic, sewaee 	
(rag /I)
Dairy cattle manure (Dale and Day, 1967)     25,600
Swine manure (Scheltinga,  1966)  	   27,000-33,000
Chicken droppings (Niles, 1967)	      24,000

                               CHAPTER 16 / LIVESTOCK OPERATIONS / 233

     Nitrogen and  phosphorus  are the  plant nutrients of primary
concern.  These elements are present in sufficient quantities  to in-
crease nutrient concentrations  in surface water  bodies  and  thus
stimulate the growth of aquatic plants.  In addition, nitrate toxicity
due to increased nitrogen  concentration in groundwater is important
in many rural areas.
     Livestock wastes are sources of infectious agents that may infect
other animals and,  in  some instances,  man. Among the  potential
water-borne diseases transmissible from animals are anthrax, brucel-
losis, coccidiosis, encephalitis, erysipelas, foot rot, histoplasmosis, hog
cholera, infectious bronchitis, mastitis, Newcastle disease, ornithosis,
gastroenteritis, and salmonellosis (Wadleigh, 1968).  Although water-
borne diseases are relatively rare in our country, increasing emphasis
on water-based recreation creates new opportunities for this mode of
infection. Leptospirosis has been spread from cattle to swimmers by
the water-borne route (Diesch  and McCulloch,  1966).  Samples  of
cattle feedlot runoff, as small as one ml,  showed the  presence  of
Salmonella organisms even though there were no symptoms of infec-
tion  observed in the cattle (Miner et  al., 1967).  By using  the  fecal-
coliform—fecal-streptococcus ratio (Kenner et al., 1960) it is possible
to distinguish between livestock  and human wastes. When  stored in
a lagoon or applied to the  soil, pollutional bacteria—coliform and en-
terococcal—die off rapidly (McCoy, 1967). Thus, little public health
hazard would appear due to lagooned livestock wastes.  It was further
noted that for bovine wastes the predominant enterococci were  Strep-
tococcus durans and S. faecium rather than S. faecalis found in the
human intestine.  This suggests a different interpretation of entero-
coccal counts for animal than for human waste sources.
    Since livestock wastes are not usually collected, transported, treat-
ed, and discharged into a receiving stream, as municipal sewage al-
most always  is, a quantified prediction of water-quality deterioration
caused by animal wastes cannot be  made as it can for municipal
sewage.  Calculation of a  population equivalent for the wastes from
various animals assumes that the total wastes from  these animals are
discharged into streams and released at a uniform rate either with or
without treatment.  Neither assumption is valid except in a most un-
usual situation.
     However, the potential for livestock wastes to pollute water is
influenced by the ways in which it is collected, stored, and treated as
well  as the final method of disposal. Seven major potential pollution
sources exist in connection with livestock wastes.

     Runoff from Range and Pasture Operations

     Where animals graze a vegetated land area (range or pasture),
little interest  has been shown by  water pollution control  agencies.
Manure is uniformly distributed in  a light application, liquids are ab-
sorbed by the soil,  and the  vegetative cover  utilizes  the added nu-
trients and  inhibits erosion. Low-intensity rainfalls  are usually ab-


 sorbed by the soil and high-intensity rainfalls in excess of soil infiltra-
 tion rates provide sufficient dilution water to minimize the concentra-
 tion of potential pollutants in the runoff.
     In range and pasture systems, one can visualize extensive waste
 treatment taking place as any runoff-carried pollutants pass over the
 soil surface.  Vegetative cover provides effective screening as well as
 settling areas for particulate matter.  Mixing and aeration stimulate
 biological treatment of soluble organic matter.  Thus, with respect to
 water pollution potential, range or pasture livestock  production is of
 less concern than confinement production.  However, when  one  con-
 siders the use of a farm pond as a domestic water source, utilization
 of the watershed as a pasture is discouraged  because  of the high-
 quality water requirements and the relatively long die-off periods ex-
 hibited by  pathogenic organisms in  such a system (Andre et al.,
     Runoff from  Cropland following Manure Application

     When manure is spread on frozen or snow-covered fields, or
 when  heavy rainfall occurs immediately following manure applica-
 tion, considerable runoff and a resulting organic matter and nutrient
 loss is possible.  Data from Wisconsin indicate  that spring applica-
 tion of manure  caused no increase in loss of nitrogen in runoff.
 Manure application on snow-covered ground that was  followed by a
 rain increased nitrogen losses from  a normal 3 to 4 pounds per acre
 annually to over  23 pounds (Hensler et al., 1969).  Additional runoff
 losses  are  possible where manure is stockpiled prior to spreading in
 such a way that runoff has direct access to a surface stream.
     Runoff from  Feedlots and Similar  Unroofed Enclosures

     Animals  produced in feedlots, pens, and other  uncovered en-
closures in such a concentration as  to remove the vegetative cover
present pollution  hazards unlike the  pasture  systems.  During  and
immediately after rain and spring thaws, water flows over manure-
covered feeding areas and carries both particulate and soluble manure
components with it. This pollution source has received  considerable
public interest due to the occurrence  of dramatic fish  kills and other
gross pollution incidents. The action of animal hooves on a feeding
surface creates an area void of vegetation and one through which in-
filtration rates are greatly reduced.  However,  considerable surface
storage capacity is available on feeding areas in the hoof depressions.

     Data exist on the quality of runoff from cattle feedlots (Smith
and Miner, 1964; Miner et al., 1966; Loehr, 1968). They indicate cat-
tle feedlot runoff to be of highly variable quality,  depending upon

                               CHAPTER 16 / LIVESTOCK OPERATIONS / 235

such  factors as rainfall intensity, temperature and feedlot surface
moisture  content, and manure accumulation.  Organic  content  as
COD  in cattle feedlot runoff ranged from 3 to 11 times the COD in
untreated domestic  sewage  (Miner et  al.,  1966).   Although runoff
from  feeding areas confining animals other than cattle may be ex-
pected to be high in organic matter, no data are currently  available
concerning  these sources. In addition to the high-strength character-
istics of feedlot runoff, the slug effect upon a receiving stream is par-
ticularly damaging.  When feedlot runoff is uncontrolled, particularly
from  a lot located adjacent to receiving streams, the large volume of
relatively high-strength  wastewater enters  the stream quickly and
consequently allows little time for dilution by runoff from clean areas.
Thus, one technique proposed for the reduction of feedlot runoff dam-
age is the construction of flow control structures that spread the dis-
charge of runoff over a longer time period.  Of particular concern to
pollution agencies have been  large feedlots (capacity over 1,000 head),
lots located  near or adjacent  to streams, or lakes and lots whose run-
off enters groundwater supplies through abandoned wells,  springs,
sinkholes, or other openings.
     In assessing the significance of cattle feedlot runoff compared
with  other waste sources within a  drainage basin,  one must look at
both  the quantity and quality of runoff.  Assuming an  earthen lot
with a 2% slope, about 11 inches of annual runoff might be expected
from  30  inches of annual rainfall, with runoff occurring during 30
days  of the  year.  At an average of 1,000  mg/1 of  BOD, the runoff
from  a feedlot on each of these 30 days would be equivalent to the un-
treated sewage from a community of 500 people per acre of feedlot
surface.  Although such an  average is  of little help in actual situa-
tions, it  indicates that runoff from  cattle  feedlots is a significant
source of organic wastes, but it  is not of the same magnitude as one
gets if he bases his predictions on standard population equivalents for
various livestock.

     In response to fish kills attributable to feedlot drainage (Loehr,
1968) and for other reasons, such as stream enrichment, various pol-
lution control measures have been devised.  The first step in most pro-
grams is to divert any water falling outside the feedlot so that it will
not flow across the feedlot and thereby minimize the quantity of pol-
luted runoff. The second step is generally  the construction of a run-
off collection and impoundment system that will prevent  the im-
mediate and uncontrolled entry of runoff into a stream.  Facilities to
settle manure solids are frequently incorporated into either the run-
off collection system by the design of channels for low flow velocities
or by the construction of separate settling basins.  Settling facilities
are designed for flow velocities of 1 foot per second  or  less and for
dewatering so collected solids  will  dry  more rapidly  and thus more
easily.  Where solids are to be  removed from a  settling basin with a
dragline, a maximum basin width of 50 feet is desirable.


      Runoff impoundment basins generally provide sufficient capacity
 to hold 3 to 6 inches of runoff from the contributing area.  The final
 design capacity is a function of the climatological  features of the
 area and the proposed method for disposing of collected runoff.  In
 some parts  of the country where seasonal and annual evaporation
 losses sufficiently exceed rainfall quantities,  it is possible  to design
 runoff impoundment basins so that most or  all collected water will
 be lost by evaporation and seepage. This approach is not applicable
 in humid regions, however.
      Where evaporation and seepage losses are not sufficient for run-
 off disposal, collected wastewater may be spread  on land or treated
 prior to release into a stream or surface water body.  Problems as-
 sociated with wastewater treatment are (1)  the necessity of frequent
 operator  attention, (2) the difficulty in producing a  high-quality ef-
 fluent, and (3) the costs involved in such treatment.
     Discharge from  Waste Storage or Treatment  Units

     Roofed livestock confinement units offer advantages to the pro-
 ducer in ease of mechanizing feed and water distribution and manure
 collection as well as offering the possibility of environmental control.
 Such units range from unheated structures with natural ventilation
 to totally enclosed  buildings with mechanical ventilation as well as
 heating and cooling equipment.
     To perform satisfactorily, an enclosed livestock building must in-
 corporate a compatible manure management system. A manure man-
 agement system may logically incorporate (1) a means to separate the
 manure from the animal and to collect it in some logical place, (2) a
 method to transport it, (3) a storage device, (4) one or more treatment
 units, and (5) a final disposal or utilization scheme. These functions
 must be mutually compatible as well as being compatible with  the
 remainder of  the production unit.  They  must not only control  the
 escape of potential water pollutants but also minimize  the potential
 for odor, insect, and rodent nuisances, and operate  with a minimum
 of labor, capital investment, and operating costs.
     Totally roofed animal units eliminate the open-lot runoff prob-
 lem but they offer the greatest potential for water pollution of  all the
 livestock production schemes.  They also offer the greatest potential
 for essentially pollution-free operation.  System design and manage-
 ment determine the degree of pollution that will develop, if  any.  As
 an example, a 1,000-head beef unit would be equivalent to a com-
 munity of 6,000 people, based  on BOD, if the  raw  wastes were
dumped into a stream every day, or a community of up to 600,000 if
 the accumulated wastes were dumped every 100 days.  However, with
proper waste collection, transport,  and  application to cropland,  the
manure from this operation need not contribute to water pollution.
     Liquid manure systems are most common in roofed confinement
units. Liquid manure may be applied to the soil with or without treat-
ment as just discussed. Treatment for release into high-qualitv sur-
face waters has not been recommended due to  the inability of cur-

                               CHAPTER 16 / LIVESTOCK OPERATIONS / 237

 rently available systems to produce an acceptable effluent at a rea-
 sonable cost.

     Percolate  from Feedlots  and Similar  Unroofed Enclosures

     Whenever water passes through a layer of manure  and perco-
 lates into the underlying soil, it will carry certain components of the
 manure with it. Because of soil puddling and compaction by animal
 hooves, however, the infiltration rate in an animal feeding area will
 usually be low. Thus, only a  very small quantity of water would be
 expected to enter the groundwater supply as long as the lot is in con-
 tinuous use to confine animals. Where soil and groundwater samples
 have been collected near old feedlcts, elevated  nitrate-nitrogen  con-
 centrations have been detected (Smith,  1967).  Data collected from
 beneath feedlots and irrigated fields of the South Platte Valley in
 Colorado also indicated elevated nitrogen concentrations  in ground-
 water near feedlots (Stewart et al., 1968).  They also noted high or-
 ganic carbon concentrations in groundwater samples as much as 35
 feet beneath feedlots.  High organic carbon  concentrations caused
 much of the nitrogen to be present as ammonium nitrogen. Thus, lo-
 calized  pollution of the water-table  aquifer with nitrogen near and
 under animal feeding  areas does  take  place.  However,  due  to  the
 limited  acreage being used for feedlots,  widespread  groundwater
 pollution due to infiltration from animal feeding areas is not likely.
     Percolate from  Disposal  Areas

     Most animal manure is spread on cropland.  This includes not
only manure and other wastes scrapped from open feedlots but also
that hauled, both solid and liquid wastes, from confinement buildings
and barns as well. This manure is field spread not only because of its
fertilizer value but also as a convenient and least-cost disposal tech-
nique in most  situations. Current manure-spreading techniques in-
clude not only  conventional solid  manure spreaders but also liquid-
hauling tanks and irrigation equipment.  Two potential modes of pol-
lution exist for manure applied to cropland:  (1) runoff due to rainfall
or snowmelt  carrying it to surface streams or impoundments and (2)
percolation into the groundwater.
    Where collected feedlot runoff or liquid manure is spread on crop-
land, forest land, or pasture, the greater portion of pollutants will be
removed from  the  wastewater before it becomes a  portion of the
groundwater recharge.  Soil has the ability  to remove  all suspended
solids and much of the dissolved  material.  BOD and  COD removal
should present  no problem as long as the infiltration capacity of the
soil is maintained. Soil also has the ability to absorb large quantities
of phosphorus.  Nitrogen, however, can escape to the groundwater
and thus sufficiently increase the nitrate concentration in a localized


 area so that the groundwater would be of inferior  quality for some
     Recent work  with application of anaerobic animal wastes to
 grassland  indicates that with proper management extensive biologi-
 cal  denitrification  is possible (Koelliker  and Miner, 1969). In  one
 trial, using anaerobic lagoon effluent, 2,300 pounds of nitrogen per
 acre were  applied in 30 inches of lagoon effluent. Losses to ground-
 water (250 Ib/A) and in runoff (170 Ib/A) were 420  pounds per acre.
 A net  nitrogen loss within the soil profile  of 400 pounds per acre was
 measured.  Thus, a loss of 2,020 pounds  of nitrogen per  acre due to
 denitrification took place during the 3-month trial period.
     Percolate from Field-spread  Manure

     Groundwater pollution due to field-spread manure has generally
 been of b'ttle significance, due to the associated organic matter which
 tends  to release nitrogen over an extended  time  period and due to
 the  conventional rates  of manure application.  This mechanism al-
 lows the nutrients greater opportunity to be used by crops or be in-
 corporated into the soil.  The  soil is also effective  in removing po-
 tentially infectious bacteria; 14 inches of silt loam soil removed the
 initial concentrations  of  1 X  105/rnl of  Escherichia coli  and  of
 1 X  10G to 1 X 107/ml of enterococci (McCoy, 1969).

 1.   Potential water pollutants from animal manures are oxygen-de-
     manding matter, plant  nutrients,  infectious  agents, and color-
     and odor-contributing substances.
 2.   Total solids in animal manures are about 300  times more con-
     centrated  than in municipal  sewage.  The BOD  of  undiluted
     animal manures  is about 100 times greater than the BOD of
     municipal sewage.
 3.   Ammoniacal nitrogen concentrations in diluted  and decomposed
     animal wastes, such as lot runoff and lagoon effluent, are suffi-
     cient to exert a major oxygen demand or produce a toxic level
     to fish in a receiving stream.
 4.   The incidence cf  water-borne diseases transmitted from animal
     to man is low even though a dozen or more diseases can be trans-
     mitted by this route.  Fecal  enterococcal counts must be inter-
     preted differently  for animal-manure-polluted water than for hu-
     man-waste-polluted  water since nonpathogenic enterococci  ap-
     parently predominate in some animal  wastes.  Most infectious
     agents die off rapidly when animal wastes are treated or applied
     to the soil.
5.   Data concerning  pollutants  removed by runoff from livestock
     range and pasture operations  are  sparse. Logic indicates that
     this potential source is relatively insignificant  when compared
     to other sources.

                              CHAPTER 16 / LIVESTOCK OPERATIONS / 239

6.   Runoff from manured cropland will transport greater quantities
     of pollutants if the manure has been spread on frozen or snow-
     covered fields.
7.   Highly concentrated  open feeding areas offer the potential  for
     runoff-caused pollution problems, due to the low infiltration rates
     and high manure density.  Runoff control is one key to pollution
     prevention. Manure  cleaned from lots and collected runoff re-
     quires some means of disposal. Land application is the current
     disposal means of preference.
8.   Roofed confinement  livestock  buildings make possible  a high
     degree of control over manure disposal. A proper means for con-
     trol of this material requires systems  for  manure collection,
     transport, storage, treatment, and/or disposal or utilization. Hy-
     draulic manure transport  systems offer improvements in  labor
     requirements but unless some means of water reuse is planned,
     excessive waste disposal expense is encountered. Improper ma-
     nure disposal  from  such  a unit  causes the  greatest pollution
     threat of the systems mentioned.
9.   The application of livestock manure to the  soil is both a logical
     and historically verified practice. Technological, social, and eco-
     nomic factors  have in recent years made  this practice less  ac-
     ceptable.  Applied in proper quantities with alert management,
     and with improved methods of application, manure disposal by
     return to the soil should be encouraged. This disposal may  ne-
     cessitate  treatment and conditioning prior to  disposal to  mini-
     mize odors or water pollution.

Andre, D. A., Weiser, H. H., and Maloney, G. W.  1967.  Survival of
     bacterial pathogens in farm pond water.  /. Am. Water Works
     Assoc. 59(4): 503.
Dale, A. C., and Day, D. L. 1967.  Some aerobic decomposition prop-
     erties of dairy-cattle manure.  Trans.  Am.  Soc. Agr. Engrs.
     10 (4): 546-48.
Diesch, S. L., and McCulloch, W. F.  1966. Isolation  of pathogenic
     leptospires from waters used for recreation. Public Health Kept.
     81 (4): 299-304.
Dornbush J. N., and Anderson, J. R.  1965.  Lagooning  of livestock
     wastes in South Dakota.  Proc. 1964 Ind. Waste Conf. Lafayette,
     Ind.: Purdue Univ. Eng. Ext. Ser.  117, pp. 317-25.
Hart, S. A., and Turner, M. E.  1965.  Lagoons  for livestock manure.
     J. Water Pollution Control  Federation 37(11): 1578-96.
Hensler R F.  Olsen, R. J., Witzel, S. A., Attol, O. J., Paulson, W. H.,
     and Johannes, R. F.  1969. Effect of method of manure han-
     dlino- on crop yields, nutrient recovery and runoff  losses.  Pre-
     sented at meeting of Am.  Soc.  Agr.  Engrs., 22-25  June  1969,
     W. Lafayette, Ind.
Jeffrey  E A.  Blackman, W. C., and Ricketts,  R. L  1964  Aerobic
     and  anaerobic  digestion  characteristics  of  livestock  wastes.
     Univ. of Mo. Eng. Ser. Bull.  57.


 Jones, D. D., Jones, B. A.,  and Day, D. L.  1968, Aerobic digestion of
     cattle wastes.  III. Res. 10 (2): 16-18.
 Kenner, B.  A., Clark, H. F.,  and Kablet, P. W.  1960. Fecal strepto-
     cocci:   quantification of streptococci  in feces.  Am. J. Public
     Health 50 (10): 1553-59.
 Koelliker, J. K.  1969.  Soil  percolation as a renovation means for
     livestock  lagoon  effluent.  Unpublished Master's  thesis, Iowa
     State Univ., Ames.
 Koelliker, J. K., and Miner, J. R.  1969. Use of soil to treat anaerobic
     lagoon effluent renovation  as a function of depth and  applica-
     tion rate. Paper 69-460 presented at meeting of Am. Soc. Agr.
     Engrs., 22-25 June 1969, W. Lafayette, Ind.
 Law, J. B.,  and Bernard, H.  1969. The impact of  agricultural pollut-
     ants on subsequent users.  Paper 69-235 presented  at  meeting
     of Am. Soc. Agr. Engrs., 22-25 June  1969,  W. Lafayette, Ind.
 Loehr, R. C.  1968. Pollution implications of animal ivastes—a  for-
     ward oriented re-view. U.S. Dept. of Interior, Fed. Water Pollu-
     tion Control Admin., Robert S. Kerr Water  Res.  Center, Ada,
 McCoy,  E.   1967.  Lagooning of liquid manure (bovine):  bacterio-
     logical aspects. Trans. Am. Soc. Agr. Engrs. 10 (6): 748-87.
 	.  1969. Removal of  pollution bacteria from animal wastes by
     soil percolation.  Paper 69-430 presented at meeting of Am. Soc.
     Agr. Engrs., 22-25 June 1969, W. Lafayette, Ind.
 Miner, J. R., Lipper, R. I.,  Fina,  L. R.,  and Funk, J. W.  1966. Cattle
     feedlot  runoff:  its nature and variation. /. Water Pollution
     Control Federation 48 (10):  1582-91.
 Miner, J. R., Fina, L. R., and Piatt, C.   1967.  Salmonella infantis in
     cattle feedlot runoff.   /. Appl. Microbiol. 15 (3): 627-28.
 Niles, C. F.   1967.  Egglaying house wastes. Proc. 22nd  Ind. Waste
     Conf. Lafayette, Ind.: Purdue Univ. Eng. Ext. Serv. 129, p. 334.
 Scheltinga,  H. M. J.  1966.  Aerobic purification  of farm waste. /.
     Proc. Inst. Sewage Purification, pp. 585—88.
 Smith, G. E. 1967.  Fertilizer nutrients  as contaminants in water
     supplies.  In Agriculture and the quality of our environment,
     ed. N. C. Brady, pp. 173-86.  Norwood, Mass.:  Plimpton Press.
 Smith, S. M., and Miner, J. R.  1964.  Stream pollution from feedlot
     runoff.   Trans. 14th Ann.  Conf. Sanit. Eng.,  pp. 18-25. Univ.
     of Kans., Lawrence.
 Stewart, B. A., Viets, F. G., and Hutchinson, G. L.   1968. Agriculture's
     effect on nitrate pollution. /. Soil Water Conserv. 23 (13): 13-15.
Taiganides,  E. P., Hazen, T. E., Baumann, E. R., and Johnson, H. P.
     1964.   Properties and pumping  characteristics of hog waste.
     Trans. Am. Soc. Agr. Engrs. 7 (2): 123-29.
Wadleiffh. C. H. 1968. Wastes in relation to agriculture and forestry.
     USD A Misc. Publ. 1065.
Witzel, S. A., McCoy, E., Polkowski,  L. B., Attoe, O. J.,  and Nichols,
     M.  S.   1966.  Phvsical, chemical  and bacteriological properties
     of  bovine animals.   In  Management of farm animal  ivastes.
     St.  Joseph, Mich.:  Am. Soc.  Agr. Engrs.  SP-Oe66,  pp.  10-14.

     IN  SOIL

        UGE quantities of animal waste are accumulating in small
areas because of the increasing confinement  of  animals in large
numbers for meat, milk, and egg production (Wadleigh, 1968). The
production of enormous amounts of urine and fecal material has
caused unparalleled disposal problems and a threat to water  quality
(Commoner, 1968). The best way to dispose of animal waste is to
put it on the land for decomposition and mineralization. But what is
the highest concentration of animal waste  can be applied  to  the
land without upsetting favorable microbial decomposition patterns,
producing  a toxic effect  on the crop,  or polluting the runoff and
     There are some waste treatments that can be  applied to  animal
manure to remove its high oxygen demand and inorganic nutrients,
but these treatments are not yet economically feasible.
     Much can be done in the management of the animal waste on
site (for example, on beef cattle feedlots)  to create a favorable en-
vironment for decomposition  so that a considerable amount of the
manure can be decomposed to COo and N2, which will dissipate into
the atmosphere  (Dale and Day, "1967; McCalla and Viets,  1970).
Phosphates are readily adsorbed by the soil and thus may be removed
effectively  from  solution.  Therefore,  correct management can re-
duce eutrophication1 in streams and lakes.
     T. M. McCALLA is Microbiologist, USDA, Lincoln, Nebraska. L. R.
     FREDERICK is Professor, Department of Agronomy, Iowa State Uni-
     versity. G. L. PALMER is Instructor, Department  of Agronomy, Iowa
     State University.
     Contribution from the Northern Plains Branch, Soil  and Water Con-
     servation  Research  Division, ARS, USDA, in cooperation with the
     Nebraska Agricultural Experiment Station, and from the Agronomy
     Department, Iowa State University, Ames. Published as  Paper No.
     2742, Journal Series, Nebraska Agricultural Experiment Station; and
     Paper No. J6431, Iowa Agricultural Experiment Station, Project 1378.
     1.  Eutrophication is an excessive enrichment of water  with nutri-
     ents, such as nitrates and phosphates, which will promote a luxuri-
     ant growth of algae (algal bloom).


 TABLE 17.1.  Chemical  composition  of various  fresh  manures,  litter  free.
Chemical Constituents
Ether-soluble substances 	
Cold water-soluble organic matter .
Hot water-soluble organic matter . ,
Total nitrogen 	

Sheep Horse
Manure* Manure
(percent of dry
2.83 1.89
, 19.19 3.19
5.73 2.39
18 46 23.52
18.72 27.46
20.68 14.23
4.08 1.09
. 17.21 9.11

t Manure*

 Source:  Waksman (1938).
 * Solid  and liquid excreta.
 t Solid excreta only.


     Fresh manure contains from 30 to 85% water.  The rest of the
 constituents in manure are inorganic and organic solids, liquids, and
 gases.  The composition of manures is  shown in Tables 17.1, 17.2,
 17.3, and 17.4.
     Manure contains all  the inorganic nutrients needed  by plants.
 These nutrients are worth  slightly more than $1 per ton (Table 17.4).
 When putting large quantities of manure on land, materials such as
 ammonia may accumulate in concentrations toxic for  the growth of
 plants (McCalla and Haskins, 1964; Megie et al., 1967). Using average
 figures  for production of manure per animal unit and agricultural
 statistics for the number  of animals present  in  the various  states,
 an  estimate of the N, P, and K in manure produced by livestock in
 the  north-central  region of the  United States was made.  For the
 western north-central states, the manure contained per year 2,100,000
 tons N, 300,000 tons P, and 1,300,000 tons K. Similar figures were
 obtained for the eastern north-central states.  These figures  are ap-
 proximately  comparable  to the nutrients applied as fertilizer in
 1968, except that about 50% more phosphorus was applied as fertil-
     Roughly, 90%  of  the dry matter in  manure is  organic waste
 material from  animal  digestion of  feeds.  Animal  rations  consist
 largely  of carbohydrates (sugars, starches, celluloses, and hemicellu-
 loses), some  proteins, fats, small amounts  of lignin, and  numerous
 inorganic nutrients,  such as nitrogen, phosphorus, potassium, and a
 number of micronutrients  (Hemingway,   1961;  Gilbertson  et al.,
 1970).  In a high-concentrate ration, about 70 to 80%  of the organic
 nutrients are utilized by  the animal.  The substances used  by the
 animal  are mostly carbohydrates, some proteins, small amounts of
minerals, and other  substances. The animal waste is  more concen-
 trated than the feed in lignin and minerals upon  deposition in feed-
lot  or confinement structure and is less  concentrated in carbohy-
drates.  But the manure retains about 60 to 75% digestible materials.
Some fats are  present, and also humiclike substances resistant to

TABLE  17.2.   Characteristics of animal manures.

Dairy cattle 	
Fattening cattle • •
HOGS 	 ,
Horses 	 ,

, . . 79
, . . . 80
. . . . 75
. . . . 60
. . 65

11 2

2 0

10 0

1 0

Ca Fe
(Ib/ton manure)
56 0 08
2.4 0.08
11.4 0.56
15.7 0.27
11.7 0.32

2 2



 Source:  Loehr (1968).

 TABLE  17.3.  Trace element content of manures (as ppm, dry-matter  basis).
 Source:  Atkinson et al. (1954).
 Note: Data from 44 samples of farmyard manure, representing fresh cow,
 horse, swine, sheep, poultry, and mixed manures, and composted cow and
 mixed manures.
 * With one exceptionally high value omitted.
 TABLE 17.4.   Chemical analysis of slurry manure (a mixture of feces  and
              urine) from confined beef cattle feeding in Nebraska.
Volatile solids
Total solids
Wet Weight
85.0 %
11.6 %
0 %
121,000 mg. (Vliter
4.5 mmhos/cm2
Each Ton
5.8 Ib.
3.6 Ib.
6.2 Ib.
116 Ib.
72 Ib.
124 Ib.
Note:  Acknowledgment is made of the assistance of J. R. Ellis, USDA-ARS-
SWC, in making these determinations.
              TABLE 17.5.   Particle size analysis of fresh ma-
                           nure (oven-dry weight basis).
                 Particle Size
               Percent of Total
             4.00 mm or greater
             4.00 mm but >2,000ju
             Source:  Unpublished data, T. M. McCalla and
             J. S. Boyce (1969).

                       CHAPTER 17 / MANURE DECOMPOSITION IN SOIL / 245

 decomposition (Jansson, 1960a, 1960b; Alexander, 1961). Antibiotics
 may also occur in the animal waste (Morrison et al., 1969).
     The mechanical size  of  the particles in manure is  shown  in
 Table  17.5.  The solids consist of undigested fragments of grain,
 bran, fibrous materials, and about 30% colloidal  materials.
     The microbial population of animal waste is composed mainly
 of bacteria, fungi,  actinomycetes, and protozoa.  Cells of microbes
 and cells from  the lining  of the  intestinal tract of the animal  in
 feces amount to about 40%  of the feces (Crampton and Harris,
 1969).  Among the bacteria,  the enterococci and  coliforms are very
 numerous, with coliform counts as high as 18 billion excreted per
 animal per day (Table 17.6).
     Fresh manure, a manure-soil-urine  mixture from next  to the
 concrete feeding apron, and dry manure from the  middle of the feed-
 lot were collected in eastern Nebraska. A manure suspension, 5%
 by weight (oven-dry basis), was made by shaking manure and dis-
 tilled water for  1 hour. After standing for 0,  1, and 24 hours, both
 solubility of substances and suspension of the manure, and the num-
 ber  of microorganisms  in  the supernatant, were  determined  (Table
 17.7).   Highest  solubility  and suspension  of  combustible  material
 were found  in fresh manure,  and greatest solubility and suspension
 of noncombustible  material were  found in samples collected  next
 to the  feedlot bunkers.  Appreciable numbers  of microbial pollution
 indicators were present, and they remained in  suspension even after
 24 hours of settling. The amount of phosphorus  and nitrate in sus-
 pension and solution remained high after settling. Concentration  of
 total P  of material in  suspension and solution was approximately
 68 to 113 ppm and for NO:fN was 8.5 to 23 ppm.  The pH  decreased
 sharply when  the fresh manure suspension was allowed to stand for
 24 hours. Thirty percent of the manure  was in particle  sizes less
 than 2/ji.  Salter and Schollenberger  (1939)  found up  to  50%  of
 manure was humus. Shigella and Salmonella were not found in the
 manure samples. The orders  of magnitude of the microbial counts
 in the manure suspension  were: total count, 10s; anaerobes,  105  to
 10C; Escherichia coli, 105;  enterococci, 104 to 10°;  and total fungi,
 103 to 105 per ml of manure suspension.

     Decomposition in  Storage

     Animal waste may remain where deposited in feedlots and con-
finement buildings for considerable time before disposal, and much
decomposition may occur.  For  example, manure from beef cattle
with a high-roughage diet was incubated in a growth chamber simu-
lating spring and summer climatic conditions at Lincoln, Nebraska.
Urine was added twice weekly to the incubated manure samples to
equal two stocking rates: 50  and 250 ft2 per animal. After 3 weeks
of decomposition, 90% of the nitrogen added initially in the manure

TABLE 17.6.  Estimated daily per  capita discharge of coliforms in animal feces.


Moisture content (%) 	
Average weight of 24-hr fecal
discharge (wet weight in
grams) 	 ,
Coliforms per gram (millions) . . . . ,
Total coliforms discharged
per day (millions) 	 	

. . 77.0
. . . . 150.0
, . . . 13.0
. . . . 1,950.0







Source: Geldreich et al. (1962).

  TABLE 17.7.   Numbers of microorganisms and chemical tests on 5%  suspension of manure  in distilled water.  Numbers pere
                ml or mg/ml or ppm of manure suspension.
Kind of
Sample H,O
Do 95
Mo 95
Fo 95


per ml
per ml
per ml
E. coli
per ml
per ml
res. wt.
res. wt.
In suspension
(XlOr') (xlO7') (xlO*) (mg/ml)

(ppm for solution)
   Source:  Unpublished data, T. M. McCalla and J. R. Ellis, (1968).
    * = combustible at 550° C.
    t = numbers below or above dilutions made.
    D — dry manure
    M = mixture of soil, manure, and urine near feed bunker
    F = fresh manure
    0 = 0 times of standing after shaking
    1 = 1 hour of standing after shaking
    24 = 24  hours of standing after shaking


 or subsequently in the urine was lost into the atmosphere with  the
 stocking rate of 50 ft2  per  animal.  In  the  decomposing manure,
 NH3 concentrations were high, pH ranged from about 8 to 9, nitrates
 accumulated only to a slight extent, COD  values remained high, and
 salt concentration increased.  About 50%  of volatile solids were lost
 in 4 months (McCalla et al., 1969b).  In Connecticut, 3 bushels of
 fresh manure from dropping pits lost 55% of the organic matter and
 77% of the N when stored for 20 weeks in a laying house (Perkins
 et al., 1964). With the loss of carbon and nitrogen, mineral content
 increased, readily available organic materials decreased, and resistant
 materials such as lignin accumulated (Burnett and  Dondero,  1969).
 Manure oxygen demand is characterized as BOD2 and COD3  values
 (McCalla et al., 1969b).
      The BOD:COD ratio generally is about 8.5:10 for beef cattle
 manure (Lipper, 1969).  Morrison et al.  (1969) showed that excreted
 chlortetracycline in beef cattle feedlot waste, arising from antibiotic
 supplementation of the ration, had a half-life of 1 week at 37° C and
 greater than  20 days  at 28° C and 4° C.  By altering decomposition
 patterns, antibiotics or other chemicals may  affect release of nui-
 sance odors.
     The microorganisms found in manure during decomposition  are
 bacteria, fungi,  actinomycetes, and  protozoa  (Witzel et  al.,  1966;
 McCoy,  1967, 1969).  Many  of the E.  coli, enterococci,  and other
 intestinal  and  disease microorganisms are short-lived in the  soil
 (King, 1957; Burroughs, 1967; Klein and Casida, 1967).  In a manure
 decomposition study at Nebraska, the  E. coli  and  enterococci  dis-
 appeared rapidly, and none  remained  after the second  and third
 months. Fungi numbers were very low initially, but increased during
 the incubation period.  Bacilli decreased;  total bacteria increased
 (McCalla et al., 1969b).
     Decomposition in Soil

     The  addition of large amounts of manure  will stimulate the
growth of saprophytic bacteria, fungi, and actinomycetes in the soil.
Aerobic, mesophilic  bacteria metabolizing cellulose are much  more
numerous in manured fields.  Protozoan and actinomycete numbers
and COo  production are increased by manure additions (Alexander,
     Manure from a high-concentrate  ration  contains about 10 to
15%  lignin. Most of the other energy material  decomposes rather
rapidly in the soil.  Polysaccharides, including cellulose and starch,
and most protein materials decompose rapidly,  although some of
the proteinaceous material, probably associated with lignin or  kera-
tin, is fairly resistant to decomposition (Polheim, 1965).   Consider-

     2. Biological oxygen demand is the oxygen consumed by microbes in
     the process of oxidizing the organic materials during a 5-day incuba-
     tion  period.  This is basically an indication of the readily oxidizable
     material present.
     3. Chemical oxygen demand is a chemical evaluation  of the total
     oxidizable material  using sulfuric acid  and potassium  dichromate,
     the measure being the quantity of oxygen used in this process.

                        CHAPTER 17 / MANURE DECOMPOSITION IN SOIL / 749
                                 Unlobe led carbon,  mean of all
                Labeled carbon, mean of all
                            Period of incubation, years
      FIG. 17.1.   Losses  of  unlabeled (soil) and  labeled (ryegrass)  carbon
      from soils  incubated in  the field with labeled  ryegrass.  (Jenkinson,

 able carbon and nitrogen  are found in microbial cells formed during
 decomposition.   Of labeled C  added as 0.6% ryegrass to Broadbalk
 field soils, Jenkinson (1966) found that 30 to 33% remained in the
 soil after 1 year,  of which about  one-third was in microbial cells.
 After 4 years, 19% of the labeled C still remained, and only about
 19%  of  that was in  microbial  cells.  The carbon turnover rate
 appeared to vary with stage of decomposition:  the original residues
 decomposed rapidly with  a half-life estimated to be 14  to  30 days.
 After the first year, the biomass has a half-life of 1 year, and the
 residual C has a half-life of about 4 years, while the soil humus has
 a half-life of  about 25 years (Figure 17.1).  About 50 to  60% of the
 nitrogen in manure applied to soil will  be mineralized the first year.
     Factors  Affecting Decomposition
     Under aerobic conditions,  carbonaceous  materials  are rapidly
oxidized to CO2; microbial  cells are synthesized;  and nitrates,  sul-
fates, and inorganic phosphate  tend to  accumulate.  Manure added
in large quantities to the soil has  a tremendous O2 demand.  Well-

TABLE  17.8.  Generalized  presentation  of breakdown products of manure decomposed under aerobic and anaerobic conditions.
Type of

Carbon Nitrogen Phosphorus Sulfur
corn- com- com- com-
pounds pounds pounds pounds

CO,, CH,

N2, NHa
Phosphorus Sulfur
com- com-
pounds pounds
H,POr S=

                       CHAPTER 17 / MANURE DECOMPOSITION IN SOIL / 251

 drained soil is aerobic, but the soil environment may become anaero-
 bic, particularly if conditions are favorable  for decomposition  and
 there is an excess of water.
     Under anaerobic conditions, which will occur in a very wet soil
 (as  in  over-irrigation),  denitrification  can  occur.  A  considerable
 amount of the nitrogen may  be  lost into the atmosphere, because
 1 unit of N can be lost for each 3.1 units of carbohydrate metabolized
 to CO,.
     Many anaerobic decomposition products, such as organic acids
 (acetic,  butyric, propionic, isobutyric) and  other  compounds, may
 be unfavorable for plant growth.  The iron may be reduced to a fer-
 rous  condition.  Foul-smelling compounds, such  as  indole, skatole,
 mercaptans, hydrogen sulfide,  and amines, are byproducts of protein
 decomposition (Table 17.8).
     Temperature is another important  factor. When the soil is cold,
 decomposition is slow.  Rothwell (1955) found that breakdown rates
 at 45°, 60°, 70°, and 80° F were about 30, 60, 70,  and 80% , respec-
 tively, of the rate at 95° F.
     The maximum amount of manure that the soil will accommo-
 date in decomposition has not been determined.  Indeed, if the land
 were  covered  with  several inches  of manure, considerable decompo-
 sition would occur in the manure pack where thermophilic bacteria
 may be  active in the decomposition. Temperature in manure  packs
 will reach  160° F even in winter.

     Application of animal waste  to the surface or incorporation in
the  soil is followed by  further decomposition.  Manure should  be
immediately  plowed under to minimize  N loss  (Table  17.9).  About
three-fourths or more of the  organic materials W7ill be decomposed
in the first year.  The mineralization of the animal waste will result
in nitrogen,  phosphorus, potassium, and micronutrients  becoming
available to plants, but there is no evidence that manure is superior
to inorganic fertilizer (Tables 17.10  and 17.11).  Further evidence is
needed to  evaluate any  contribution due to organic matter present

TABLE!  17.9.  Effect of plowing manure under at different times after appli-
            cation on crop yield.

                                            Relative  Value in
                                          Increasing Crop Yields
(15 experi-
Manure plowed under immediately 	 100
Manure plowed under 6 hours after spreading 79
Manure plowed under 24 hours after spreading 73
Manure plowed under 4 days after spreading . . 57
(1 experi-
Source:  Salter and Schollenberger (1939).


 TABLE 17.10.  Long-time yields with manure and fertilizer are comparable.
England ....

. Wheat

> 75
> 50

(bill a)


(bul a)
12 6
32 0

 or microorganisms  carried and growth-promoting or growth-inhibit-
 ing effects possible. Embleton and Jones  (1956) showed  that yield
 of oranges was the same when 2 pounds of nitrogen  were applied
 per tree as manure or  as commercial fertilizer  annually  when  the
 soil was  tilled, but was lower with manure  applications  when  the
 soil was not tilled. Manure was also an efficient source of phosphate
 and potash for the trees.
     Excessive mineralization of animal waste in the soil may result
 in leaching of nitrate into the groundwater and  runoff with  N and
 P. Huge quantities of animal waste applied to the  land may result
 in accumulation of some organic and inorganic constituents in con-
 centrations that may become toxic to plants, particularly under an-
 aerobic decomposition conditions  (Megie et al., 1967).  For example,
 corn seeds planted into manure will not germinate (Figure 17.2).
     Gaseous products, such as CO2, NH3, NOo, NoO, and N2, become
 a part of the soil  air and  may  return to the  atmosphere.  Small
 amounts  of organic acids  and other  odor-forming compounds  are
 gaseous.  When  released in  the  soil,  some (e.g., NH3, NOo, H2S)
 may be sorbed, and others (e.g., organic acids) may be metabolized,
 lowering the volatilization.
     The  readily decomposable organic constituents will be rapidly
 utilized by microorganisms.  Of the materials remaining, humuslike
 substances become  a part of the  humus complex of the soil .(Dorr,
 1965).  The  ultimate  accumulation  of organic constituents  in  the
 soil, however, will  be only a  small fraction of the total  organic
 material applied to  the land.  Salter and Schollenberger (1939) indi-
 cated  that  the beneficial physical  effects on  the soil  of  adding
 manure are probably overestimated.
     There is a considerable backlog of information on the applica-

 TABLE  17.11.   Corn  yield  of three varieties  with  manure  and fertilizer on
              sandy clay  loam.

         (N 4- P.O5 + K;O)            AES704   C103XB14  WF9XB14
(Ib/a)                            (bu/a)
                                   159        143
                                   168        144
                                   154        144
(1)  1240 -f 730 -f 1030  ..............   147
(2)  1270 -f 750 + 600 -f 66 T manure . .   143
(3)  800 + 515 -f 600 + 66 T manure . .   141
Source:  Unpublished data, D. G. Woolley and L. R. Frederick (1960).
Note: Nos. 1 and 3 have comparable amounts of N, P, K.

                       CHAPTER 17 / MANURE DECOMPOSITION IN SOIL / 253
        FRESH   MANURE
                                           FIG. 17.2.  The  influence
                                           of manure on the germina-
                                           tion of corn after 5-day in-
                                           cubation  at  25° C.  The
                                           control was planted in soil
                                           and the other seeds were
                                           planted in fresh  manure.
tion of animal waste in small to moderate amounts to the land  and
its effect on crops and on the physical, chemical, and biological prop-
erties of the soil (Hastings, 1938; Salter and Schollenberger,  1939).
But there are many unanswered questions in regard to the  applica-
tion of large amounts of animal waste to the land, such as the effect
on crop growth and on the pollution of surface and  groundwaters.

Alexander, Martin. 1961. Soil microbiology.  New York: John Wiley.
Atkinson, H. J.,  Giles, G. R., and  Desjardins, J. G.  1954.  Trace
     element content of farmyard manure. Can. J. Agr. Sci. 34:76-80.


 Burnett, W. E., and Dondero, N.  C.  1969.  The microbiology and
      chemistry  of poultry waste decomposition and associated  odor
      generation. (Mimeo.) Presented Cornell Animal Waste Man-
      agement Conf., Syracuse, N.Y., 13-15 Jan. 1969.
 Burroughs, A. L.  1967. Viral  respiratory infection in commercial
      feedlot cattle.  Am. J. Vet. Res. 28:365-71.
 Commoner, Barry.  1968. Threats  to the integrity of the nitrogen
      cycle:   nitrogen  compounds in  soil,  water,  atmosphere,  and
      precipitation.  (Mimeo.)  Presented Ann. Meeting  Am.  Assoc.
      Advan. Sci., Dallas, Tex., 26 Dec.  1968.
 Crampton,  E. W., and Harris, L. E.  1969.  Applied animal nutrition.
      2nd ed. San Francisco: W. H. Freeman.
 Dale, A. C., and Day, D. L.  1967.  Some aerobic decomposition prop-
      erties  of dairy cattle  manure.  Trans.  Am. Soc. Agr.  Engrs.
 Dorr, R. 1965.  The  characterization of  the organic substance  in
      manures. II.  Groups of organic  manures and residues, and a
      proposed scheme of a simple  analysis based on  the oxidizable
      carbon.  Landwirtsch.  Forsch.  18:238-46.   (#10096,  Biol.
     Abstr., 1968)
 Embleton, T. W., and Jones, W. W. 1956.  Manure as  a source  of
     nitrogen. Calif. Agr. 10:14-15.
 Geldreich, E. E., Bordner, R. H., Huff, C. B., Clark, H. F., and Kabler,
     P. W.  1962.  Type distribution of coliform bacteria in the feces
     of warm-blooded animals.  /. Water  Pollution Control Federation
 Gilbertson, C. B., McCalla, T. M., Ellis,  J. R., Cross, O. E., and Woods,
     W. R.   1970.  Beef  feedlot wastes:  characteristics of runoff,
     solid wastes and nitrate movement on dirt feedlots as affected
     by animal density and feedlot  slope. /. Water Pollution Control
     Federation.  (Submitted.)
 Hastings, Stephen  H.   1938.  Influence of farm manure  on yields
     and sucrose of sugar beets. USDA Tech. Bull. 614.
 Hemingway, R. G.  1961.  The mineral  composition of  farmyard
     manure.  Empire J. Exptl. Agr.  29:14—18.
 Jansson,  S. L.  1960a.  On the properties of organic manures.  I.
     Actual  humus  properties.  Uppsala Lantbrukhogsholans Ann.
 	.  1960b. On the properties of organic manures.  III.  Potential
     humus   properties.    Uppsala   Lantbrukhogsholans   Ann.
 Jenkinson, D. S. 1965.  Studies on  the  decomposition of plant ma-
     terial in  soil.  I.  Losses of carbon from 14C-labelled  ryegrass
     incubated with soil in the  field.  /. Soil Sci. 16:104-15.
 	.  1966.  Studies on the decomposition of plant material in soil.
     II. Partial sterilization of soil and the soil biomass.  /. SoiZ Sex.
 King, N. B.   1957.  The  survival of  Brucella abortus in manure.  /.
     Am. Vet. Med. Assoc. 131:349-52.
 Klein, D. A., and Casida, L. E., Jr.  1967.  E. coli die-out from normal
     soil as related to nutrient availability and the indigenous  micro-
    flora. Can. }. Microbiol. 13:1461-70.
Lipper, R. I. 1969. Design for feedlot waste management—history
     and characteristics.  Presented  at seminar, Design  for Feedlot
    Waste Management, Topeka, Kans., 23 Jan. 1969.
Loehr, Raymond C.  1968. Pollution implications of animal wastes—

                       CHAPTER 17 / MANURE DECOMPOSITION IN SOIL / 255

     a fonvard-oriented review.  U.S. Dept. Interior, Fed. Water Pol-
     lution Control Admin., Robert S. Kerr Water Res.  Center, Ada,
McCalla, T.  M.,  and Haskins, F. A.  1964. Phytotoxic substances
     from  soil  microorganisms  and crop residues.  Bacterial. Rev.
McCalla, T. M., and Viets, F. G., Jr. 1970.  Chemical and microbial
     studies of ivastes from beef cattle feedlots.  Nebr. Exp. Sta. Publ.
     (In press.)
McCalla, T.  M., Ellis, J.  R., Gilbertson, C. B., and Woods,  W. R.
     1969a.  Chemical  studies of runoff from  rain and snowmelt
     from beef cattle feedlots.  Agron.  Abstr.,  pp. 84-85.
McCalla, T. M., Ellis, J. R., and Woods, W. R.  1969b.  Changes in
     the  chemical and  biological  properties of beef cattle  manure
     during decomposition. Bacterial. Proc., pp. 4—5.
McCoy, Elizabeth.   1967.   Lagooning of  liquid manure (bovine):
     bacteriological aspects. Trans. Am. Soc. Agr. Engrs. 10:784-85.
	.  1969. Removal of pollution bacteria from animal waste by
     soil  percolation.  (Mimeo.) Paper  £69-430, presented  at the
     Ann.  Meeting of  the  Am.  Soc. Agr.  Engrs., Lafayette, Ind.,
     22-25 June 1969.
Megie, Christian A., Pearson, R.  W.,  and Hiltbold,  A. E.  1967.
     Toxicity of decomposing crop residues to cotton germination
     and seedling growth.  Agron. J. 59:197-99.
Morrison, S.  M., Grant, D. W., Nevins, M. P., and Elmund, K. 1969.
     Role  of excreted  antibiotic  in modifying  decomposition of
     feedlot waste.  (Mimeo.) Paper presented at the Cornell Animal
     Waste Management Conf.,  Syracuse, N.Y.,  13-15 Jan.  1969.
Perkins,  H. F., Parker, M.  B., and Walker, M. L.  1964.  Chicken
     manure—its production,  composition,  and use as  a fertilizer.
     Ga. Agr. Exp. Sta. Bull. N.S. 123.
Polheim, P.  1965.  Characterization of  the organic matter in  ma-
     nures.  I.  Classification  of organic manures on  the  basis of
     solubility of organic substances and  of nitrogen in  organic bond.
     Landwirtsch. Forsch.  18:228-37. (#10099, Biol. Abstr.; 1968)
Rothwell, D. F.  1955.  The influence  of temperature  and  nitrogen
     on the decomposition of plant materials mixed with soil. Ph.D.
     thesis.  Purdue Univ.  Library.
Salter, Robert M., and  Schollenberger,  C.  J.  1939.  Farm  manure.
     Ohio Agr. Exp. Sta. Bull. 605.
Stewart, B. A.,  Viets, F. G., Jr., Hutchinson, G. L., Kemper, W. D.,
     Clark, F. E., Fairbourn, M. L., and Strauch, F.  1967.  Distribu-
     tion of  nitrates and other water  pollutants under fields  and
     corrals in  the  middle South Platte  valley  of Colorado.  USDA,
     ARS 41-134.
Taiganides, E. P., and Hazen, T. E. 1966.  Properties of farm  ani-
     mal excreta.  Trans.  Am. Soc. Agr. Engrs.  9:374—76.
Wadleigh,  Cecil H.  1968.  Wastes in relation to agriculture  and
     forestry. USDA Misc. Publ. 1065.
Waksman,  Selman A.   1938.  Hinnus.  2nd  ed.  Baltimore:  Williams
     and Wilkins.
Witzel, S. A., McCoy, E., Polkowski, L. B., Attoe, p. J.,  and  Nichols,
     M. S.  1966.  Physical, chemical and bacteriological properties
     of farm wastes (bovine animals).  In Proc.  Symp. Management
     of Farm Animal Wastes, pp. 10-14.


          HEN animal manure is mixed with water, the biochemical
 reactions are both rapid  and predictable. The keys to the chemical
 transformations in aqueous suspension of manures lie in the chemi-
 cal composition of the manures, the microbes present, the  environ-
 mental conditions, and the time of exposure. It is important to under-
 stand each of these major variables and their interactions.  Too cften
 engineers and scientists examine only a portion of the problem and
 fail to recognize the fundamental concepts that underlie all manure
 transformations in water.

     The chemical characteristics of manure are primarily dependent
upon the chemical characteristics of  the feed processed through the
animals. Only a  small fraction of the feed eaten by any animal is
processed into animal tissue.  The feed is transformd internally into
materials which can be either absorbed or passed through the animal.
Waste products of metabolism are largely collected in the urine and
are passed out of  the animal along with the solid manure.
     Not very much research  has  been carried out correlating the
chemical characteristics of feed and  the chemical characteristics of
manure, yet a major part of the problem in evaluating the  chemical
characteristics of  manure  lies in knowledge of the  chemical varia-
tions in feeds. With hogs,  approximately 30% of the feed consumed
is  converted to animal tissue and  70% is excreted in  the form of
urine and manure (Irgens  and Day, 1965).  With cattle, the conver-
sion rate is lower, approximately 10%.  With a feed  consumption of
5.0 Ib/day for a 100-lb hog, the manure should contain 3.5 pounds of
the feed along with the excess water, which will be approximately 1
gallon. This would indicate that hog  wastes should contain 300,000
     Ross E. McKiNNEY is Professor, Department of Civil Engineering,
     University of Kansas.



mg/1 total solids. Generally, additional wasted water results in lower
values of solids  than indicated.  It is important that complete mate-
rial balances be made to determine the fate of all materials fed to the
animal.  It may well be that attention to the complete animal system
could result  in more efficient feeds and lower manure productions.
     Interest in  animal manure  pollution problems has stimulated
interest in chemical analysis of manures with respect to water pollu-
tion parameters. A recent study on hog  manure  (Schmid and  Lip-
per, 1969) indicated that the volume of urine and  manure should be
approximately 0.9  gal/day/100 Ib live weight for hogs in  confine-
ment fed a sorghum-grain-soybean meal ration. The COD of the ma-
nure was 0.52 lb/day/100 Ib live weight, while the BODU was  0.20
lb/day/100  Ib live  weight.  A review of published data (McKinney
and Bella, 1968) indicated an ultimate BOD of 0.50 lb/day/100 Ib
live weight.  There  is  no doubt that variation in feed composition is
a significant factor in  the variation in manure characteristics.
     Beef cattle manure collected at  the University  of Wisconsin
(Witzel et al., 1966) yielded 1.0 Ib BODr/day/1,000 Ib live weight and
3.3 Ib COD/day/1,000 Ib live weight.  Since cattle tend to be fed more
varied rations than  hogs, the chemical characteristics of the manure
will also be more varied.

     Manure contains  a  tremendous population of microorganisms.
Unfortunately there have been few studies to delineate the  various
types of microorganisms  in manure. Beef cattle being  rumen orga-
nisms have a more complex microbial flora than hogs.  One of  the few
studies (McCoy, 1967) on beef cattle manure indicated a wide variety
of bacteria related to the feed consumed by the cattle.  As expected
there were proteolytic microorganisms, lactic acid producers,- as well
as cellulose and pectin fermenters. The presence of methane-produc-
ing bacteria has been well established in cattle manure.  Special mi-
crobial techniques are required to enumerate the rumen bacteria due
to the anaerobic environment in the rumen and  the varied metabolic
characteristics  of these microorganisms.
     It suffices to say that the microbial population in animal manure
is more than adequate to bring  about  the chemical transformations
which will occur when the manure is mixed with water.

     Anaerobic Environment

     When manure is mixed with  water, microbial activity is  very
rapid.  The oxygen is removed so quickly that it has no significant
effect on the anaerobic bacteria which were growing in  the manure
prior to its discharge from the animal.  The complex organics are hy-
drolyzed  further to yield organic acids from proteins and cellulose.


 The ammonia released from the protein helps prevent the pH from
 dropping too rapidly.  But  as  the cellulose is decomposed, the pH
 can drop sharply, causing the environment to retard further bacterial
 action.  As the pH drops below 5.5, microbial activity slows and only
 the acid-forming bacteria continue metabolism. Eventually the acid
 buildup will cause  all  microbial activity to cease.  The acidified ma-
 nure will remain stable until the acids are either removed or neutral-
 ized.  If the  acidified  manure is agitated, numerous odorous com-
 pounds will be discharged from the liquid phase.
      If the pH of the manure does not drop below 6.0, the methane
 bacteria  will  metabolize the  organic acids, producing a satisfactory
 environment  for further metabolism.  The ammonia released from
 protein metabolism reacts with carbon  dioxide to form ammonium
 bicarbonate, which acts as a buffer to held the pH at a favorable level.
 Under these conditions the acid-forming bacteria continue to metabo-
 lize the complex organics, forming organic acids which are immedi-
 ately neutralized by the ammonium  bicarbonates.  The neutralized
 acids are metabolized by the methane bacteria to reform  the ammo-
 nium bicarbonate buffer. The nonbiodegradable organics remain un-
     The extent of degradation  of organic matter by the acid-produc-
 ing bacteria and the methane bacteria depends largely upon the time
 of contact and the extent of mixing.  These two environmental fac-
 tors are  very important in  determining the  extent  of metabolism.
 There is  no simple  formula  for determining the time of contact for
 metabolism to be carried to  completion. It suffices to say that the
 more microbes present, the  higher the temperature, and the  better
 the mixing, the shorter will be the time for metabolism.  There are
 basic limits to this concept. The amount of organic food present will
 limit the maximum microbial population which can  be maintained.
 Temperatures above 37° C will become toxic  to the mesophilic bac-
 teria and will require the development of thermophilic bacteria ,if me-
 tabolism  is to continue. Generally it is not  possible to  obtain too
 much mixing  but it  should be recognized that  little additional benefit
 can be derived from mixing  above the optimum level.
     One of the most critical environmental factors affecting anaero-
 bic metabolism has  been shown to be toxicity related to soluble ca-
 tions (McCarty and McKinney, 1961).  It was  demonstrated that am-
 monium ions were toxic to the methane bacteria in anaerobic  diges-
 tion systems.  Since animal urine contains considerable amounts  of
 urea which is readily hydrolyzed to form ammonium  ions, ammon-
 ium ion toxicity should be an important factor in the complete anaero-
 bic metabolism of concentrated manures.  This fact has been con-
 firmed recently (Schmid and  Lipper,  1969) in studies on controlled
 anaerobic digestion of hog manure.
     It. should be recognized  that the acid-forming bacteria merely
hydrolyze the  biodegradable  components of  i^e  organic  manures.
There is no change in the COD or the BOD of  the wastes.  If the ma-
nure contains  large  quantities of inert, nonbiodegradable materials,
there will be little apparent change. BOD and COD reductions  occur
when the  methane bacteria convert the soluble organics to methane,


 an insoluble gas that is discharged to the atmosphere above the liq-
 uid. If the methane were not lost from the liquid phase, there would
 be no decrease in COD  or BOD of the system, only transformation of
 the biodegradable organics from one form to another form.
     If the water diluting the manure contains nitrates or sulfates, the
 bacteria will reduce the nitrates to nitrogen gas while oxidizing the
 organic matter or will  reduce  the sulfates to sulfides.  Nitrogen gas
 dees not create a BOD or COD and would  result in stabilization of the
 organic matter. On the other hand, the hydrogen sulfide would exert
 an oxygen demand unless it was lost to the atmosphere.  It should be
 recognized from an energy standpoint  that microbes will reduce ni-
 trates  completely before reducing sulfates (McKinney and  Conway,
 1957).  Both nitrates and sulfates will be reduced before methane will
 be formed.   This relationship  is very important in understanding
 anaerobic transformations.
     Aerobic Environments

     In an aerobic environment free dissolved oxygen is present for
the microbial reactions. Initially the organic matter is oxidized to car-
bon dioxide, water, and ammonia.  The oxidation reaction results in
energy transfer from the manure to the microbial cells. The microbes
use this energy to synthesize new microbial protoplasm.  Aerobic me-
tabolism results in approximately one-third of the organic matter me-
tabolized being  oxidized and two-thirds of the organic matter being
converted to cellular protoplasm.
     With true aerobic conditions the bacteria growth stimulates the
growth of protozoa.  Like the bacteria, the protozoa oxidize a definite
amount of organic matter while converting a portion of the organic
matter to new protozoan cells.  The protozoa use bacteria as  their
source of food, thereby reducing the total amount  of bacteria in the
     As long as  dissolved oxygen remains in  the liquid  the bacteria
will metabolize  all of the biodegradable organics contained in the
manure. The metabolism of the protein components produces am-
monium bicarbonate which holds the pH at  the proper level, between
6.5 and 8.5, for good bacterial growth. If sufficient time is allowed,
nitrifying bacteria will grow and oxidize the  ammonium ions to ni-
trites and then to nitrates. Since this reaction results in the conver-
sion of a base to an acid, there will be a definite pH drop.  The degree
of pH drop depends primarily upon the amount of buffer present that
is not  related  to ammonium ions.  If oxygen should become limiting
after nitrification has  occurred, denitrification will result.  The bac-
teria metabolizing the organic  manure  will use nitrates  almost as
readily as oxygen and will reduce the nitrogen gas. Denitrification  is
as odorless as aerobic metabolism since metabolism is complete.
     One of the major problems  in aerobic metabolism of animal ma-
nure is supplying sufficient oxygen  to maintain  the aqueous  system
aerobic.  This can be done only with dilute aqueous suspensions of
manure.  Failure to maintain aqueous manure systems aerobic has


 caused numerous problems in trying to arrive at satisfactory aerobic
 treatment systems.  Dilution can  be carried out by the addition  of
 fresh water or by the use of treated effluent.
      The  aqueous  suspension of manure after aerobic metabolism
 will contain the bacteria produced from the metabolic reactions  as
 well as the inorganic salts in the  urine and in the manure and the
 nonbiodegradable organics, both suspended and dissolved.  The bac-
 teria will undergo endogenous metabolism with  time until only  an
 inert residual  of dead cells will remain.  The inert residual of dead
 cellular solids will contain  about one-fifth of the volatile solids in the
 original microbial mass produced and all of the in organic solids in the
 original microbial mass.  Thus it is that neither aerobic nor anaerobic
 metabolism will result in complete  degradation of manure in aqueous
 systems.  Yet, microbial transformations can convert unstable manure
 which  is  difficult to handle into  a fluid material which is easy  to
 handle and contains all of the elements in the original manure. This
 permits easier spreading of treated manures onto fields without the
 creation of obnoxious odors and nuisance conditions.

     There is no doubt that mixing animal manures with water will
 result in serious environmental pollution problems as a result of un-
 controlled microbial reactions which will be predominantly anaerobic.
 For this reason it is necessary to develop systems  which employ con-
 trolled microbial reactions in order to transform  the manure into a
 form where it can be returned into the environment without creating
 a serious pollution problem. A number of aqueous treatment systems
 have been developed and  studied to date.
     Oxidation  Ponds

     The simplest form of liquid treatment for animal manures is the
oxidation pond. In all respects, oxidation ponds have not proved suc-
cessful for animal wastes. Oxidation ponds have tended to produce
obnoxious  odors and poor quality  effluents  (Clark,  1965; Hart and
Turner,  1965).
     Fundamentally,  there is no reason why oxidation  ponds could
not be used to treat animal manures satisfactorily.  The  problem lies
in use of inadequate dilution of the concentrated manures. Generally,
animal manure lagoons are designed to operate on  minimum water
volumes in order to eliminate any effluent.  The net result is that con-
centrated manure is  discharged  at a  single point in the oxidation
pond.  The heavy manure solids tend to settle out around the inlet,
creating an anaerobic environment at a single point. The building
up of solids results in an acid environment due to anaerobic metabo-
lism and failure to distribute the acids into the liquid so that methane
fermentation can occur. Anaerobic  metabolism also results in  carbon
dioxide formation. The carbon dioxide  is released as a  gas as acids


 build up and depress the pH. The carbon dioxide gas causes solids to
 rise to  the surface and  permits  odorous  compounds to be released
 into the environment.
     The important concept to recognize in the use of oxidation ponds
 is dispersion of the organic  matter throughout the pond liquid.  It
 must be recognized that simple oxidation ponds cannot be  used  for
 animal manures. There  is no way for the manure to be dispersed in
 a simple pond system.  It is possible to use a large pump to dilute  the
 manure prior to its addition to the oxidation pond.  The treated efflu-
 ent could be used to flush the manure from the animal house.  Un-
 fortunately, the quantity of liquid which  would have to be recycled
 is quite large, around  200 gallons per day per hog or 1,000 gallons
 per  day per steer.
     It is essential that there is adequate volume in the oxidation pond
 for  good  metabolism.   A 4-foot deep  oxidation pond can handle
 around 40 Ib BOD-/acre/day. This would require 1  acre of oxidation
 pond for 200 hogs or 40 steers. For either confined hog growing or
 cattle feedlots, the use of oxidation ponds requires far too much land
 area. A 50-sow farrowing house would require 0.3 acres of oxidation
 ponds to treat the wastes, provided there was good mixing. Larger
 installations  create even greater problems due to mixing.
     Aerated Lagoons

     Mechanical aerators have been added to oxidation ponds in an
effort to produce better mixing and to add additional oxygen.  In an
effort to reduce power costs to a minimum, the mechanical aerators
are generally undersized for the pond volume.  Best results are ob-
tained when mixing relationships are balanced  against oxygen trans-
fer characteristics.  Some  research (McKinney and Benjes,  1965)
indicated  that a mechanical  surface aerator was capable of trans-
ferring 1.5  pounds of oxygen per HP-hr with  a residual DO  of 1.0
mg/1 and that 14 HP was required to reproduce  good mixing in 1,000
cu ft of pond volume.  This meant that a 5-HP  surface aerator could
transfer 7.5 pounds of oxygen per hour and  mix 20,000  cu ft of
wastes. The 5-HP surface aerator  could treat  the wastes from 225
hogs or 45 head of cattle. The very long detention time, over 50 days.
would result in a high degree of stabilization of the BOD but would
still produce a large mass of solids  for disposal.
     Oxidation Ditch

     One of the most effective forms of the  aerated lagoon concept
has been the use of the oxidation ditch under a slotted floor. Mechan-
ical  rotor aerators circulate the wastes under the  slotted floor and
aerate the mixture. The aerobic environment results in stabilization
of the manure in an odor-free  environment.  Although the concept of
the oxidation ditch was originally tried in Europe, the first successful
field scale unit for treating animal wastes in the United  States was


 put into operation at the Paul Smart hog farm near Lawrence, Kansas,
 in January 1966. The results obtained from the study of several units
 (McKinney and Bella, 1968)  indicated that mechanical rotor aerators
 were capable of treating the  wastes from up to 275 hogs with  a 5-HP
 unit.  This would mean that the same unit could treat  the  wastes
 from  55 head of cattle.
     Foaming has been a serious problem in starting oxidation  ditches
 as well  as in improperly  loaded  units. A number of investigators
 (Sheltinga, 1966; Moore et al., 1969) have reported foaming problems.
 Aside from start-up, foaming occurs only when the unit is  overloaded.
 Maintenance of proper aerobic conditions with adequate mixing elim-
 inates foaming due to the manure.  Excessive use of  detergents or
 disinfectants could result in foaming but this would occur only under
 abnormal conditions.  Foaming should never be a problem in  a well-
 operated treatment unit.
     The oxidation ditch system is capable of metabolizing  all of the
 biodegradable components of the manure but normally will contain a
 large quantity of living microbial cells which would exert  a hi eh oxy-
 gen demand, around  1,000 mg/'BOD-.  These microbial cells  can be
 further  treated by oxidation ponds  or mixing  with soil. Further
 aeration alone  will also result in their stabilization.
     Anaerobic  Lagoons

     Recently it has been shown that properly designed anaerobic la-
goons can be used for pretreatment  of manure.  Field  units for hogs
(Curtis,  1966; Willrich, 1966) indicated  100 cu ft  of anaerobic la-
goon per 100-lb hog. By and large the anaerobic lagoons were merely
large sludge-holding ponds.  Periodically solids  were removed  and
placed on the land. An anaerobic lagoon for dairy cattle (Lrehr  and
Ruf, 1968)  operated  at 9 Ib BOD,-/day/l,000 cu  ft. Cattle-waste
anaerobic lagoons have been operated at higher organic loadings than
hog-waste lagoons  as  a result of the  nature of the wastes.   Cattle
wastes generally have a higher population of methane  bacteria  and a
better buffering capacity. This generally permits cattle-waste lagoons
to operate more efficiently.  The  problem with efficient operation of
anaerobic lagoons lies in adequate  mixing.  Gas production by  the
methane bacteria alone will not produce the desired degree of mixing
as heavy organic loads and untreated solids will accumulate in  the
anaerobic lagoon.  It should be recognized that anaerobic treatment
of cattle manure will result in only 20% total solids reduction.  While
the lagoon system will remove 80 to 85% of the solids, the accumu-
lated solids must be removed eventually.
     Concentration  of solids to  10%  would require  approximately
one-half cu  ft per day per  head of  cattle for sludge  storage  alone.
It is important  that sludge storage be  provided  for  a period of  6
months to a year to reduce the time  intervals for sludge removal.  In
effect,  sludge storage  capacity  should  equal  the active  anaerobic
lagoon capacity for beef cattle.



     With regard to aqueous treatment systems for animal manure, it
is apparent that aqueous treatment systems are  not desirable for
animal wastes except in special situations.  The concentrated animal
wastes are not normally mixed with water and can be handled best
as solid wastes.  This is especially true of cattle manure.
     The advent of confined  animal growing has posed some changes
in the philosophy of handling manure as a solid waste.  Chicken
houses have been designed  to collect manure as a solid on moving
belts and  to transport it from the source to a  point of concentration.
On the other hand, confined hog houses have  too much fluid manure
for handling as solids.  The oxidation ditch has proved a satisfactory
system for both collection and treatment of  hog manure; it  is de-
signed to  replace conventional collection  and disposal methods.  It
should be recognized that the treated hog manure  must be returned
to the soil the same as untreated manure. The soil is the ultimate
acceptor of all animal wastes. It is vital that this concept be acknowl-
edged and accepted as one of the basic factors in manure disposal.
     Studies are currently underway to demonstrate the use of the
oxidation  ditch for  handling cattle manure from animals grown in
confinement like hogs.  There is no reason why it should not  work.
     Regardless  of the  treatment system used, the biological  treat-
ment will  reduce only a small  fraction of the total solids of the ma-
nure. The residual solids and the soluble salts pose a major disposal
problem that must be considered as part of the total manure disposal
problem.  Fortunately,  biological  treatment of  the manure destroys
the obnoxious qualities and  results in a material which can be han-
dled relatively easily without the creation of sanitary problems.

Clark, C. E. 1965.  Hog waste disposal by lagooning.  /. Sanit. Eng.
     Div. Am. Soc. Civil Engrs. 91 (SAG):  27-41
Curtis, D. R. 1966.  Design criteria for anaerobic lagoons for swine
     manure disposal.  In Management of farm animal ivastes. Am.
     Soc. Agr. Engrs. Publ.  SP-0366, pp. 75-80.
Hart  S. A., and Turner, M. E.  1965.  Lagoons for livestock manure.
     J. Water  Pollution Control Federation 37:1578-96.
Irgens, R. L., and Day, D. L.  1965.  Laboratory studies of  aerobic
     stabilization of swine  wastes. Farm structures eng.  rept. Univ.
     of 111. Agr. Exp. Sta.
Loehr, R. C., and Ruf, J. A. 1968.  Anaerobic lagoon treatment of
     milking-parlor  wastes. /.  Water Pollution Control  Federation
McCarty, P. L.,  and McKinney, R. E.  1961.  Salt toxicity in anaerobic
     digestion.  /. Water Pollution Control Federation 33:399-415.
McCoy,  E.  1967. Lagooning of liquid  manure (bovine):  bacterio-
     logical aspects.  Trans. Am. Soc. Agr. Engrs.  10:784-85.
iMcKinney, R. E., and Bella, R.  1968.  Water qualitij changes in con-


     fined hog ivaste treatment.  Kans. Water Resources  Res. Inst.
     Rept. Univ. of Kans.
McKinney, R. E.,  and Benjes, H. H., Jr.  1965.  Evaluation of two
     aerated  lagoons.  /.  Sanit. Eng.  Div.  Am.  Soc.  Civil Engrs.
     91 (SA6): 43-55.
McKinney, R. E., and Conway, R. A. 1957. Chemical oxygen in bio-
     logical waste  treatment. Sewage Ind. Wastes 29:1097-1106.
Moore, J. A., Larson, R. E., and Allied,  E. R.  1969. Study of the use
     of oxidation ditch to stabilize beef animal manure in cold  cli-
     mates.   In  Animal ivaste  management, pp.  172—77.  Cornell
Scheltinga, H. M.  1966.  Biological treatment of animal wastes.  In
     Management of farm animal wastes. Am. Soc. Agr. Engrs. Publ.
     SP-0366, pp. 140-43.
Schmid, L. A., and Lipper, R. I.  1969.  Swine waste characterization
     and  anaerobic digestion.   In  Animal ivaste  management,  pp.
     50-57.  Cornell Univ.
Willrich,  T.  L.  1966. Primary treatment of swine wastes  by  la-
     gooning.  In Management of farm  animal wastes.   Am. Soc.
     Agr. Engrs. Publ. SP-0366, pp. 70-74.
Witzel, S. A., McCoy, E., Polkowski, L.  B., Attoe, O. J.,  and Nichols,
     M. S.  1966.  Physical, chemical and bacteriological  properties
     of farm  wastes (bovine animals).  In Management  of  farm ani-
     mal  wastes. Am. Soc. Agr. Engrs.  Publ.  SP-0366, pp.  10-14.


     V*ENTURIES  prior to  the  era of bacteriology, man realized
that water was  somehow involved  in  the  transmission of disease.
     Before  consideration of  current problems of disease  transmis-
sion related to water, the historical implication of water and disease
will be briefly reviewed.  The early miasmatic  theory of  disease
taught that all disease was due to emanations from water, earth, and
influence of the stars, moon, winds, and seasons. More than 2,500
years ago during the pre-Christian era,  the role of water was further
described by Hippocrates, the "father of medicine,"  in his treatise,
"Airs, Waters and Places" (Chadwick and Mann,  1950). He related
causes of disease to different waters, the wind, and  to the slope  of
the land.  These  findings were further advanced during  the  early
Christian period and the Middle  Ages.  During this  time epidemics
of certain diseases such as typhoid and cholera were  associated with
floods  and  the  rise  and fall  in the  level of  groundwater.   The
theory of poisonous miasmata and  vapors  (arising  from decaying
filth) held  until  the end of the nineteenth century.  Some  early
observations were inadequate and unsound, but others represented
correct observations of fact.
     During the nineteenth century, researchers, including Henle,
Snow, Budd, and Pasteur, developed the germ  theory of disease.  In
1876 Robert Koch proved the germ theory by his classical work  on
anthrax. Historical  aspects of bacteriology are well  described in a
book by Bulloch (1938).  The golden era of bacteriology has existed
and developed for nearly 100 years. Major emphasis has been  on
the importance  of the  microbiologic  agent in  causation of  com-
municable diseases.
     Recognition that predisposing or contributing factors of disease
must be identified,  and  multiple  causes  of  disease ex;st,  has
broadened man's efforts  to consider the  total perspective of disease—
the interrelationship of  the agent, host, and environment complex.
     STANLEY L. DIESCH is Associate Professor, Department of Veterinary
     Microbiology and Public Health, University of Minnesota.




      Knowledge  of epidemiology  as  related to disease  transmission
 is essential.  Epidemiology is the study of disease  as related to  the
 host, agent,  and total environment—or the ecology of disease.
      Of the many infectious diseases affecting  animals, more than
 150 are classified as zoonoses,  or those infections or infectious dis-
 eases transmitted under natural  conditions between vertebrate ani-
 mals and  man  (WHO,  1967b).  Zoonoses associated with  food-
 producing animals are usually considered occupational.  An increas-
 ing number of zoonotic diseases  associated with recreational activi-
 ties are being reported.
     The infectious disease process  contains  six necessary  factors.
 These factors are considered as links in a chain  and all are essential
 in disease development.  The six  essential factors are  as follows:
 1.  Causative or etiological agent
        Infection represents entry and development or multiplication
    of an infectious agent in the body of man or animal.  The para-
    sitic agent usually lives at the expense or detriment of the host.
    Fortunately, many  organisms are not  pathogenic for man and
    animals. Certain organisms have  specificity and will infect only
    a selected species. For example, hog cholera virus will not infect
    man or other animals.
 2.  Reservoir of the infectious agent
        Reservoirs are man,  animals, plants,  soil,  or inanimate or-
    ganic  matter.  Here  an infectious  agent  lives  and multiplies.
    With  few exceptions, pathogens  are  not  capable of prolonged
    growth  or  multiplication  outside  the  living body.  Significance
    of the animal  reservoir depends upon man's  direct or indirect
    association.  Man has much greater direct exposure to domestic
    animals  than  wild  animals.   Animals  and man are potentially
    and indirectly associated with animal pathogens through waters.
    Man remains the most significant reservoir of  infection for  his
    species, and animals for their kind.
3.  Escape of organisms from the reservoir
        Escape and subsequent discharge  of the organism  into  the
    environment may occur through natural body openings  (respira-
    tory, intestinal, urinary), by way of open lesions, and by mechani-
    cal means (blood-sucking arthropods).  A variation exists in time
    duration of escape of pathogens. This is dependent on the course
    of disease in  the  host animal. In general  the duration  of com-
    municability of an infectious agent varies inversely with  the
    degree of communicability.
4.  Transmission of the infection  from the  reservoir to the new host
        Transmission occurs by direct and indirect methods.  Direct
    contact  occurs  when organisms  pass  immediately to  the new
    host by  physical contact.  Indirect contact  occurs when  there is
    a  transfer of infectious agents between  the reservoir  and  the
    new host without direct association.  These organisms must be
    capable  of  surviving  outside  the body  and  a  vehicle or  vector

                               CHAPTER 19 / DISEASE TRANSMISSION / 267

    must transfer the organisms.  The classification of indirect meth-
    ods of transmission are vectors (arthropods or other inverte-
    brates) and vehicles (all nonliving objects or substances that are
    contaminated and transfer the infectious organisms).  Vehicles
    include water, milk, other foods, air, and fomites.
5.  Entry of organisms into new hosts
        Before entry, the organisms must pass defensive  barriers of
    the host.  With exceptions, the mode of entry into man or animal
    corresponds with the mode of exit.
6.  Susceptible host
        Man and animals possess defense mechanisms or  resistance,
    which protect against invasion of the pathogenic microorganisms.
    Immunity implies  the development of absolute protection in a
    susceptible host  against  disease by artificial or natural means.
     Development of a disease in man or animal depends on comple-
tion of several concurrent events  and includes the strength of the
six essential links of the chain.
     The following important factors concern  the  susceptible  host.
Age usually increases resistance, for the longer man or animal  lives,
the greater is the opportunity for contact with  specific microorga-
nisms and for development of  immunity. Incidence of disease in a
community is significant, for greater occurrence of  disease increases
opportunities for exposure.  Opportunities for spread include biologi-
cal, sccial, and  physical factors.  Environmental  factors  such as
water  supply, sanitation, housing, and crowding are  involved.
     The occurrence of disease in  populations has been inadequately
reported.  Cases (with clinical  signs or symptoms)  may be reported
but many  infected carrier (subclinical or inapparent) animals may
exist in a population.   Because the carriers are  often not recognized,
they are more capable of transmitting  disease to populations.  The
case-carrier concept may be viewed as a floating iceberg, with a  small
fraction of the ice observed (cases) and  the remainder under  water
inapparent  (carriers).  This phenomenon varies with each disease
     If a disease outbreak is manifest in a population, the outbreak
exists  until there is a  death, disablement, recovery, and/or develop-
ment of resistance against the specific disease.  Infected animals
may shed  millions of organisms  into the environment,  and  these
organisms  may find a  susceptible host.  Errington (1963) has stated,
"Nature's way is any way that works."
     To prevent, control, and  eradicate animal disease, treatment
with antibiotics and  chemotherapeutics, and  prevention  by use of
vaccines and bacterins, have been developed.   Quarantine, test, and
slaughter programs of identified infected animals have been used.

     A few decades ago a predominantly rural America existed with
wide dissemination of livestock populations. The raising of livestock


  and rural living remain predominant in many areas of the world.
  Today, fewer farms have greater concentrations of livestock.  In the
  United States in 1937, 24.3% of 128,649,000 people lived on farms
  containing 94,694,000 animal units.  In 1967,  5.4%  of  198,608,000
  people lived on farms containing 120,439,000 animal units.  In three
  decades there has  been  a 21.3%  increase  in  animal units, with a
  65% decrease in human farm population and a 35%  increase in total
  population (USDA,  1968b).
      An increasing number  of  animals are raised  in confinement.
  In the United States on January  1, 1969, there were 23,040,000 cattle
  on feed in lots.  Of  these, 10,823,000 were found on 2,080 lots, each
  with 1,000 or more  head of cattle (USDA, 1969). It is not uncommon
  to find feedlots  of  10,000  cattle or broiler  farms of 100,000. This
  concentration can greatly enhance disease-prevention programs, but
  may by increased contact  cause greater problems in disease trans-
      The environment of the agricultural worker  allows greater ex-
  posure to infectious and parasitic  diseases than  is encountered in
  urban surroundings (WHO, 1962).
      As man migrated from the farms to the cities, controlled sewage
  disposal and chlorination  of water supplies have reduced  the in-
  cidence of illnesses such as typhoid fever, paratyphoid,  dysenteries,
  and  cholera.  Perhaps man, as a result of control of specific  water-
 borne diseases, has developed a placid attitude concerning  water-
 associated disease.
      Living in the city, man has increasingly been seeking his  out-
 door recreational activities  in  rural  areas.  Most  people  seeking
 outdoor recreation wish to be near water. Swimming will  be the most
 common form of outdoor recreation by the year 2000 (U.S.  Outdoor
 Recreation Res. Rev. Comm., 1962). Being  exposed to the environ-
 ment of domestic and wild  animals and surface waters will increase
 man's exposure to waterborne infections.
     Water is absolutely essential to maintain the bodies of both man
 and animal.  In  the United States much of man's water supply for
 household use is from deep  wells  or  chlorinated, treated supplies.
 Man  continues to be exposed to  surface water through occupational
 and  increasing  recreational  activities.  Confined animals receive
 much of their water from deep wells, but those  on ranges in pasture
 largely consume  water  from ponds,  streams, rivers,  and  lakes.
 Economics  of  agriculture demands  the fullest utilization of land.
 Often land adjacent to surface water can be used only for pasturing
 of livestock.  Millions of food-producing and  wild animals are found
 here.  If infected, pathogens escape into surface waters via respira-
 tory discharges, drainage of wounds, feces  or urine, or dead animals.
 Transmission of organisms from reservoir to water also occurs by soil
 runoff, flooding, wind, and  other ways. Due to the dilution, patho-
 gens discharged into water  may  be of  relatively low densities.  The
 general concept that  running water undergoes purification is counter-
 acted by the fact that infected animals may shed millions of patho-
gens for days, weeks, or months.

                               CHAPTER  19 /  DISEASE TRANSMISSION /  269


     In  consideration  of  agriculture's  role  in  maintaining  clean
 water, concern is for the  cause and effect or the effect and  cause
 relationship  of  waters  contaminated with  pathogenic  organisms.
 When total  ecology of disease is studied, complexity is greatly
 increased and by  definition decreased by the numerous interrelated
 factors involved.  To document water's  role  as a vehicle in disease
 transmission, information  gathered from a literature review will  be
 used.  Specific disease entities are grouped by classification based  on
 etiology of the causative organism.
     In view of the  scope  of  this subject it will be impossible  to
 discuss all diseases individually. No reference will be  made to pre-
 vention, control,  and treatment.  This  information  is available  in
 literature cited.   Specific  examples  in  each category are  briefly
 described,  with emphasis  on  resistance  and transmission of  the

     Bacterial Diseases

     Species  of  vegetative bacteria  vary greatly in  their ability to
survive away from the host.  Spore forms are very resistant to physi-
cal and chemical agents whose action can greatly affect  the growth
rate and death (Merchant and Packer, 1967).
     In 1854 water first  assumed an important role  in the transmis-
sion of disease when John  Snow  was able  to demonstrate the  rela-
tionship between human cholera and water from  the Broadstreet
pump  in London.  Since the development of the bacteriologic era,
numerous  documentations  of  wrater transmission of disease  have
been made.

     In the United States the major zoonotic disease is salmoiiellosis.
Approximately 20.000  human  cases  are reported each  year,  but
estimates are  that between 1 to 2 million cases occur (Steele, 1968).
The disease is widespread in food-producing animals, poultry,  and
other animals (Edwards and Galton,  1967).   These are the major
reservoirs for  man.  In acute cases in calves,  10,000,000 organisms
per gram of feces have been reported.1
     Salmonella survive in water and the environment for extended
periods of time (Kraus and Weber, 1958; Andre et al., 1967; Gibson,
1967).  The bacteria could  survive  several weeks  to  3  months in
drinking water and natural surface water (Kraus and Weber, 1958).
liibbs and Foltz in 1964 isolated Salmonella from  two calves, creek
water, and a human being. Schaal (1963) reported enzootic salmonel-

     1, K. L. Loken, 1967, personal communication.


 losis in cattle as a result of drinking contaminated brook water.  In
 May 19135 a serious epidemic of waterborne Salmonella typhimurium
 occurred, with  three human deaths (NCDC,  1965).  Of the human
 cases reported  each year, more  than half  are sporadic. The re-
 mainder are associated with epidemics that  can usually be  traced
 to contaminated foods  of animal origin or to  water (McCroan et al..
 1963; Steele, 1968).
      More than 1,300 serotypes of Salmonella have been  identified.
 These  bacteria are ubiquitous  and shed in  the feces of infected
 animals. Surface waters serve as  potential vehicles  for transmission
 of Salmonella to other animals or man.
      In  1966 a  large waterborne outbreak of  human cases occurred
 at Riverside,  California,  from  a  Snhnonenn-contarninated  water
 supply.  Although the source of contamination was  not identified, it
 was  speculated the water  may have been contaminated by seepage
 from distant cattle feedlots  (Decker and Steele, 1966).  Due  to the
 widespread occurrence of  reservoirs and environmental contamina-
 tion, salmonellosis continues to be a major disease entity.

     Leptospirosis,  caused by a spirochete,  has been classified as a
 waterborne zoonosis.  In the  United States and  many areas  of  the
 world, leptospirosis is found  in domestic animals and wildlife.  In
 domestic animals  the bacteria are found  primarily  in  cattle  and
 swine and may be  shed  in the urine for several  months.  Counts of
 100 million leptospires per ml of urine have been reported (Gillespie
 and Ryno, 1963).
     Leptospires may live  in water for several weeks  (Chang et  al..
 1948;  Gillespie and Ryno, 1963; Ryu and Liu. 1966).  However,  the
 changing environment may complicate survival (Diesch et al., 1969).
 Fresh water in all forms in nature is a major factor in the  circulation
 of leptospires in enzootic foci. The conventional idea that stagnant
 waters  and  slow-moving streams  are potentially infectious  is  not
 necessarily  valid.   The  infectiousness of rapid-flow  water in  the
 jungle and  increased infectiousness with flooding has been  shown
 (WHO, 1967a). Leptospires  have  been isolated from fast-moving
 streams (Gillespie and Ryno, 1963).
     Human outbreaks have  occurred when  people have come in
 contact with contaminated water through swimming or occupational
 exposure.  In the  United States  since  1941  approximately 1,000
 human cases have  been  reported.   Swimming has accounted  for 10
 outbreaks that involved  233 human cases.2  In  1964  Leptospira
 pomona was isolated from the  swimming site in a  creek where human
 cases occurred in  1959 and  1964 following swimming.  Cattle  and
 other animals frequented this  stream (Diesch and McCulloch,  1966).
 Sixty-one human cases occurred in  Washington following  swimming
in water contaminated by infected cattle (NCDC, 1965a).
     Between  1951  and 1960  the estimated  annual loss to the dairy
 and milk industry was more than $12 million per  year (USDA,  1965).
In 1969  the  Leptospirosis  Committee  of  the  United States Animal

     2.  W. F. McCulloch, 1969, personal communication.

                               CHAPTER 19 / DISEASE TRANSMISSION / 271

Health Association stated that leptospirosis is not amenable to eradi-
cation.  It is likely that water will continue to  serve as a vehicle of
transmission of leptospirosis to animals and man and remain one of
the major sporadic diseases associated with water transmission.

     In addition to being one of the oldest known diseases affecting
man and animals,  anthrax was the first zoonotic disease associated
with an etiologic agent.
     Anthrax  spores  are one of the most resistant of pathogenic
bacteria.  Spores stored in soil contained in a rubber-stoppered bottle
remained viable for 60 years (Wilson and Russell,  1964).  Field ob-
servations indicate  similar duration of viability in alkaline, undrained
soils in warm climates (Blood and Henderson, 1968).   There have
been instances of  animals becoming infected  on anthrax areas 25
years after the original cases of disease (Merchant and Packer, 1961).
     A major mode of dissemination of spores is by surface \vaters
flooding contaminated ground, causing transfer of spores to wide-
spread areas.  Many water courses in anthrax districts in the United
States are contaminated (Stein, 1942; Jones, 1963).
     Reported human cases of anthrax have declined steadily during
the past  50 years (Brackman,  1964).  Most  of the human cases
reported in the  United States in recent years  have been associated
with imported goat hair and coarse wool.  Estimates indicate  that
a decade ago the worldwide yearly  incidence was 20,000 to 100,000
cases (Classman, 1958).
     Animals are most commonly infected by ingestion  of contami-
nated food and water. Potential infection will  exist for many years,
especially in the contaminated anthrax  districts where surface water
plays a major role in transmission.

     Tularemia is a widespread, highly contagious disease  that has
been isolated from more than 100 kinds of wild and domestic animals
(Steele, 1968).  In U.S. agricultural animals, the disease is  reported
most commonly  in  sheep.  The bacteria do not  form spores. Re-
searchers reported water and mud  contamination and the occurrence
of tularemia in beaver and  muskrat as  widespread phenomena  in
northwestern United States. The tularemia organism has been found
in all streams  tested with any frequency in the Bitter Root Valley
(Hamilton area) of Montana (Parker et al., 1951).
     The organisms  are believed to be able to multiply in the mud,
leaf  mold,  and materials that make up  the beds  and shores of the
streams. The aerobic organism is recoverable from running waters
only, and never  found in still or stagnant streams. During  a  7-
year period in one stream, Francisella tularensis has been recovered
from approximately 30% of the specimens tested.3

     3.  Cora R. Owen, 1969, personal communication.


      Tularemia can be transmitted by many different routes (Shaugh-
 nessy,  1963).  There  is evidence that the bacteria will penetrate the
 intact skin (Quan et al., 1956).
      Four clinical and four probable human cases were  associated
 with contaminated water (Jellison et al., 1950). Two of the cases
 were associated with contaminated water supply (spring water); the
 bacteria  of tularemia were isolated  from water collected from the
 faucet.  In another report (Jellison  et al., 1942) contamination of
 four streams was found. One stream remained contaminated for 33
 days after any beavers were known to be present.  Since contamina-
 tion of water may persist for months and perhaps for years, drinking
 of water from streams in endemic regions should be  avoided. During
 a tularemia epidemic that occurred in Vermont. 47 human cases were
 linked to contact with muskrats; the tularemia organism was isolated
 from the mud and water of a trapping site (Young et al., 1969).  This
 was North America's largest outbreak of tularemia in man linked to
 aquatic mammals. Since the disease is established  in wildlife popu-
 lations it does not presently appear amenable to control.

     Brucellosis is a contagious  disease of cattle,  swine,  and goats
 and a major occupational disease of man.
     The bacteria are shed in the excretions and secretions, especially
 uterine, of infected animals.  In pastures and barnyards, brucellae
 have survived 65 to 182 days or more in dead fetuses and fetal mem-
 branes, and 2 months in manure (Bosworth, 1934).  In tap water the
 organism remains viable 10 to 120 days at 25° C and in bovine urine
 up to 4 days  (Van Der Hoeden,  1964).  Brucellae  survived in grass
 for 100 days in winter and 30 days in summer.  It survived freezing
 temperature over 824 days in cattle urine, lake water, tap water, raw
 milk, bovine feces, and soil (Ogarkov,  1962).  In the United  States,
 1975 is the target date  for the eradication of brucellosis. According
 to Harris  (1950),  water,  except when grossly  contaminated with
 brucella organisms, seems to be an unlikely  source of  human in-

     Erysipelas  is  of major  importance  and  widely  distributed.
causing swine erysipelas and affecting turkeys.  The bacteria occa-
sionally cause erysipeloid in man.
     The organism is resistant to drying and remains viable a month
or more in  the  dark and 10 to 12 days in sunlight (Morse, 1963).
It  exists in  soil as  a saprophyte and retains  virulence.  Persistence
in soil is  variable and determined by  temperature, pH, and other
factors. It is reported viable 4 to 5 days in drinking water and 12 to
14 days in  sewage  (Reed, 1965).  Soil, food,  and water are readily
contaminated by infected animals through large  numbers being dis-
charged in  the urine.  From soil experimentally inoculated, the or-

                               CHAPTER 19  / DISEASE TRANSMISSION / 273

ganisms were recovered to a maximum of  21  days.  Persistence was
longer during  winter  and spring (Rowsell, 1958).  Surface waters
may transmit the disease from one farm to another (Karlson, 1967a).
Since the bacteria can pass through  the  stomach without loss of
viability,  carrier animals  may continuously  contaminate  the soil
(Rowsell,  1958).

     Although in the United States bovine tuberculosis is no longer of
major importance,  the disease is still of major importance in  some
areas of the world.
     The  bacteria  are resistant  to  chemical and physical agents
(Middlebrook,  1965).  In some instances  virulent bovine tubercle
bacilli can survive  6 months exposure in soil, in soil-dung mixture,
and in dung (Maddock,  1933). It is reported (Christiansen, 1943;
Blood and Henderson, 1968) that stagnant drinking water may cause
infection  up to 18  days  after  being used by a tuberculous  animal.
Viable organisms were isolated from the  soil 6 or  8 weeks after feces
were dropped,  but  the duration varies widely—being longer in wet
weather.  According to Karlson (1967b), the bacillus is  transmitted
through feed and sometimes water.
     In the United States in recent years  only a rare human case was
caused by the bovine strain (Feldman, 1963).

     The disease is widespread and usually associated with the entry
of the bacteria into a wound. The organism, a sporeformer, is widely
distributed in nature and is abundant in animal  or  human feces,
especially of horses and other herbivorous animals (Sterne and Van
lieyningen, 1965;  Merchant  and Packer, 1967).  The  spore form
resists boiling for more than 1 hour.  The spores are capable  of per-
sisting in the soil for a number of years (Blood and Henderson, 1968).
With the rapid increase of the horse population in the  United States,
the subsequent contamination of the soil will likely increase.  Surface
water may play a  major role  in the  dissemination of  the  tetanus

     Colibacillosis has worldwide distribution and it is, under certain
conditions, associated with enteric infections in man and animals.
It is found universally in the intestinal tracts of man and animals.
The organism is usually destroyed at 60° C for 30 minutes.  Heat-
resistant strains may survive (Merchant and Packer,  1967) and in-
dividual cells survive freezing in ice for 6  months. The  organism is
transmitted by  water,  feces,  and  flies   contaminated  with  fecal
material.  Some strains are hazards to both  man  and animal and may


 cause illness of the newborn (Morgan, 1965).  The number of E. coli
 organisms found in water indicate the extent of fecal contamination.
 Attempts  to document association between  cases  in  agricultural
 animals and man appear to be inconclusive.
     Rickettsial  Diseases

     Agents  of  rickettsial diseases other than  Q fever depend on
 arthropod vectors  for  transmission  of  disease  and on  human  or
 animal hosts for their mechanisms (Fox, 1964).
     Q  FEVER

     Q  fever is found in man and animals on every continent of the
 world.  It  has a widespread  host range  (Babudieri,  1959).  In  the
 United  States it has agricultural significance in  sheep,  goats, and
 cattle.  The organism,  an intracellular parasite, has greater resist-
 ance to physical and chemical agents than other pathogenic  Rickett-
 sia  and has  more  resistance  than most  nonsporogenic bacteria.
 The agent  is viable in skim milk for 42 months and tap water for 36
 months (Ignatovich,  1959). Welsh et al. (1959) isolated the organism
 from standing water (surface pools) on  infected  sheep ranches in
 California  over a 6-week period during the lambing season, and from
 the  soil up to 148 days.  Stoenner (1964) reported that the role of
 microenvironments on  mobile fomites was significant in extending
 the  hazards of  the disease to diverse occupational groups not nor-
 mally considered at  risk.  He  estimated that in the United States at
 least 25%  of the dairy herds  and a higher percentage of sheep and
 goat herds are infected.  Q fever usually appears  as inapparent in-
 fection  in domestic livestock.
     The exact mode of transmission is unknown but dust-laden air,
 containing animal waste, and ticks are considered important.  One
 organism has been suggested  as an infectious dose for man  (Tigertt
 et al.,  1961).   The  role of   water  in  transmission  has not  been
     Viral Diseases

     There are  an estimated  500 known  animal  viruses  (Green,
1965). Counterparts of major groups of viruses known to infect man
are also found in  domestic  animals.  According to Abinanti (1964),
no thorough investigation has been made to determine which virus
may be present in milk and other animal by-products or under what
conditions they are destroyed.
     In an extensive review of enteroviruses of animals it was con-
cluded that the enterovirus problem of animals parallels that  re-
ported in  man and that  a multitude  of organisms may be isolated
from feces of  different animal species (Kalter, 1964). In recognition

                               CHAPTER 19 / DISEASE TRANSMISSION / 275

that human  health is closely related to health of animals, viruses
are considered  the least-explored  infectious  agents (Sinha  et  al.,
     In general, viruses do not survive for long periods of time out-
side the animal host (Gratzek,  1967).  Viruses possess about  the
same degree of resistance to heat, drying,  and chemical agents  as
many  of the vegetative forms of bacteria.  Most  are unaffected  by
concentrations of antibiotics that will destroy bacteria.
     According to Frier and  Riley (1965), some information on sur-
vival of viruses in water is available, but most of the data have been
obtained from distilled-water studies under controlled temperatures.
They stated that compared to bacterial and  protozoal agents, viruses
in natural waters,  except under unusual  circumstances,  are  pre-
sumed to survive only for a short period of time.  However,  Brown
and McLean (1967) stated that enteroviruses are more  resistant to
halogens than bacteria and unless residual free chlorine is sufficiently
high, water,  although free of viable bacteria,  may  contain active
    Joyce  and Weiser (1965) reported  that a  study  of  farm ponds
over a 6-month period revealed no enteroviruses or specific  bacterio-
phages.  They  experimentally inoculated pond water with entero-
viruses which survived for long  periods of time (present  up to  91
days) at simulated temperature  extremes  and over pH ranges  of
extremes found in  natural pond  waters.  The virus survived longer
in slightly  and heavily polluted waters  than in moderately polluted
waters. Chemicals  found in farm  ponds did not appear to  inhibit
viral survival.  Based on experimental findings, the conclusion was
that farm pond water poses a definite site for storage of enteroviruses.
    Less is known regarding the role  of water in  transmission  of
viruses than  bacteria.  Many of the viral diseases are transmitted  by
arthropods. Approximately 200 viruses  have been classified as arbo-
viruses (Merchant and Packer, 1967). The nonarthropod-borne viral
diseases are fewer in number but many are associated in the  United
States with animal industries.
     Many classes of viruses are excreted in the feces  of  animals.
Included are picornaviruses (enteroviruses), reoviruses (respiratory-
enteric viruses), herpes viruses, adenoviruses, and myxoviruses (Grat-
zek, 1967).
    According to Prier and Riley (1965), natural  water  is of minor
significance  when  compared with  other  factors  that  affect viral
transmission of disease between individuals and herds.
    Geldreich (1965), in describing the origin of microbiologic pol-
lution in streams, stated that water contaminated by fecal  pollution
may also contain viruses excreted by warm-blooded animals.
    Although man is primarily involved in viral hepatitis transmis-
sion, this agent, with high resistance  and capable of being trans-
mitted via surface waters, can serve as a study model.  Mosley (1963)
reports  a total of 31 human epidemics presumed to have been trans-
mitted by water. The hepatitis virus is not destroyed by  chlorination
or pasteurization  (Anderson et  al., 1962).  According to  Mosley
(1963) only the viral agent of infectious hepatitis has been clearly  as-


 sociated with  waterborne  transmission  in man's  drinking water.
 Water may also have a role in transmission of poliovirus, Coxsackie,
 ECHO, and adenovirus (Clarke and Chang,  1959;  Brown and Mc-
 Lean, 1967; Chang, 1968).
     The role of water in transmission of viral diseases has not been
 adequately documented or perhaps considered.  The following viral
 diseases of domestic animals are examples to show the variation in
 viral resistance.

     This virus causes an  acute  systemic  infection of fowls,  may
 infect man, and is highly resistant to detrimental factors of the en-
 vironment.  In chicken down and dust the virus remains  active for
 many weeks at ordinary temperatures (Bernkoph, 1964).

     Hog cholera  is an acute, highly contagious disease of swine.
 caused by a relatively stable virus.  One report stated that the virus
 at 37° C  survived for  7 but not  for 15  days (Bruner and Gillespie,
 1966).  Survival time may be longer and varies with environmental
 conditions.  Transmission is believed to be primarily by contact with
 infected swine, or indirectly by secretions and excretions.  The target
 date for eradication of hog cholera in the United States is 1975.

     Foot-and-mouth  disease is an extremely acute contagious dis-
ease of all cloven-footed animals that rarely infects man.  The virus
is resistant to external  influence including common disinfectants.
It may persist for more than 1 year in infected premises.  The virus is
rather susceptible to heat and  pH change and insensitive  to cold.
Many  methods  of transmission occur,  with the  common  method
believed to  be  ingestion of  contaminated  feedstuff  (Blood  and
Henderson, 1968).

     The role of water in the transmission of many agents of disease
is unknown. Of recent interest are three  disease  entities causing
similar chronic neurologic disorders:  scrapie in sheep, mink enceph-
alapathy,  and Kuru in man. The etiologic  agents are extremely re-
sistant and have  long incubation periods  (McDaniel,  1969).  The
agent of scrapie in sheep resists exposure to  75° C for 1 hour, is ether
resistant,  and brain tissue in 10 to 12% formalin is still viable after
4 to 28 months (Merchant and Packer,  1967).  The role of water
transmission is unknown.

                               CHAPTER 19 / DISEASE TRANSMISSION / 277

     Fungal Diseases


     The agents of systemic mycotic diseases of importance are acti-
nomycosis, nocardiosis, aspergillosis, phycomycosis, candidiasis, his-
toplasmosis, North American blastomycosis, coccidioidomycosis, cryp-
tococcosis, and sporotrichosis. With the exception of candidiasis the
others are found free living in nature and are not considered to be
zoonotic.  These diseases  are  known as occupational fungi.   The
fungi are cultured with  ease from soil containing chicken manure,
starling roost, and pigeon  feces (Harrell,  1964).  Animals and man
are susceptible to  these fungi found in the environment. The spores
are airborne-transmitted.  Infected animals are not considered res-
ervoirs  for the transmission of  disease to  man (Menges, 1963;
Maddy, 1967).

     Histoplasma infection  of  man and  animals is widespread in
midwestern United States. It is reported sporadic in animals (Blood
and Henderson, 1968).  Evidence is lacking on the role  of water in
transmission of fungal  disease. Gordon et al. (1952) first  reported
the isolation of the spore of H.  capsulatum from river water.
     Experimentally the fungus will  remain viable  as long as 621
days in water (Metzler  et al.,  1956).  The  fungus will grow in ordi-
nary river water.   Ordinary water  purification processes uniformly
removed spores from the w7ater. It was found that  the spores are more
resistant to chlorine than polio virus  or enteric bacteria.  According
to Furcolow (1965), present evidence of transmission by the water
supply  is  not  considered important.  Since  spores can easily be
washed  into streams,  spore  content  in  water  storage  should be
considered in endemic areas.

     Certain  ringworms  are  transmitted  from  animal  to  man
(Bridges, 1963). The ringworms are  considered  as major zoonoses.
Direct transmission  is the method of common  spread.  The fungal
spores remain viable for years in a dry environment (Blood and Hen-
derson, 1968).  The  role  of water in  the transmission of the spores
has not been determined.
     Parasitic  Diseases

     The diseases  associated with helminths and other parasites is
an old science.  Helminths were considered important until the dis-
covery of the  microscope.  Then the era of bacteriology rapidly de-
veloped and  pushed  the  macroscopic  forms of parasites  into  the


 background.  Approximately 60 years ago, with the development of
 tropical medicine, worms again  become  prominent  as  causative
 agents of disease (Cameron, 1962).
     Protozoan  and helminth  diseases  are  widespread in animals
 associated with  agriculture.  Helminths include the trematodes or
 flukes, cestodes or tapeworms, and  nematodes or roundworms.
     Trematodes or flukes are rare in North America except  for
 "swimmers itch" in northern  lakes.  In the United States  cestode
 diseases are not public  health problems of  magnitude. The nema-
 todes  cause  many  diseases in  man and  animals, including  fish
 (Steele, 1968).
     In general, larvae and eggs of parasites are relatively resistant
 to the external environment.  It has been reported (Blood  and Hen-
 derson,  1968) that during  comparatively  dry seasons  and  short
 pasture, dung pats can act as reservoirs for larvae for up to 5 months
 in the summer and 7 to 8 months in the winter.  Under warm and
 wet conditions,  helminth parasites survive in large numbers for as
 long as 6  to  8 weeks, appear relatively resistant  to cold, and may
 survive through the winter.
     Ascarids and the larvae of hookworm may be  contracted from
 water  or soil (Faust et al., 1968). The life cycle of the  fluke evolves
 in a mollusk, usually a snail. The fluke of "swimmers itch" develops
 in a snail.   From a  single egg, thousands of cercariae  emerge in
 water  and attack any warm-blooded animal, including man.  Tape-
 worm  eggs pass through  in feces and all require an intermediate host
 to complete the cycle.

     Balantidiasis is  a protozoan disease  of cosmopolitan  distribu-
 tion, usually observed in warm climates.  It is a parasite of the in-
 testine  and most commonly  found in swine, monkeys, and man
 (Faust,  1963; Van Der Hoeden, 1964).  Human infection results fr^m
 ingesting contaminated food and water. According to Hoare (1962),
 over 90%  of the people are infected  in some countries.

     Toxoplasmosis is an intracellular protozoan  infecting animals
and  man. The disease has a wide host range.  The mode of trans-
mission is not known (Jacobs, 1964). According to  Jacobs  (1956),
despite the sea of toxoplasma infection around us the mode of trans-
mission is still in doubt.

     Up to 200,000 eggs per day are produced by one female ascarid
(Faust et al., 1968).  The eggs are very resistant to cold and survive

                               CHAPTER 19 / DISEASE TRANSMISSION / 279

 most readily in moist surroundings.  Survival up to 5 years has been
 recorded (Blood and Henderson, 1968).

     Strongyloides is a dermatitis developed in trappers, hunters, and
oil workers from  swampy areas of southern Louisiana.  Infectious
larvae of the Strongyloides species infecting swamp-inhabiting mam-
mals were associated with the disease (Burks and Jung, 1960).

     Beef infected with cysticercosis,  the  beef tapeworm, causes
taeniasis  in  man. The tapeworm is  spread  to  cattle by  human
defecation in feed pens and cattle pasture or through distribution of
human sewage and septic  tank effluent to pastures. Researchers in
Great Britain concluded that tapeworm  eggs can survive most urban
and rural human sewage treatment processes  and then pass on in
final effluent or air-dried sludge.  This material, if used on  pastures
or if it finds its ways to streams, can infect livestock (Silverman and
Griffiths,  1955).  In  fiscal  1968 in the United States, 12,723 beef
carcasses were  reported  infected on slaughter  (USDA,   1968a).
Prevalence in man is unknown.

     There is a growing public concern for the environment and the
need for a reevaluation of water's role as a vehicle in transmission
of animal  diseases associated with agriculture.  Documented cases
of infectious diseases of  animal origin  in man and animals have
been associated  with water transmission. Following a literature re-
view, it is apparent that adequate consideration of water transmission
has not been made. In many case reports reviewed, no epidemiologic
studies were made  to determine  the  source  of infection.   In  this
chapter an effort is  made  to indicate the potential epidemiologic
significance based on the variability of the resistant characteristics of
various kinds of pathogenic organisms and their potential for water
     Much  of the past documentation  of water transmission  has
been associated with bacterial agents. The role of animal viruses and
other agents is practically unknown, and with  available methods, the
significance in disease transmission via  water cannot be measured.
Pathogenic organisms of  animals  are found  in  surface  waters, but
for a multiple of  factors, disease only occasionally occurs in man
or animals. Factors involved may be the dilution of water, with a
low  density of  organisms found;  the chain  of  events necessary to
produce the infectious disease process does not develop;  or man and
animals are not exposed.  If the disease does develop, it is not always
diagnosed or reported.


     Although  in  recent years chronic diseases  of  man have been
 of major consideration, the potential of zoonotic diseases through
 occupational and  recreational exposure may be  increasingly signifi-
 cant in the future.  Water is only one of the methods of disease trans-
 mission, but water is essential for life—all animals and man have ex-
 posure to water.
     In the United States the predicted concentration of populations
 of food-producing animals  may better  facilitate the  control  and
 eradication of animal disease by preventive medicine practices rather
 than treatment.  Developing problems,  such as  animal waste dis-
 posal  and the  subsequent environmental effect,  increase  with  live-
 stock  concentration. Future population  growth  and  new develop-
 ments will change occupational and recreational  methods,  and these
 factors will upset the ecologic  systems  in nature that  exist today.
 One cannot predict what will  happen in the future due to  these
 ecologic changes.
     Ecologic  studies of disease in the  environment of nature  are
 filled with  the variabilities of the agent-host-environment complex
 and are difficult to define.  This research needs new approaches.
     The future effect of changing agricultural practices, growth and
 concentration of animal and human populations, and man's increas-
 ing  exposure to water will  effect a challenge  to  all scientific  disci-
 plines to assess the interrelated disease associations.

Abinanti, F. R.  1964.  Respiratory viruses of animals.  In Occupa-
     tional diseases acquired from animals, ed. H. J. Magnuson, pp.
     53-71.  Ann Arbor:  Univ. of Mich. School of Public  Health.
Anderson, G. W.,  Arnstein, M.  G., and Lester, M.  R.  1962.  Com-
     municable disease control. 4th ed. New York:  Macmillan.
Andre, D. A., Weiser, H. H., and Malaney, G. W.  1967. Survival of
     bacterial enteric pathogens in farm pond water.  /. Am. Water
     Works Assoc. 59:503-8.
Babudieri, B. 1959. Q fever a zoonosis.  In Advances  in  veterinary
     science, ed.  C. A. Brandly  and  E.  L. Jungherr, pp. 81—182.
     New York and London: Academic Press.
Bernkopf, H.  1964.  Newcastle disease. In Zoonoses, ed. J. Van Der
     Hoeden, pp. 396—400. Amsterdam, London, New York: Elsevier.
Blood, D. C., and Henderson, J. A.  1968.  Veterinary medicine.  3rd
     ed.  Baltimore: Williams and Wilkins.
Bosworth, T. J.  1934.  Persistence of Brucella on the aborted foetus
     and its membranes.  Univ. of Cambridge, Inst.  of An.  Pathol.,
     Rept. of the Director 4:65-71.
Brackman, P. S.  1964. Anthrax. In Occupational diseases acquired
     from animals, ed. H. J. Magnuson, pp. 216-27.  Ann Arbor:
     Univ. of Mich. School of Public Health.
Bridges;  D. H.  1963.   Fungous  diseases.   In Diseases transmitted
     from animals  to  man, ed. T. G.  Hull,  5th ed., pp.  453—507.
     Springfield, 111.: Charles C Thomas.
Brown, J. R., and McLean, D.  M.  1967.  Water-borne diseases, an
     historical review.  Medical  Services J.  Can., pp.  1011-26.
Bruner, D. W., and  Gillespie, J.  H., eds.   1966.  Hagan's  infectious

                               CHAPTER 19 / DISEASE TRANSMISSION / 281

     diseases  of  domestic  animals.  5th ed. Ithaca:  Cornell Univ.
 Bulloch, W.  1938. The history of bacteriology.  London, New York,
     Toronto: Oxford Univ. Press.
 Burks, J. W., and Jung, R.  C.  1960.  A new type of water dermatitis
     in Louisiana. Southern Med. J. 53:716-19.
 Cameron, T. W. M.  1962. Helminths of animals transmissible to
     man. In Progress of medical science, pathology and bacteriology,
     ed.  R. W. Reed and G. C. McMillan. Am. ]. of Med. Sci.  130/
     354, 157/381.
 Chadwick, J., and Mann, W. N., collaborators.  1950. The  medical
     ivork of Hippocrates. A new translation from the original Greek
     made especially for English readers.  Oxford:  Blackwell Scien-
     tific Publ.
 Chang, S.  L.  1968. Waterborne viral infections and their preven-
     tion. Bull. World Health Organ.  38:401-14.
 Chang, S. L., Buckingham, M., and Taylor,  M. P.  1948.  Studies on
     L. icterohaemorrhagiae.  IV. Survival in water and sewage.  De-
     struction in water by  halogen compounds, synthetic  detergents
     and heat.  /. Infect. Diseases 82:256-66.
 Christiansen, M. J.  1943. Graemarks infektion og kvaegtuberkulose.
     Maadskrift Dyrlaeger 54:241-305.
 Clarke, N.  A., and Chang,  S. L.  1959.  Enteric viruses in water. /.
     Am. Water Works Assoc. 51:1299-1317.
 Decker,  W. M.,  and Steele, J. H.  1966. Health aspects  and vector
     control associated with animal wastes.  Proc. Nat. Symp.  Ani-
     mal Waste Management, pp.  18—20. Mich. State  Univ.,  East
 Diesch, S. L., and McCulloch, W. F.  1966.  Isolation of  pathogenic
     leptospires from waters used for recreation.  Public Health Rept.
 Diesch, S. L., McCulloch, W. F., Braun, J. L., and Crawford, R.  P.,
     Jr.  1969.  Environmental studies on the survival of leptospirosis
     in a farm creek following  a  human leptospirosis outbreak in
     Iowa.  Proc. Ann. Conf. Bull. Wildlife Disease Assoc.  5:166—73.
 Edwards, P.  R., and  Galton, M.  M.  1967.  Salmonellosis.   Advan.
     Vet. Sci. 1:63.
 Errington, P. L.  1963.  The phenomenon of predation. Am. Scientist
 Faust, E. C. 1963.  Infections produced by animal parasites.  In Dz's-
     eases transmitted from animals to man, ed.  T. G. Hull, 5th ed.,
     pp.  433-52. Springfield,  111.:  Charles C.  Thomas.
 Faust, E. C., Beaver, P. C., and Jung, R. C.  1968.  Animal agents
     and vectors  of human disease,  3rd ed. Philadelphia: Lea  and
 Feldman, W. H.  1963.  Tuberculosis. In Diseases transmitted from
     animals to man, ed. T. G. Hull, 5th ed,  Springfield, 111.:  Charles
     C Thomas.
 Fox, J. P.  1964.  Rickettsial diseases  other than Q fever as occupa-
     tional hazards. In Occupational diseases acquired from animals,
     ed.  H. J. Magnuson, pp.  98-109.  Ann Arbor:  Univ. of Mich.
     School of Public Health.
Furcolow, M.  L.  1965.  Environmental aspects of histoplasmosis.
     Arch. Environ. Health  10:14-10.
 Geldreich, E. E.  1965.  Origins  of microbial pollutions in streams.
     In Transmission of viruses by the water route, ed. G. Berg, pp.


      355—61. New  York,  London, Sidney: Interscience  Publishers.
 Gibson, E.  A. 1967. Disposal of farm effluent.  Agriculture 74(4):
 Gillespie, R. W. H.,  and Ryno, J.  1963. Epidemiology of leptospiro-
      sis. Am. ]. Public Health 53:950-55.
 Classman, H. N.  1958.  World incidence of anthrax in man.  Public
      Health Kept. 73:22-24.
 Gordon, M.  A., Ajello, L., Georg, L.  K., and Zeidberg, L. D.  1952.
      Micro sporum gypseum and Histoplasma  capsulatum spores in
      soil and water.  Science 116:208.
 Gratzek, J. B. 1967.  General aspects of viral diseases.  In Veterinary
      bacteriology and virology, ed. I.  A. Merchant and R. A. Packer,
      5th ed., pp. 582—88.  Ames:  Iowa State Univ. Press.
 Green, M.  1965. Major  groups of animal viruses. In Viral and rick-
      ettsial infections of man, ed. F. L. Horsfall and  I. Tamm, 4th
      ed., pp.  11-18.  Philadelphia, Toronto:  J. B. Lippincott.
 Harrell, E. R.  1964. The known and  the unknown of  the occupa-
      tional  mycoses. In Occupational  diseases acquired from  ani-
      mals, ed. H. J. Magnuson, pp.  176-78.  Ann Arbor:  Univ. of
      Mich. School of Public Health.
 Harris, H. J.  1950.  Brucellosis (undulant fever).  2nd ed. New York:
      Paul B. Hoeber.
 Hibbs, C. M., and Foltz, V. D.  1964. Bovine salmonellosis associated
     with contaminated creek water and human infection. Vet. Med.
 Hoare, C. A. 1962.  Reservoir hosts and natural foci of human proto-
     zoal infections.   Acta  Trop. 19:281-317.
 Ignatovich,  V. F. 1959.  The course  of inactivation  of Rickettsia
     burneti in  fluid media.  ]. Microbiol.  Epidemiol. Immunol.
 Jacobs, L.  1956. Propagation, morphology, and biology  of toxoplas-
     mosis.  Ann. N.Y. Acad. Sci. 64:154-79.
 	.  1964.  Actual and potential importance  of protozoal  and hel-
     minth zoonoses  as  occupational  hazards.  In Occupational dis-
     eases acquired  from animals, ed. H. J. Magnuson, pp. 344—43.
     Ann Arbor:  Univ. of Mich. School of Public Health.
 Jellison, W. L., Kohls, G. M., Butler, W. J., and Weaver,  J. A.  1942.
     Epizootic tularemia  in the beaver, Castor canadensis,  and the
     contamination of stream water with Pasteurella tularensis. Am.
    J. Hyg. 36:168-82.
 Jellison, W. L., Epler, D. C., Kuhns, E., and Kohls, G. L.  1950.  Tu-
    laremia in  man from a  domestic rural  water supply.  Public
    Health Rept., pp. 1219-26.
 Jones, T. L.  1963. Diseases of cattle. 2nd ed. Santa Barbara, Calif.:
    Am. Vet. Publ.
 Joyce, G., and Weiser, H.  H.   1965.  Survival  of enteroviruses  and
    bacteriophage in farm pond waters. /. Am. Water Works Assoc.
    pp. 491-501.
 Kalter, S. S.   1964.  Enteroviruses in animals  other than man.  In
    Occupational diseases acquired from animals, ed.  H. J. Magnu-
    son, pp. 126-59. Ann Arbor:  Univ. of Mich.  School of Public
Karlson, A. G. 1967a. The genus Erysipelothrix.  In Veterinary bac-
    teriology and virology, ed. I. A. Merchant and R. A.  Packer, 7th
    ed., pp. 466-74. Ames.- Iowa State Univ. Press.

                               CHAPTER 19 / DISEASE TRANSMISSION  / 283

 	•. 1967b. The genus Mycobacterium.  In Veterinary bacteriology
     and virology, ed.  I. A. Merchant and R. A. Packer, 7th ed., pp.
     441-65.  Ames:  Iowa State Univ.  Press.
 Kraus, P.,  and Weber, G.  1958.  Untersuchungen liber die Haltbor-
     heit   von  Krankheitserregern  intrink-und   oberflacherwasser.
     Zentr. Bakteriol. Parasitenk.  Abt.  I.  Orig. 171:509-23.
 McCroan, J. E.,  McKinley, T. W., Brin, A., and Ramsey, C. H.  1963.
     Five salmonellosis outbreaks related  to poultry products.  Pub-
     lic Health Rept. 78:1073-80.
 McDaniel,  H.  A. 1969. Comparative chronic neurological disorders.
     In Midwest interprofessional seminar on diseases common to
     animals and man. (Abstr.) Ames.-  Iowa State Univ.
 Maddock, E. C.  G.  1933.  Studies on the survival time of  the  bovine
     tubercle bacillus in soil, soil and dung, in dung and on grass,
     with experiments on  the preliminary treatment of infected or-
     ganic  matter  and the cultivation  of the organisms.  J. Hyg.
 Maddy, K.  T.  1967. Epidemiology and ecology of deep mycoses of
     man and animals. Arch. Dermatol. 96:409—17.
 Menges,  R. W.  1963.  A  review and recent  findings on histoplas-
     mosis in  animals.  Vet. Med. 58:331-38.
 Merchant,  I. A., and Packer, R. A.  1961. Veterinary bacteriology
     and virology. 6th ed. Ames:  Iowa State Univ. Press.
 	.  1967.  Veterinary bacteriology and virology.  7th ed.   Ames:
     Iowa State Univ. Press.
 Metzler, D. F., Ritter, C., and Culp, R. L.  1956. Combined effect of
     water  purification processes on removal of Histoplasma  capsu-
     latum from water. Am. ]. Public Health 46:1571-75.
 Middlebrook, G.  1965.  The mycobacteria. In Bacterial and mycotic
     infections of man, ed. R.  J. Dubos and J. G. Hirsch, 4th  ed.,
     pp. 490-521. Philadelphia, Montreal: J. B. Lippincott.
 Morgan, H. R.  1965.  The  enteric bacteria. In Bacterial and mycotic
     infections of man, ed. R. J. Dubos and J. G. Hirsch, 4th ed., pp.
     610-48.  Philadelphia, Montreal: J. B. Lippincott.
 Morse, E. V.  1963. Swine erysipelas. In  Diseases transmitted from
     animals to  man, ed.  T. G. Hull, 5th  ed., pp.  186-209. Spring-
     field, 111.:  Charles C Thomas.
 Mosley, J.  W.  1963.  Epidemiologic aspects of viral agents in rela-
     tion to water-borne disease.  Public Health Rept. 78:328-30.
 National Communicable Disease Center.  1965a.  Leptospirosis.   Zo-
     onosis Surveillance Rept. 7.
	.  1965b. Morbidity and mortality  iveekly rept., vol.  14, no. 22,
     June 5.
 Ogarkov, V. I.  1962. Infectiousness of various objects and materials
     contaminated with Brucella. J. Microbiol.  (Moscow)  4:88.
Parker, R. P., Steinhaus, E. A., Kohls, G.  M., and Jellison, W. L. 1951.
     Contamination of natural  waters and mud with Pasteurella tu-
     larensis and tularemia in beavers  and muskrats in  the  north-
     western United States. U.S. Nat. Inst. of Health Bull. 193. Pub-
     lic Health Serv. 1-61.
Prier, J.  E., and Riley, R.  1965.  Significance of water  in natural
     virus  transmission.  In Transmission of viruses  by the ivater
     route,  ed. G. Berg, pp. 287-300.   New  York,  London, Sidney:
     Interscience Publishers.
 Quan,  S. F., McManus, A. G.,  and  von  Fintel, H.   1956.  Infectivity


     of tularemia  applied  to intact  skin  and ingested in drinking
     water.  Science 123:942-43.
 Reed, R. W.  1965  Listeria and Erysipelothrix.  In Bacterial and my-
     cotic infections of man, ed. R. J. Dubos and J. G. Hirsch, 4th ed.,
     pp. 757-62. Philadelphia,  Montreal:  J. B. Lippincott.
 Rowsell, H. C. 1958.  The effect of stomach contents and the soil on
     the viability of Erysipelothrix rhusiopathiae. }. Am. Vet.  Med.
     Assoc. 132:357-61.
 Ryu, E., and Liu, C-K. 1966.  The viability of leptospires in the sum-
     mer paddy water. Japan. J. Microbiol. 10:51—57.
 Schaal, E.  1963.  Enzootic salmonellosis in a herd of cattle  caused
     by infected brook water.  Deut. Tieraerztl. Wochschr. 70:267-
 Shaughnessy, H. J.  1963.  Tularemia.  In  Diseases transmitted from
     animals to man, ed. T. G. Hull, 5th  ed., pp. 588-604. Spring-
     field, 111..- Charles C Thomas.
 Silverman, P. H., and Griffiths,  R. B. 1955.  A review of methods of
     sewage disposal in Great Britain with  special  reference to the
     epizootiology of Cysticercus bovis. Ann. Trap. Med.  Parasitol.
 Sinha, S. K., Fleming, L. W., and Scholes, S. 1960.  Current con-
     siderations in public health of the role of animals  in relation
     to human viral diseases. /. Am. Vet. Med. Assoc. 136:481-85.
 Steele, J. H.  1968. Occupational health in agriculture.  Arch.  En-
     viron. Health 17:267-85.
 Stein, C.  D.  1942.  Anthrax.  In Keeping livestock healthy.  USDA
     Yearbook of Agriculture, pp. 250-62.
 Sterne, M., and Van Heyningen, W. E.  1965.  The clostridia. In Bac-
     terial and mycotic infections of man, ed. R. J.  Dubos  and J. G.
     Hirsch, 4th ed.,  pp. 454-72. Philadelphia, Montreal: J. B. Lip-
 Stoenner, H. G.  1964. Occupational hazards of  Q  fever.  In  Occu-
     pational diseases acquired from animals, ed. H. J.  Magnuson,
     pp. 36-52. Ann Arbor: Univ. of Mich. School of Public Health.
 Tigertt, W.  D.. Benenson, A.  S., and Gochenour,  W.  S.  1961. Air-
     borne Q fever. Bacterial. Rev. 25:285-93.
 U.S. Dept. of Agriculture.  1965. Losses in  agriculture.  ARS Agri-
     cultural Handbook 291.
 	.  1968a.  Livestock  Slaughter Inspection Division Kept.
 	.  1968b.  Statistical report.
 	.  1969.  Cattle on feed.  Statistical Report, 1 Jan.
 U.S. Outdoor Recreation Resources Review Commission. 1962.  Out-
     door recreation for America, a report to the President and  to the
     Congress. Wash., B.C.
 Van Der Hoeden, J.  1964. Brucellosis.  In Zoonoses, ed. J. Van Der
     Hoeden, pp. 95-132.  Amsterdam, London, New York:  Elsevier.
 Welsh, H. H., Lennette, E. H., Abinanti, F. R., Winn, J. F., and Kap-
     lan, W.  1959.  Q fever studies.  XXI.  The recovery  of Coxiella
     burnetii from soil and  surface waters of premises harboring in-
    fected sheep. Am. ]. Hyg. 70:14-20.
Wilson, J. B.,  and Russell. K.  E.  1964.  Isolation of Bacillus an-
    thracis  from soil stored for  60 years. /. Bacteriol. 87:237.
World  Health Organization.  1962.   Occupational  health  problems
    in agriculture.  Fourth  report of the joint ILO/WHO committee
    on occupational health. Tech. Rept. Ser.  246.

                               CHAPTER 19 / DISEASE TRANSMISSION / 285

    -.  1967a. Current problems in leptospirosis research.  Report of
     a WHO  expert group. Tech. Kept. Ser. 380.
       1967b. Joint FAO/WHO expert committee on zoonoses. Third
    Kept. Tech. Kept .Ser.  378.
Young, L. S., Bicknell, D. S., Archer, B. G., Clinton, J. M.. Leavens,
    L. J., Feeley, J. C., and Brachman, P. S.  1969.  Tularemia epi-
    demic:  Vermont,  1968.  Forty-seven  cases  linked to  contact
    with muskrats.  Nezu Engl.  J. Med. 280:1253-60.

      I HE practice of managing  animal waste to control pollution
 began when  animals were confined.  Today livestock operations tend
 to be more confined and continue to increase in size.  This requires
 a higher degree of waste management. Social attitudes are changing
 the definition of pollution and the degree of acceptability, thus requir-
 ing more waste management.  Taste and color, odors, dust, organic
 and inorganic  matter, plant nutrients, insects, and pathogenic bac-
 teria are all pollutional factors which can result from the mismanage-
 ment of animal waste.
     Management is defined by Webster as "the act or art of planning,
 organizing, coordinating, directing, controlling, and supervising any
 project or activity with the responsibility for  results." Looking par-
 ticularly at animal  waste management this act may be broken down
 into four separate functions: collection,  storage,  treatment, and uti-
 lization or disposal.  Not all systems contain all of the above processes
 and for any one system the order may be changed.  This chapter will
 look at these four steps as they affect water pollution.
     While all  the steps will be discussed separately there is a very
 definite relationship among the functions. In most livestock opera-
 tions, the ultimate utilization or disposal practice will strongly dictate
 the nature of the other processes employed in the waste management
     Manure  varies  in composition and characteristics because of dif-
 ferences in specie, breed, age of the animal,  and the ration. Number
 of animals, geographic locations, climatic conditions, proximity to
 populated areas, and land availability should be considered in select-
 ing a workable and satisfactory management system.

     The collection process can be divided into two types:  wet or dry.
The  dry system can be defined as that which does not add any dilu-

     J. A. MOORE is Instructor, Department of Agricultural Engineering,
     University of Minnesota.


                          CHAPTER 20 / ANIMAl WASTE MANAGEMENT / 287

tion or conveying water to the waste.  Dry systems minimize the vol-
ume of waste material that must be  further processed, while wet
systems utilize the efficiency obtained with liquid-carried transporta-
tion.  The low cost of water and the  efficiency of pumping systems
can make liquid collection very attractive if utilization or disposal of
this additional volume of wastewater is  available.
     In dry systems the manure is usually deposited on the floor, pen,
or under the cage and collected and removed to the next process at
some given frequency. In the open feedlot operation, the manure may
be stored in  a lot for several months before being collected and  re-
moved.  Many dairy operations use mechanical equipment to remove
waste from the  building  on a daily basis.  Gutter cleaners,  shuttle
stroke and endless belt conveyors, powered carts, and small and large
tractors are examples of  some of the mechanical equipment  which
has been developed to reduce the labor required for  collection.
     Flushing gutters have been used successfully in  poultry,  swine,
and dairy operations which use liquid collection  systems. If disposal
is no great problem, clean water can be used for flushing; in other
operations  some treatment can be employed to permit the recycling
of flushing water.  In operations using flushing systems the installa-
tion of impervious channels or conduits is essential.  If  any of this
wastewater is allowed to escape, either by design or otherwise, unde-
sirable conditions result.
     Manure solids which are allowed to settle out on the bottom or
sides of waterways will continually be rewet and  stink, attract flies
and rodents, and  be very unsightly.
     Since  water is being used as a carrier for  the  manure, it is
important that this liquid does not seep into the soil or through  cracks
in the conveying system.  If the above occurs,  the solid  will be left
high and dry and can  create the nuisance conditions mentioned
above.  A loss of water will result in a buildup of solids on the surface
and "polluted" water moving into the soil and eventually the ground-
     Some operations use sloping  bottom  ponds to collect and hold
the wraste slurry for bimonthly flushing.  If water is  allowed to seep
out, the collection process has failed and the above-mentioned  condi-
tions result.
    The development of slatted floor structures has expanded the
use of storage tanks under the housing area to collect and store ma-
nure. The use of fully slatted floors can eliminate the need for labor,
either hand or mechanical, in the collection process.
    In almost all construction the under-the-floor storage tank is  de-
signed  to function as  part of the structural members  of the building
foundation.  These components are almost always concrete  and serve
as a water-tight  storage unit,  thus eliminating any water  pollution.

     Storage may be the first process in the waste management sys-
tem.  Many beef feedlots and poultry operations allow manure  to


 build up and employ only annual or semiannual clean-out schedules.
      In some poultry operations shallow liquid  pits are constructed
 under  the cages and  manure is collected and stored in these for a
 bimonthly flushing. This storage process in liquids for a short period
 of time minimizes odor production and eliminates flies.  When these
 shallow pits become full, the slurry may be flushed to a larger tank
 for some additional  storage or  immediate  removal and disposal.
 These  units are usually concrete lined, which prevents any  deep
 percolation losses.
      Storage of the manure collected in tanks under slat floors may
 extend over long periods of time and  accomplish several purposes.
 Storage tanks eliminate the need for labor in the collection process.
 Feces deposited in the slats by the animals are worked through so the
 livestock operator need exert no energy in getting this waste into the
 tank. These tanks can contain the waste until the land, or some other
 treatment or disposal system, is in condition to accept  the manure.
 The effectiveness of adding this organic matter and plant nutrients
 to the soil is reduced in the winter when heavy snowfall  and freezing
 temperatures are encountered.
      If  animal waste is collected and spread in the winter, the waste
 may actually be stored above the land in a frozen condition until
 spring.  Because of the cold weather experienced in northern climates,
 very Little processing  in the way of  anaerobic  or aerobic microbial
 activity takes place during  the winter.  Depending upon the amount
 of precipitation, slope of the ground, and overland flow from higher
 elevations, this site may continue to serve as storage until the animal
 waste is completely stabilized, leached,  or mechanically  incorporated
 into  the  soil.
     Storage of animal manures is also required after processing in
 some cases.  In operations in which effort and energy are expended
 to reduce the moisture content, storage  can be employed to maintain
 the product in its  postprocess condition. Processes such  as  drying,
 composting, and dehydration reduce the moisture content which gen-
 erally results in lower nuisance levels.  This material  can  then  be
 further processed into feed  or fertilizer  or applied on the land as the
 demand and  time allow.   Storage requires that external  moisture
 sources, such as rainfall and snow, be  kept from the processed ma-
 nure. If  this objective is met, then any water pollution  threat is

     Many may consider processing as a step or method involved in
preparing a marketable product, and treatment as any step or method
involved in stabilizing or reducing waste products.  Here,  treatment
and  processing will be used  to mean  any method involved in  an
attempt to make the product marketable or reduce or stabilize the

                         CHAPTER 20 / ANIMAL WASTE MANAGEMENT / 289
     Dry Systems
     By far the most effective way to minimize the water pollution is
to remove moisture from the manure and then provide safeguards to
eliminate or minimize its subsequent contact with water. The major
treatment systems which remove moisture from the waste are drying,
dehydration (which differ only in the amount of moisture removed),
incineration, and  composting.
     Natural drying is extensively used as a treatment process in the
arid regions of the Southwest. This method is employed because of
the low humidity  and high temperature which  are  encountered in
this area. The very conditions which  allow this system to be  used
also reduce water pollution potential. Dust can be a  major nuisance
in these areas.
     When wet periods in winter or high intensity summer thunder-
showers occur, dikes or catchments can be used to collect and con-
tain the runoff until evaporation can remove the water. By removing
the solids from catchments or sedimentation chambers and frequently
scraping the pens, runoff water quality can be improved. In many
areas in the Southwest continuously running surface waters are not
common,  and generally the  water table is very deep.  These  two
factors  greatly reduce the  water pollution hazard.
     In systems employing dehydration  or incineration, water contact
and subsequent pollution are avoided.  Most operations process the
waste directly from the defecation site.  Since dehydration is relatively
expensive, the resulting product is usually stored  or further processed
without the opportunity to contact and pollute water.
     In  the  incineration process the remaining  ashes possess  little
threat to water pollution when compared to the original  product. In-
cineration can be accomplished without polluting the air,  but this is
an expensive operation  and is not widely used as a  disposal system
for animal  manures.
     Composting is a process of promoting aerobic degradation of or-
ganic wastes in a relatively dry condition.  This can be accomplished
in a pile or windrow, much  the same as you might make leaf or
vegetable waste compost in your backyard, only on a  larger  scale.
This process is usually mechanized and can  take place in  a large re-
volving  drum which may be  heated  and  ventilated.   When  this
method  is employed on  a large scale it is not  a  great contributor to
water pollution.
     Wet Systems
     Water pollution hazards are increased in liquid waste manage-
ment systems.  The relatively inexpensive price of water allows oper-
ators to use liquid systems for the advantage  obtained in the trans-
portation  of this material.  Liquid  systems have  many things in
common with  municipal sewage  treatment plants and  all of the
engineering and biological principles  apply.


      For economic reasons the livestock operator is unable to handle
 and  treat animal waste  in the same manner  as domestic  sewage.
 Actually sewage is about 99%  clean water and  1% waste. To dilute
 animal waste to the same consistency and employ similar treatment
 systems would be economically prohibitive.  It has been reported that
 homeowners pay about 0.9 cents per pound for municipal refuse col-
 lection,  treatment, and disposal, while a similar operation would cost
 the dairyman $200 per cow, per year (Hart,  1964a).
      Since many treatment operations were developed by Civil Sani-
 tary Engineers, we will use their terms to describe three of the basic
 treatment processes which do  apply to animal manure.

     The first treatment process is called primary treatment. In this
 process floating, suspended,  and settleable solids of untreated waste
 are reduced by  sedimentation and  screening.

     Screening.  Screening of animal waste as a treatment process
 has not been used in any commercial livestock operations.  Research-
 ers have  evaluated it  and reported  that dairy cattle waste strained
 through a No.  4 (4.76-mm opening)  sieve removes 50% of the solids
 by weight and 36% of the BOD5 an a 2% solids slurry (Dale and Day,
 1967). A similar study found that only 12%  of the total solids of the
 dairy waste and none of the  solids of chicken waste were held above
 a No. 8  sieve (2.38-mm opening) (Sobel,  1966).  While they do  not
 agree, both  of these studies indicate  that screening can serve to  take
 out some of the undigested corn kernels, hay stems, and silage which
 are common to most feeds. These materials are  relatively inert and
 not amenable to biologic treatment.

     Sedimentation.  Sedimentation  has been and can be  a very ef-
 fective treatment method of  animal waste.  Gravity is the principal
 force causing matter to settle in water. While the principles of this
 phenomenon are well defined and understood, no formula, theoretical
 or empirical, has been devised  that is applicable  to practical sedi-
 mentation-basin design because  of the widely varying conditions oc-
 curring during operation.  Some of  the conditions which  affect  the
 efficiency of the operation  are size of particles (the greater the size.
 the more rapid  is the rate of settling), specific gravity of the particles.
 concentration of the suspended  matter, period of retention, and  the
 velocity of flow  through the basin.
     Sedimentation can serve as  a treatment scheme before a second-
 ary system or be designed to  function at the same time.   Beef cattle
feedlots serve as a good example to employ either or both of these
 systems.   Natural precipitation  that falls  on  or is allowed to  run
through open feedlots becomes polluted and may have a suspended
solids concentration as high as 10,300 mg/1 (Miner et al., 1966).
     Several  studies  (Miner et al., 1966;  Loehr,  1969; Norton  and
Hansen, 1969) have reported the relationships and effects of the in-

                          CHAPTER 20 / ANIMAL WASTE MANAGEMENT / 291

tensity, duration,  slope, etc., on the quality and quantity of runoff
water  from cattle feedlots.
     Laboratory studies have been conducted on animal manure to
determine the settleable matter.  The test is denned in standard meth-
ods, but basically measures the  solid material that will settle from a
1-liter sample  in  60 minutes.
     The suspended  and dissolved solids were found to be a function
of dilution, ration, and  detention time, with dairy  manure settling
from 20% to 95% in 1 hour as the dilution ratio changed from 2:1
to 10:1 (Sobel, 1966). Similar  settling  curves have been plotted for
chicken manure with dilution ration and settling time as the  varia-
bles.  When looking at this method as a  treatment system it is helpful
to realize  that it is the organic nonsettleable suspended solids and the
organic dissolved solids which leave  the settling chamber and exert
the BOD in the effluent.  In beef cattle waste this amounts to 39%  of
the manure added to the system (Ward and Jex, 1969).

     There are two different biological processes which constitute sec-
ondary treatment systems: aerobic and anaerobic systems. However,
it is customary to recognize three major subtypes of energy-yielding
metabolism:  fermentation, aerobic respiration, and anaerobic respira-
tion (Stanier et  al.,  1965).  These three processes are distinguished
from one another by differences  of the ultimate electron acceptor.

     Anaerobic Systems.  Anaerobic respiration  can be  defined  as
those biological  oxidations which  use  an inorganic  compound other
than oxygen as the final electron acceptor.  Nitrates, sulfates, and
carbonates are commonly used as  the electron acceptor by anaerobic
     One of the  main advantages  this  type of system has to offer in
the treatment of animal waste is the high degree of stabilization that
is possible.  Unlike the aerobic oxidation, the  anaerobic  conversion
to methane gas yields little energy to the microorganisms. This low
energy conversion does not support the growth of a  large number of
new cells and the resulting end products are primarily carbon dioxide
and methane gas.
     Since cell growth is slow there is a low production of waste  bi-
ological sludge. Nutrient requirements are low for this type of system
and since oxygen is  not required,  the power  requirements for opera-
tion are reduced.  In many municipal operations methane gas is col-
lected and can be used as a heat source for the waste  digestor, heating
buildings, or  generating  electricity.
     It has been shown that as much  as 90%  of the  degradable  or-
ganics of a waste can be stabilized in anaerobic treatment while only
about  50%  is stabilized in an  aerobic  system (McCarty,  1964a).
While the system  has advantages, the disadvantages begin to weigh
very heavily  when treating wastes with BOD  concentration  of less
than about  10,000  mg/1.


      The major disadvantage of the anaerobic system is the high tem-
 perature required for optimum operation; temperatures about 90° F
 are preferred. While heating the digesters is a common practice  for
 municipality and some industrial wastes, agriculture has yet to make
 widespread use of this technique. Most livestock operators are not
 interested in developing the stalls required to run a good anaerobic
 digester.  This  usually involves a knowledge of mixing  ratios,  pH
 control, etc.
      Using other than oxygen as the electron acceptor results in  the
 production of some foul odors.  With the present public  awareness
 and demand for high environmental quality, the anaerobic systems
 are definitely handicapped because of the odor production character-
 istics.  Because of the low  energy realized in the process, the treat-
 ment is not rapid and requires a longer period  of time for start-up
 and adjustment to temperature and loading changes.
     However, the use of unmanaged anaerobic lagoons to treat ani-
 mal waste is  widespread in this  country. In this sense an anaerobic
 lagoon can be defined as a tank, pit, or reservoir over 5 feet  deep
 which receives  animal waste in some dilute concentration. The 5-
 foot minimum depth eliminates the transfer and mixing  of oxygen
 from the surface by thermal current or wind action.
     Many authors (Hart,  1963; McCarty, 1964b; Hart and Turner.
 1965; Curtis, 1966; Willrich, 1966; Loehr, 1967, 1968; Gramms et al..
 1969; Schmid and  Lipper,  1969) have studied anaerobic  lagoons
 in the laboratory and field for all the major farm animals.  The load-
 ing rates reported for each animal ranged from near zero to 4,000
 chickens, 250 hogs, or 45 cattle per 1,000 cubic feet  of liquid; the
 loading rates  most often suggested were about one-half of these maxi-
 mum values.
     There is no standard measurement by which  all of  the above
 investigators  can  compare results. Each  is  likely to have his  own
 list of objectives and criteria to measure  the success of his project
 and then to project loading rates. Sludge buildup rate is one opera-
 tional parameter which affects the frequency of cleanout.  This is a
 rather costly  operation and in some locations the  use of  additional
 land for a larger or second lagoon will eliminate this cost.
     Lagoons  can be operated on a batch or continuous basis. Gen-
 erally the effluent from an anaerobic lagoon cannot be discharged to a
 surface waterway.  Sprinkling onto pasture or a waste disposal plot
 may provide a final disposal site for excess liquid. Recycle of flushing
 or wash water is  one method of reducing effluent from  the waste
 disposal system.
     Some locations will allow the construction of a lagoon site with
 a designed seepage rate, while other sites  may be required to construct
 an  impervious lagoon.
     Temperature is perhaps the variable which  has  the greatest  in-
fluence  on the performance of an anaerobic lagoon.  Amount of mix-
ing, pH, salinity, detention time, and type of ration fed to the animals
will also affect the operation of the system.

                          CHAPTER 20 / ANIMAL WASTE MANAGEMENT / 293

     Aerobic Systems.  Respiration (aerobic metabolism) is that class
of biological oxidations which utilizes molecular oxygen as the final
electron  acceptor.
     Oxygen is transferred naturally in turbulent flowing streams and
rivers and in shallow ponds and lagoons.  Algae can also be a major
contribution of oxygen to a pond.  However, the  relationship is not
always favorable as algae produce oxygen in the sunlight and con-
sume it at night. If these  natural processes are not sufficient, then
mechanical means  can be employed to provide  additional oxygen.
     With adequate oxygen and the  waste as a food source,  aerobic
bacteria  grow rapidly and  degrade soluble organics very effectively.
In this growth some of the waste is converted to cells, which consti-
tutes a  biological floe.  In final settling this sludge material is  re-
moved and some of it  becomes a solid water product which  creates
the need for another disposal system.

     Oxidation Ponds.  Sufficient oxygen levels can be maintained in
ponds or lagoons limited to about 4 feet deep if  the loading rate is
not too great.  Generally aerobic lagoons are designed to treat 20 to
40 pounds of BOD/acre/day, depending upon location.  Using the 40-
pound rate this is equivalent to 2,600 chickens, 100 hogs, or 30 cat-
tle/acre/day.  These figures were generated by reviewing several of
the articles in the field  (Babbit and Baumann, 1958; Forges and Taft,
1964; Clark, 1965; Jeffrey et al., 1965; Loehr, 1968) and summarizing
the reported results.
     Wind action, temperature, depth, and amount of sunlight will
all influence  the  treatment. If the loading is light or the detention
time is long the effluent may be of sufficient quality to allow discharge
to a surface waterway.  Several lagoons in series  are sometimes em-
ployed to provide treatment that will  allow discharge.
     Depth to water table, soil type, crop,  rainfall, slope of the land,
and water quality will influence final disposal at  this effluent.  In
areas where evaporation is greater than rainfall, a final  liquid dis-
posal system may not be necessary.
     In many treatment systems supplemental air  has  to be provided,
and this can be done with rotors or aerators which strike the  surface
of the water  and increase  oxygen transfer or by employing com-
pressors  and bubbling air  up through the liquid.  While the above
equipment is  expensive, the freedom from noxious  odors may be
worth the price.  Waste digestion  and odor control are factors which
must be considered in a livestock  production operation today.

     Aerated Lagoons.  Some manufacturers of floating aerators have
guaranteed that  their equipment will supply about 3 pounds of oxy-
gen per horsepower hour at standard conditions (Dale et al., 1969).
An operator can  determine the total oxygen demand of his waste load
and  select the necessary equipment.  In design it is  best to  supply
twice the oxygen demand  to  ensure sufficient dissolved  oxygen in
the entire  system.


     Oxidation Ditches.  While floating aerators (vertical shaft units)
 splash, mix,  and reaerate the liquid in a lagoon, rotors (horizontal
 shaft units) are used to accomplish the same functions in an oxida-
 tion ditch.  These units are generally constructed under a slatted floor
 building. As such they are almost always water tight and prevent
 any seepage  losses.
     Most rotors can transfer 1 to 1.5 pounds of oxygen/hr/foot of
 rotor length.  This system holds much promise to contain and  treat
 the waste in  an odor-free environment; several investigators (Irgens
 and Day,  1966; Dale and Day, 1967; McKinney and Bella,  1967;
 Jones et al., 1969; Ludington et al.,  1969;  Moore et al.,  1969)  have
 studied the loading rates and operation characteristics of this system.
     There  are over 100 oxidation ditches in use in this country.  Most
 of these units are in hog operations, with only limited application to
 beef wastes.  The above researchers suggest that 10 ftVhog and  60
 ftVbeef animals are leading rates that can be applied to oxidation
     These  systems can be  operated on a batch basis or with a con-
 tinuous overflow, which requires an additional treatment system.  As
 indicated above temperature and loading rate will influence the
 pollutional  reduction and rate of solids buildup.

     Trickling Filter.  The  trickling filter is  an aerobic system that is
 widespread in the treatment of domestic  wastes.  While it has  been
 demonstrated in a laboratory study that this method can be applied
 to dairy waste,  the system  has  economic and management  require-
 ments which have limited its agricultural application (Bridgham and
 Clayton, 1966).

     Combination Systems. Investigators (Agnew and Loehr, 1966;
 Webster and  Clayton, 1966; Loehr,  1969) have explored  the  advan-
 tages of combining an anaerobic and aerobic process to form  n  com-
 plete treatment system.  It  would appear that some  combination of
 the two can utilize the advantages of both  and provide a good system.
 Field  trials  to date do not  allow the projection  of sizes and loading
 rates  required for commercial  units.

     Secondary treatment systems may have removed up to 90%  of
the original organic matter.  In the event additional treatment is re-
quired, this is called tertiary or third-degree treatment.  Since most
of the solids have been removed and much of the oxygen-dermnd ma-
terials have been oxidized, this  additional treatment may be aimed
at nutrient removal.
     Nutrients, primarily nitrogen and phosphorus, can be responsible
for the growth of algae and  other unwanted plants.  Tertiary treat-
ment is presently being Implemented in a few domestic  waste treat-
ment plants.

                          CHAPTER 20 / ANIMAL WASTE MANAGEMENT / 795

     Like most cities of several years ago, the animal industry is today
thinking about primary  and secondary  treatment and has not  yet
been encouraged to employ tertiary treatment systems.  Nutrients  are
generated in large quantities in  livestock  operations,  and these do
represent a very  real pollution potential.  While application  of  nu-
trient removal systems for animal wastes is  some  distance  in  the
future, proper management of this material can maximize the benefit
of utilization and minimize the pollution from disposal.

     Almost all of the utilization and disposal of animal manures will
be through land application. Some attempts have been made to re-
cover portions of the waste  product for the  drug industry, but these
have generally met with limited success and less  application.  Incin-
eration does a good job of disposing of manure and almost eliminating
all water pollution potential, but cost has  kept this from widespread
     It is not within the scope of this chapter to review or report any
or all of the volumes of work  that have been published  by agrono-
mists, soil scientists, engineers, and others on the effect of animal ma-
nure on soil and crop responses. Many report that animal manure
cannot compete with  manufactured fertilizers and this is very true,
but manure will continue  to be produced and we must look to least-
cost disposal systems which still maintain  our environment quality
if we wish  to continue producing livestock.
     Work done at Rutgers University shows that engineering systems
can be developed to apply liquid manure to the soil (Reed,  1966). This
plow-furrow-cover system employs  equipment which opens  a  plow
furrow, applies up  to 225 tons of liquid manure per acre and then
covers up this material, which  maximizes soil contact and stabiliza-
tion and minimizes environmental  pollution (Reed,  1969).
     Chopper pumps are now available that can  move any material
that will flow to the pump (Hart et al., 1966).  Large rubber nozzles
on sprinkler heads will allow irrigation systems to convey and spread
liquid  manure.  The  old  manure  spreader has  seen  several  new
developments in recent years to increase its capabilities.
     Plow-furrow-cover, like all  other forms of land application, needs
careful review by scientists  from all disciplines to  determine the pol-
lution  effect,  immediate  and  long-range,  on the  surrounding  soil,
water, and  air.
     Techniques are available to collect, store, and treat animal ma-
nure. The one large  remaining task and challenge is to determine
the limits of our environment to accept, utilize, or dispose of animal
wastes. It is wonderful to live  in a country that has  the capabilities
to send men to the moon and back.  It is, however, somewhat disturb-
ing to realize that some  of our people are standing knee deep in
brown gold  to do it.  Time  and effort  will solve the problem; let us
exert the effort and  shorten the  time.



 Agnew,  R. W., and Loehr,  R.  C.  1966.  Cattle-manure treatment
      techniques.  In Management of  farm animal ivastes, SP-0366,
      pp. 81-84. St. Joseph, Mich.: Am. Soc. Agr. Engrs.
 Babbit, H. E., and Baumann, E. R.   1958.  Sewerage and sewage
      treatment. 8th ed. New York: John Wiley.
 Bridgham, D.  O.,  and Clayton, J. T. 1966.  Trickling niters  as  a
      dairy-manure  stabilization component. In Management of farm
      animal wastes, SP-0366, pp. 66-68.  St. Joseph, Mich.: Am. Soc.
      Agr.  Engrs.
 Clark, C. E.  1965.  Hog waste disposal by lagooning.  ]. Sanit. Engrs.
      Div. Am. Soc.  Civil Engrs. 91 (SA6): 27-46.
 Curtis, D.  R.  1966. Design criteria for anaerobic lagoons for swine
      manure disposal. In Management of farm animal ivastes.  SP-
      0366, pp.  75-80.  St. Joseph, Mich.: Am. Soc. Agr. Engrs.
 Dale, A.  C., and Day, D. L.  1967.  Some aerobic decomposition prop-
      erties of dairy cattle manure.  Trans.  Am. Soc. Agr. Engrs. 10
     (4): 546-51.
 Dale, A. C., Ogilvie, J. R., Chang, A. C., Douglass, M. P., and Lindley,
     J. A.  1969.  Disposal of dairy cattle wastes by aerated lagoons
      and  irrigation.  In Animal waste management,  pp.  150-59.
     Ithaca: Cornell Univ.
 Gramms, L. C., Polkowski, L. B., and  Witzel,  S. A.  1969.  Anaerobic
     digestion of farm animal wastes (dairy bull, swine, and poultry).
     Paper 69-462  presented  at annual  meeting of Am. Soc.  Agr.
     Engrs., 22-25  June, Purdue Univ., Lafayette, Ind.
 Hart, S. A. 1963.  Digestion tests of livestock wastes.  /. Water Pol-
     lution Control  Federation 35 (6): 748-57.
 	. 1964a.  Manure management.   Calif. Agr., pp. 5—7. (Dec.)
 	. 1964b.  Thin spreading of slurried  manures.  Trans.  Am.
     Soc. Agr. Engrs. 7(1): 22-28.
 Hart, S.  A., and Turner, M.  E. 1965.  Lagoons for livestock ma-
     nure.  /. Water Pollution Control Federation 37(11): 1578-96.
 Hart, S. A., Moore,  J. A.,  and Hale, W. F.  1966. Pumping manure
     slurries. In Management of farm animal ivastes, SP-0366, pp.
     34-38. St. Joseph, Mich.: Am. Soc. Agr. Engrs.
 Irgens, R.  L.,  and  Day, D. L.  1966. Aerobic treatment  of swine
     waste. In Management of farm animal wastes, SP-0366, pp. 58—
     60.  St. Joseph, Mich.: Am. Soc.  Agr. Engrs.
 Jeffrey, E.  A.,  Blackman, W.  C., Ricketts, R.  1965.  Treatment of
     livestock waste—a  laboratory study.  Trans.  Am.  Soc.  Agr.
     Enqrs. 8(1): 113-17.
 Jones, D. D., Day,  D. L.,  and Converse,  J. C.  1969.  Field tests of
     oxidation ditches in  confinement swine buildings.  In Animal
     waste  management, pp. 160—71.  Ithaca:  Cornell Univ.
 Loehr, R. C.  1967.  Effluent quality from anaerobic lagoons treating
     feedlot waste.  /. Water Pollution  Control Federation 39:384-91.
	. 1968.  Pollution implications  of animal wastes—a  forward
     oriented review. U.S. Dept. of Interior, Fed. Water  Pollution
     Control Admin., Robert S. Kerr Water Res. Center, Ada, Okla.
       1969. Treatment of wastes from beef  cattle  feedlots—field
    results.  In Animal waste management,  pp.  225—41.  Ithaca:
    Cornell Univ.
Ludington, D. C., Bloodgood, D. E., and Dale, A. C. 1969.  Storage

                         CHAPTER 20 / ANIMAL WASTE MANAGEMENT / 297

    of poultry manure with minimum odor.  Trans. Am. Soc. Agr.
    Engrs. (In press.)
McCarty, P. L.  1964a.  Anaerobic waste treatment fundamentals.  I.
    Chemistry and microbiology. Public Works, pp. 107-12. (Sept.)
	.  1964b. Anaerobic waste treatment fundamentals. IV. Proc-
    ess design.  Public Works, pp. 95-99. (Dec.)
McKinney, R. E., and Bella, R. 1967. Water quality changed in con-
    fined waste  treatment. Project Completion  Report, Kans. Water
    Resources Res. Inst., Manhattan.
Miner, J. R., Fina, L. R., Funk, J. W., Upper, R.  I., and Larson, G. H.
     1966. Stormwater runoff from cattle  feedlots.  In Management
    of farm animal wastes, SP-0366, pp. 23-27. St. Joseph, Mich.:
    Am. Soc. Agr. Engrs.
Moore, J.  A., Larson,  R. E., and Allred, E. R.  1969.  Study of the
    use of the oxidation  ditch to stabilize beef animal manure in
    cold  climates.  In  Animal  waste  management,  pp.  172—77.
    Ithaca: Cornell Univ.
Norton, T. E., and Hansen, R. W. 1969. Cattle  feedlot water quality
    hydrology.  In Animal waste management,  pp. 203-16. Ithaca:
    Cornell Univ.
Porges, R., and Taft, R. A.  1964.  Principles and practices of aerobic
    treatment in poultry waste  disposal.  Paper presented at the
    Natl. Poultry Ind. Waste Management Symp., 20 May, Lincoln,
Reed, C. H. 1966. Disposal of poultry manure by plow-furrow-cover
    method. In Management of farm animal wastes,  SP-0366, pp.
    52-53. St. Joseph,  Mich.: Am. Soc. Agr. Engrs.
	.  1969. Specifications for equipment for liquid manure disposal
    by the plow-furrow-cover method.  In Animal ivaste  manage-
    ment, pp. 114—19.  Ithaca: Cornell Univ.
Schmid, L. A., and  Lipper, R.  I.   1969.  Swine wastes,  characteriza-
    tion and anaerobic digestion.  In Animal  waste  management,
    pp. 50-57.  Ithaca:  Cornell  Univ.
Sobel, A. T. 1966. Physical properties of animal manures associated
    with  handling.  In Management of  farm  animal  wastes, SP-
    0366, pp. 27-32.  St. Joseph, Mich.:  Am. Soc. Agr. Engrs.
Stanier, R. Y., Doudoroff, M., and Adelberg, E. A. 1965. The micro-
    bial ivorld.  2nd ed. Englewood Cliffs, N.  J.: Prentice-Hall.
Ward, J.  C., and  Jex, E. M. 1969.  Characteristics of aqueous solu-
    tions of cattle manure.  In Animal waste management,  pp. 310-
    26. Ithaca:  Cornell Univ.
Webster, N. W., and Clayton, J.  T.  1966. Operating characteristics
    of two aerobic-anaerobic  dairy manure treatment systems.  In
    Management of farm animal ivastes, SP-0366, pp. 61—65. St.
    Joseph, Mich.:  Am. Soc. Agr. Engrs.
Willrich, T. L. 1966. Primary treatment of swine wastes by lagoon-
    ing.   In Management of farm animal -wastes,  SP-0366, pp. 70—
     74.  St. Joseph, Mich.: Am.  Soc. Agr. Engrs. -



 T. E.  HAZEN, Leader
 R. 1.  UPPER, Reporter
      IN the moderator's opening remarks,  he mentioned two pos-
 sible areas of discussion that seemed to be suggested from informa-
 tion  presented in the papers of the Wednesday afternoon session.
 One  area concerned use of the soil as the ultimate receptor of either
 treated or  untreated  animal wastes.   The  other was whether the
 state of the art is such that more emphasis on broad-based systems
 analysis is  appropriate.
     After much discussion, it was  apparent that no one wished to
 challenge  the general concept of returning  livestock wastes to  the
 land.  However, there was ample evidence of anxiety over the many
 information gaps that affect intelligent application  of the concept.
 Much discussion was directed toward the  various aspects of nitrate
 as a pollutant.
     Tolerances of man and  domestic  animals to  nitrates in  water
 supplies are not well denned. Needs were expressed for much more
 specific information.  With respect to water for human consumption,
 the expressed need for more definite information on  tolerance  limits
 was countered by the assertion that  public water supplies must meet
 the needs of those with the lowest tolerance. J. E. Box pointed out
 that bio-contamination is always associated  with blue babies.  T. L.
 Willrich gave the history of the development of the present 10 mg/1
 (N) standard by  Comely and cited one controlled  study  under way
 involving  children of various ages in care homes.  It appears to  be
 accepted that the criterion  now  in use for babies  and pregnant
 women is very conservative for normal adults.  Suggestions that the
 pressures  of a growing  population  and increased use of  fertilizer
 may in time indicate the  need for  special drinking water were  in
 sharp contrast with other views—namely,  that increases  in ground-
water  nitrates  cannot be  tolerated  regardless  of  source and that
 algal  blooms must be precluded  in surface waters even if  those
waters are not to be  used  for drinking. A defense was  offered for
algae on the basis that they potentially have the ability to remove
     T. E. HAZEN is Professor, Department of Agricultural Engineering,
     Iowa State  University. R.  I. LIPPER is  Professor, Department  of
     Agricultural Engineering, Kansas State University.

                                 CHAPTER 21  / WORKSHOP SESSION / 299

nitrates and phosphates, and under some circumstances have other
redeeming characteristics.  The assertion that increasingly stringent
water quality standards will be required for the 1980s and  beyond
seemed to imply that little is to be gained now by looking for maxi-
mum  tolerances in  man  as  influenced  by  the many  variables
     A case was made for  the usefulness of more information regard-
ing the influence of nitrates on livestock.
     Monitoring of nitrates in groundwater is being done in numerous
localities.  High nitrates in well waters often  are associated  with
feedlots or rural home waste disposal. High nitrates have been found
in wells at depths of  80 to 90 feet.  W.  H. Walker  said that the
Illinois  State Water  Survey is obtaining background levels  of N in
wells at depths to 100 feet and that NO:! levels as high  as 1,200 ppm
have been found in Illinois water supplies.  Many are in the range of
200  to  300 ppm.  N. J.  Thul said that  the  Kansas Department of
Health  is  sampling wells around large  feedlots. Large seasonal
variations (up to 100 ppm) in nitrate levels of effluent  from field tile
drains under cultivated land were reported  by  Willrich.  The  need
for a broad and extensive interdisciplinary approach to  reconcile
potentially conflicting  demands for clean  water and  a  highly  pro-
ductive agriculture was indicated.
     Other questions concerning the returning of livestock wastes to
land were numerous and varied.  There are several current research
projects in  which very heavy applications of wastes  are  being in-
corporated into soil.  The fate of nitrogen is a common concern, but
the behavior of numerous  other possible water,  soil, and plant  con-
taminants  is also  being  investigated.   Interest  in  management
schemes to maximize nitrogen  losses from the  soil is  evident.  Dis-
cussions regarding the salt content of wastes and its  effect on soil
structure and water intake rate reflect the complexity of the research
needed. Effects on germination and on plant growth and composition
are being given some attention and appear  to require more extensive
examination.  Research relating to these questions was cited in the
Northeast  Region  by P.  E. Schleusener; in Nebraska,  Texas,  and
Colorado by T. M. McCalla; in Kansas by W. L. Powers; in Iowa by
J. Kcelliker; in Georgia by J. E. Box; and in Mississippi  by J. B. Allen.
The  need  for better delineation of objectives  to be  achieved  and
criteria for successful systems was recognized.  The emphasis again
was  on working from a broad interdisciplinary base, develonment of
better regional planning counsels,  and avoidance of parochial  con-
cerns through team  efforts. Acknowledgment of the rural-urban in-
terface is demonstrated by projects concerned with disnosal of urban
wastes on cropland as well as those directed at minimi/ing the assault
of animal waste management practices on the sensibilities of urban
     In  the other major area covered by the discussion, it was ad-
mitted that livestock waste  management  carmot  now  be  planned
with adequate consideration being given to  all  rther  important in-
teracting factors.  Continued emphas's toward  development of svs-
tem  components  was  defended  on the grounds that better  com-


ponents  are required as building blocks for systems.  On the other
hand,  studies of system concepts  can indicate where further com-
ponent development is likely to be most productive and it was  as-
serted  that rudimentary systems analysis would be in order now.
     A question was raised concerning importance of the deficiencies
in characterization of wastes with  respect  to the rations and species
of origin. There  was no disagreement with the answer that there is
an obvious relationship, but it is not known how positive the correla-
tion may be.  The choices are to become more precise or to accept
the heterogeneity and widen the margin for error.
     Other items that were discussed briefly are as follows-,  lagoon
criteria,  performance,  and pollution hazards;  pollution tracers and
indicators; potential for beef confinement feeding systems; and allow-
able cost allocations for agricultural pollution control.





       IN unbiased view of pollution of groundwater from agriculture-
related products would emphasize the fact that numerous rural wells
and springs are polluted. It would also emphasize the fact that only a
very small proportion of rural groundwater is polluted.  These facts
are merely a  starting point for a general assessment of the degree to
which groundwater may be polluted by agriculture-related products.
    An ideal assessment would include an evaluation  of cases of
polluted groundwater in relation to unpolluted groundwater to deter-
mine  specific causes of  pollution. From such  an assessment it is
hoped would  come  simple  and concrete  guidelines or standards to
prevent pollution.
    The following  considerations indicate  that the development of
simple standards for prevention of pollution of groundwater from
agriculture-related products is difficult.

1.  Substances that can become pollutants are numerous and diver-
    sified. (Common potential pollutants include animal fecal wastes,
    fertilizers, pesticides and  associated  chemicals, and  inorganic
2.  The environment  below ground  surface  in which  agriculture-
    related pollutants may occur is complex and generally not easily
    determined.  (A dry, sandy, clay deposit  in a desert  might be
    acceptable for pollutants whereas  a  rocky ground with  a  near-
    surface water table could be unacceptable.)
3.  The distribution of these potential pollutants ranges greatly from
    place to place and  time to time.  (Wastes  from small cow pas-
    tures contrast sharply with  wastes from  large feedlots,  and a
    single pesticide  application on a crop  contrasts with repeated
    application on some orchards.)
4.  The  toxicity and attenuation properties of pollutants range great-
    ly. (Some pesticides in small quantities are known  to be harmful
    to  some wildlife. The attenuation, or weakening tendencies, of
     HAKRY E. LEGRAND is Research Hydrologist, U.S. Geological Survey,
     USDI, Raleigh, N.C.
     Publication authorized by the Director, U.S. Geological Survey.



    each possible pollutant is dependent  on complex factors of its
    environment and on its own inherent characteristics.)

      Much  fruitful research has been done  on the behavior of agri-
 culture-related products in soils, but the movement of these products
 as pollutants downward into  the  groundwater system  has received
 less attention.  The approach taken here is to discuss briefly some of
 the geologic conditions and hydrologic factors  that affect the move-
 ment of pollutants in the ground environment.

     Increasing attention is being focused on the broad spectrum of
 pollution, and the effect of agriculture on environmental  quality is
 continually being assessed. A symposium presented at the meeting
 of the American Association for the Advancement of Science in 1966
 (AAAS, 1967) included a group of papers that discussed the effect of
 agriculture on the quality of our environment. An excellent summary
 report by Wadleigh (1968) discussed wastes in relation to agriculture
 and forestry.  A group of symposium papers discussing the effects of
 pesticides on soil and water was published by the Soil Science Society
 of America (1966). At the request of the President of the United
 States several government agencies  contributed to a report (A Report
 to the President, 1969) on the control of agriculture-related  pollution.
 This latter report listed the following eight pollutants of special con-
 cern:  sediment, animal wastes, wastes from industrial processing of
 raw agricultural products, plant nutrients, forest and crop residues,
 inorganic salts and minerals, pesticides in the environment, and air
 pollution. All of these except  sediment, forest and crop residues, and
 air pollution are especially pertinent to the quality of groundwater.
     Some brief facts indicating the magnitude  of agriculture-related
 pollutants at the land surface  are stated below.
     Animal  Wastes

     The volume of wastes from livestock and poultry production is
estimated at  1.7 billion tons annually. About one-half of this amount
is produced by animals in concentrated production systems. The de-
gree of concentration and the size of individual production units are
increasing rapidly (A Report to the President, 1969, p. 2).  "The daily
wastes from poultry, cattle, and swine alone are equivalent to 10
times the wastes of the human  population of the  United  States"
(Taiganides,  1967, p. 388).
     Plant  Nutrients

     "In 1967, 39 million tons of chemical fertilizers were applied in
the United States and further large increases in use are projected.

                       CHAPTER 22 / POLLUTANTS AND GROUNDWATER / 305

The principal  nutrients supplied were  nitrogen, phosphorus,  and
potassium" (A Report to the President, 1969, p. 4).
     Inorganic Salts  and Minerals

     "Though the presence of dissolved salts and minerals in waters
is universal, their presence in detrimental concentrations is generally
associated with part of the irrigated cropland in arid regions of the
country  and not  with the  relatively  humid  East.  Salinity  from
natural sources stems mainly from the saline characteristics of soils
and from the geologic formations from which  the  soils  are formed.
The salts have not been leached out because of the scarcity of pre-
cipitation.  In agricultural  operations in the  arid part of the nation,
water  is supplied to crops in the necessary  quantities to  sustain
growth.  Concentration  of  the salts occurs in the soil as a result of
water loss through evaporation and transpiration"  (A Report to the
President, 1969, p. 61). The  part of the irrigated water reaching the
water table tends to be higher in dissolved salts than it was originally;
there may be as much  as  a ton of salt per acre-foot of  water,  as is
the case of water from parts of the Colorado River (Thomas,  1956,
p. 551).  Thus, the accumulation of salts in the soil, which may re-
sult in a downward leaching  of the salts into the zone of saturation,
tends to  deteriorate the  quality of groundwater  in some irrigated arid

     "Today, in the United States  8,000 manufacturing firms mix
about 500 chemical compounds into more than 60,000 formulations
registered for use as pesticides.  In  1964, the U.S. chemical industry
produced 783 million pounds of pesticides. . .  ." (Fish, Wildlife, and
Pesticides,  1966, p. 2).  Traces of one  or more chlorinated hydro-
carbons have been reported in every major river system of the United

     If there were no appreciable attenuation, a potential pollutant
could conceivably pass  in sequence through  the following parts of
the environment:  (1) land surface, (2) zone of aeration  (the  zone
between the land surface and the water table), (3) the zone of satura-
tion (the groundwater reservoir) to a stream,  (4) stream course, and
(5) the sea.  Almost never does a pollutant  persist throughout the
sequence  of travel,  and  generally it is  dissipated in the  zone of
     The great variety of potential  agricultural pollutants differ in
their behavior in the ground. The pollutants start to move with water
from precipitation or from solutions containing toxic elements.   Pol-


 lutants in waste solutions are already mobile, but solid wastes must
 undergo  leaching before pollutants from them become entrained in
 subsurface water.  The entrainment  may be retarded, short-lived, or
 complicated by tendencies of pollutants to lose effectiveness by (1)
 decay  or some other inherent power to decrease  potency,  (2) sorp-
 tion on earth materials, and (3) dilution through dispersion and  dif-
 fusion. Assessing the degree to which pollutants  will become at-
 tenuated and  predicting the limits of individual polluted  zones  are
 central objectives.

     In the sense used here decay refers to any of the mechanisms by
 which materials foreign to the ground may be destroyed, inactivated,
 or dissipated as to toxicity. Some pollutants degrade and lose their
 potency with passing time; others  degrade in contact  with  oxgyen,
 particularly on the land surface, in surface water, or in the zone of
 aeration above the water table.  Animal wastes degrade in an oxygen-
 rich environment that favors biological decomposition.  Some pesti-
 cides are  broken  down by microorganisms in  the soil,  but others
 (Alexander, 1967, p. 335) resist biodegradation.

     Although moving in the same direction as water, some pollutants
 move slowly or scarcely at all as they are physically retained by,  or
 react chemically  with,  earth  materials. The extent to which pol-
 lutants are retained depends on the character of the pollutant and on
 that of the earth materials through which they move.  Clays tend to
 retain,  by ion exchange or some other sorptive mechanism, many
 pollutants better than do sands. Dense rocks in which permeability,
 and thus the sorbing surface, is restricted  to  fractures  and solution
 openings have poor sorption characteristics,  and in these  rocks the
 water and the entrained pollutants may move at about the same rate.

     Almost all agriculture-related pollutants mix to  a considerable
degree in water.  Dispersion and dilution are commonly  favorable
considerations, at least  at  a certain  stage or position of pollutant
movement.  However, dispersion is not desirable where dilution is in-
sufficient to lower the concentrations  of certain pollutants to limits
acceptable for organisms that use the water.  For example,  where
concentrated toxic pollutants leak to the ground, consideration may
be given to recovering and containing them before they disperse into
the ground.
     A method of evaluating collectively  all  aspects  of attenuation
has not been developed.  Precise values  for sorption and dilution in

                       CHAPTER 22 / POLLUTANTS AND GROUNDWATER / 307

the environment are difficult to  determine.  Generally we  do not
separate our reliance  on sorption, on dilution, or on  "delay  and
decay" in the ground before the pollutant reaches points of water
use. Yet, a  crude evaluation of each  method of attenuation in each
case of possible pollution might be helpful.

     A potential pollutant at the  ground surface may be considered
to be in a geologic environment of solid earth materials that include a
complex arrangement  of soils and  rocks. It is  also in  a hydrologic
environment that may give it mobility as some water from  precipita-
tion moves into the zone of aeration and down into the zone  of satura-
tion.  Thus, the hydrogeologic setting represents an environment in
which two opposing  tendencies are at work—the tendency  for a pol-
lutant to move with  subsurface water and the opposing  tendency for
it  to be almost immobile or weakened by a  combination of dilution,
sorption on earth materials, or  some "die-away" mechanism.  The
great range in geologic and hydrologic conditions prevents good rule-
cf-thumb techniques for determining the safe  distribution of agricul-
ture-related products at the land  surface.
     The soil zone is the "action zone" where fertilizers,  manure, and
pesticides may start to become pollutants of groundwater.  It is the
action zone for biodegradation and other attenuation methods.  The
chemical and biological character,  texture,  permeability, and thick-
ness of the soil zone are important features.
     Beneath surface soils in some places  are  unconsolidated  sedi-
mentary materials of clay, silt, and  sand. In other places hard, dense
rocks underlie  soils.  Rocks at considerable depth may not be sig-
nificant because  they  lie below  the  paths  of most ground-surface
     Permeability is  an important  characteristic because it controls
the rate of movement  of water and pollutants that might be with it.
The permeability of  some  clays may be many hundreds  of times less
than that of some sands. Zones of greater permeability tend to paral-
lel, or coincide with, rock formational boundaries even if  the rocks
are appreciably inclined.  Differences of permeability in the horizontal
field, although  common, are in many  cases more gradual than in the
vertical field. The point to be made is that water and included waste
will tend to take preferred paths, flowing readily through permeable
zones  and  shunning  or  flowing with  difficulty through  relatively
impermeable materials.
     The water table is an important consideration of  groundwater
pollution, especially  in view of the ease of  attenuation  of  most pol-
lutants xvhere the water table is  deep and where the overlying  zone
of aeration is composed of sands, silts,  and  clavs. The  frequency of
precipitation in humid regions is sufficient  to keep the water  table
relatively close  to the ground surface  in  areas of moderate  per-
meabilities, and the consequent  mounding  of water beneath inter-
stream areas causes  a  continuous subsurface flow of water to nearby


 perennial  streams.  Thus one can get a general idea of the gross
 direction of movement of groundvvater in humid regions; in arid re-
 gions, however, the areas of natural groundwater discharge are more
 widely scattered,  and the general movement of water may be  less
 discernible. In arid regions some  reaches of most streams lose water
 —that is, water  from the streams  may  seep into the ground as
 opposed to the gaining type of stream in humid regions.
     In many  cases  of pollution, the movement of water  in  the
 ground has been altered  by  man's activities, such  as  pumping of
 wells or adding liquids to the ground. Pumping of a well causes a
 cone of depression on the water  table, resulting in a flow of water
 toward the well from the surrounding area.   Opposite hydrologic
 conditions result when liquids are added to the ground in one place,
 as a mound on the water table is  developed and groundwater moves
 outward from the spot. Knowledge or inferences about earlier condi-
 tions may guide decisions about  remedial action on some  pollution

     Fertilizers and pesticides are spread usually over the  land sur-
face in their conventional use, and occasional applications for both,
rather than continual applications,  are the rule.  Both the lack of
concentration and the lack of continual application tend to weaken
the ability of these possible pollutants to move downward with in-
filtrating water through the soil  zone or through the  entire zone of
aeration to the water table.
     Of the fertilizer nutrients, phosphate and nitrate  are the ions of
chief concern as to possible pollution of water resources. Phosphorus
tends to be sorbed by  soils so well that it is rarely a  serious threat.
Nitrate is a common  constituent in  groundwater, generally in pro-
portions of no more than a few milligrams per liter.  In  fact, the
average sample of groundwater  in the humid  southeastern part of
the United States has less than one milligram per liter of nitrate. Yet,
in local areas and in certain groundwater systems the nitrate content
averages several milligrams per liter.  It may originate from natural
sources, livestock feeding operations, sewage disposal systems, legume
residues, manures, or excessive use of chemical fertilizers.  It is dif-
ficult to single out the source of nitrate in groundwater, but Smith
(1967, p.  184) points  out that  leachates from  highly fertile,  un-
fertilized  agricultural  lands may have a higher content  of plant
nutrients  than the percolates from nearby fertilized, well-managed
cropland low in natural fertility.  Nitrate and chloride are good pre-
cursors of pollution, and an increase in these ions with time may aid
planning in avoiding serious pollution. With  the exception of isolated
cases,  there is little evidence to support  statements that  fertilizer
nutrients are polluting  water supplies (Smith, 1967, p. 185).  A special
studv of a fertilized terrain in southwestern Wisconsin (Minshall et
al., 1969,  p. 713) confirms this view;  it was found  after 2 years of
data collecting that water of streams during their low-flow period
(representing outflow of groundwater) in this region appeared to be a

                        CHAPTER 22 / POLLUTANTS AND GROUNDWATER / 309

 relatively unimportant carrier of plant nutrients.  With excessive and
 improper use, however, in the future nitrates could become a problem
 when excess nitrogen is added to some soils.
     Analyses of groundwater for pesticide content are relatively rare;
 most of the analyses are related to local research programs not yet
 completed or are from  isolated samples of water from wells near
 places where concentrated pesticides may have spilled to the ground.
 The absence of a systematic sampling  and monitoring program of
 groundwater is probably based on the assumption that the soil zone
 and zone of aeration are effective in attenuating pesticides above the
 water table.  The work of  Sheets  (1967),  Alexander  (1967), and
 other workers indicates the tendency for the bulk of  pesticides  to be
 degraded, volatilized, or fixed on soils. That microbial decomposition
 of pesticides is less rapid in subsoil than in surface soil appears to be
 a valid assumption that should be investigated (Sheets, 1967, p.  322).
 Even in the subsoil,  sorption is still effective, as indicated by research
 on DDT (Scalf et al., 1968).
     Some groundwater samples from  Arkansas  and Mississippi in
 the Mississippi River Delta were analyzed  for pesticide residues as a
 result of a  research project  undertaken by the USDA (ARS 81-13).
 This study reports analyses made in 1964  and indicates  that most of
 the well water  sampled contained  no  detectable  pesticide  residue.
 However, detectable residues were  identified in a few samples, gen-
 erally in quantities of only a fraction of a microgram  per liter.  At
 the time the report  was completed, the presence of  the residues in
 the well water was not explainable.  It should be noted that none of
 the wells contained  pesticide  residues  throughout  the year.   This
 study of pesticides in the Mississippi River Delta  serves to show the
 difficulty of evaluating the possibility of contaminating groundwater.
 Iverson (1967, p. 161) reports that analyses "of hundreds of samples
 of water have produced evidence that neither deep nor shallow  wells
 are being contaminated  by insecticides if  the well is constructed in
 such  a  manner as  to provide  water fit for  human consumption."
 Evidence  from  different sources  suggests the general  freedom  of
 groundwater from pesticides, but a monitoring program to  determine
 the distribution of pesticides in groundwater seems justified.
     There  are certain conditions  that  could readily lead  to  local
 pollution of groundwater by pesticides and related chemicals. Where
 such materials are dumped on the  ground in concentrated form, es-
 pecially near  shallow wells  or in  areas where  the  soil is thin  or
highly permeable, pollution of groundwater could be serious.   Soils
 are thin in  many areas  underlain by limestone,  where rolling  sink-
hole topography results in  quick  drainage  of surface water into
caverns.  In some limestone terranes the groundwater moves rapidly
 to  streams, and  pollutants have  little chance  to  be  attenuated
(Deutsch,  1963, p.  33). Attenuation  of pollutants is  much  better
where the pollutants are in contact with the ground only occasionally,
as with agricultural pesticide use.
  Although there is  a tendency for animal organic wastes  to become
more localized  each  year, both animal and  human wastes in  rural
areas are  much more dispersed than those in urban areas.  Thus,
unlike procedures in urban areas where organic wastes are generally


contained,  diluted with water, transported,  and treated,  the  waste-
handling procedures in rural areas result in wide distribution to the
ground; the percentage of water added to wastes in rural areas is
generally much less than that in urban areas.  It is difficult to assess
the potential of rural organic wastes to pollute water. Overland run-
off can leach wastes and result in stream pollution.  There is vertical
leaching into the ground environment of animal and human wastes.
Gillham and Webber (1968) reported a  significant increase in  the
nitrogen content in the groundwater as it passed beneath a barnyard.
Two counter  tendencies prevail.  The  tendency for pollution  from
leached wastes to move downward and to become entrained with the
subsurface  water is  mostly offset by the tendency of the waste mate-
rials to be attenuated by degraduation in the  soil, by sorption, and by
dilution.  Hence, pollution  of groundwater  is  less common  than
might  be expected in view of the widespread occurrence  of surface
contaminants.  Yet,  serious problems do exist.  The potential for pol-
luting  groundwater  is great where  concentrated wastes,  as at  feed-
lots, are exposed to  thin soils  on cavernous limestone formations or
to thin or sandy  soils on fractured rocks.
     Where pollutants  escape  attenuation by sorption  and decay in
the zone  of aeration, attenuation in the  underlying zone of satura-
tion may be chiefly by dilution in groundwater. As might be expected
in the groundwater  reservoir,  the  polluted  zone is normally  more
     FIG. 22.1.   Generalized   block-diagram,  showing  isolated  polluted
     zones at the land surface (dots in block on right) and the relative  ex-
     tent to which pollution is carried downward to the zone of saturation
     in block  on  left. Sites A, B, and C represent  pollution concentrations,
     such as feedlots, great enough  to  pass through the  zone of aeration
     and along the top of the zone of saturation for some distance toward
     the surface stream.

                        CHAPTER 22  / POLLUTANTS AND GROUNDWATER / 311

pronounced at the water table than at greater  depths, and the pol-
luted zone tends  to  be elongated in  the  direction of groundwater
movement.  Patterns of polluted zones on the water table have been
described schematically by  LeGrand (1965) and are shown in Figure
22.1. Although zones of pollution from agriculture-related products
have rarely been described,  their patterns on  the water table  are
similar to patterns formed by industrial pollutants. A large but com-
monly shaped polluted zone (Fig. 22.2) resulted from  waste-disposal
practices at a chemical factory in Colorado (Walker, 1961); the map
shows the movement of chlorates and 2,4-D-type compounds, as well
as the anticipated  area of influence from waste basins. Very rarely
would a contaminated  groundwater  zone in agricultural  areas be
as large as that shown in Figure 22.2, but the "down-gradient" shape
is typical.
     The volume of groundwater polluted by plant nutrients, animal
wastes, and  pesticides appears  to be  small.  Admittedly, there are

   Anticipated  area  of influence

   Chlorate  toxicity

   2, 4-D-type toxicity

   Waste basins

— Contours  on water
   (feet above sea
     FIG. 22.2.  Patterns of polluted groundwater formed from seepage of
     chemicals from waste  basins.  (Modified after Walker,  1961, p.  492.)


 numerous cases of farm wells being polluted, and numerous small
 polluted zones of water occur in the upper part  of the zone  of
 saturation. Sufficient safeguards are available to minimize ground-
 water pollution to the extent that good agricultural practices should
 not be deterred.
     The zone  of  aeration above the water table, which normally
 contains in its upper part the soil zone, attenuates almost all of the
 foreign bodies  that are potential pollutants of the underlying ground-
 water. Chemical fertilizers, animal wastes, and pesticides vary great-
 ly in  their tendency to  degrade in ground  environments. They all
 degrade better under a set of hydrogeologic conditions.  The follow-
 ing environmental factors  tend to reduce  the chances of pollution
 of water from  wells and springs:

 1.  A  deep water table, which (a) allows for  sorption  of pollutants on
    earth materials, (b) slows subsurface movement  of  pollutants,
    and (c) facilitates oxidation or other beneficial "die-away" effects.
 2.  Sufficient clay  in the path  that pollutants will move  so that re-
    tention or sorption of pollutants is favorable.  (However, excessive
    clay may result in poor  surface permeability, thereby allowing
    much water and pollutants to move overland to surface streams.)
 3.  A  gradient of the water  table beneath  a  waste  site away from
    nearby wells.
 4.  A  great distance between wells and wastes so that  advantages  of
    the above factors can accumulate.

     Dispersion has been a major factor in minimizing the pollution
 of groundwater in agricultural regions of the United  States. In their
 conventional uses  both  fertilizers and pesticides  have been widely
 but  thinly applied. Both  human and animal wastes have  caused
 only minor pollution problems until recent  years, but the increasing
 concentrations of  animal wastes  in  large  feedlots  is a matter  of
 growing concern.  Disposal of containers  of pesticides  and other
 toxic chemicals in  rural areas by design or  accident will pose ques-
 tions of the possibility of groundwater pollution.
     Soils maps and results of hydrogeologic studies should furnish
 a good background for evaluating the potential of certain  agriculture-
related products to pollute groundwater. Yet no magic or simple
quantitative system for predicting accurately the fate  of the variety
of pollutants in the ground  environment is likely to  be developed
soon.  Although extremely  unfavorable ground  conditions are easy
to determine, most earth materials have the capacity to attenuate
pollutants to some degree.  The exercise of good judgment in manag-
ing  agriculture-related  products  that  can  become  pollutants  of
groundwater is essential.

Alexander, Martin.  1967.  The breakdown  of pesticides in soils.  In
     Agriculture  and the  quality  of our  environment, ed.  N.  C.
     Brady, pp. 331-42.  Norwood,  Mass.:  Plimpton Press.

                       CHAPTER 22 / POLLUTANTS AND GROUNDWATER / 313

American Association of  Advancement of Science.   1967.  Agricul-
     ture and the quality  of our environment, ed. N. C.  Brady, Nor-
     wood, Mass.:  Plimpton Press.
Deutsch, Morris. 1963. Ground-water contamination and legal con-
     trols in Michigan.  U.S. Geol. Survey Water-Supply  Paper 1691.
Gillham, R. W., and Webber,  L. R.  1968. Groundwater contamina-
     tion. Water Pollution Control 106 (5).
Iverson, L. G. K. 1967. Monitoring of pesticide content in  water in
     selected areas of  the United  States.  In Agriculture  and  the
     quality of our environment, ed. N. C. Brady, pp. 157-62.  Nor-
     wood, Mass.:  Plimpton Press.
LeGrand, H. E. 1965.  Patterns of contaminated zones  of  water in
     the ground. Water Resources Res. 1 (1): 83-95.
Minshall, N. M., Starr,  Nichols, and Wetzel, S.  A.  1969. Plant nu-
     trients  in base  flow of  streams in  southwestern Wisconsin.
     Water Resources Res. 5 (3): 706-13.
Scalf, M. R., Hauser, V.  L., McMillion, L. G.,  Dunlap, W. J., and
     Keeley, J.  W.   1968.  Fate of DDT and nitrate in ground water.
     Robert S. Kerr Water Res. Center, Ada, Okla., and Southwestern
     Great Plains Res. Center,  Bushland, Tex., Spec. Publ.
Sheets,  T. J.  1967.  Pesticide buildup in soils.  In Agriculture and the
     quality of our  environment, ed. N. C. Brady, pp. 311-30.  Nor-
     wood, Mass.:   Plimpton Press.
Smith,  G. E.  1967. Fertilizer nutrients in water supplies.  In Agri-
     culture and the  quality of our environment, ed. N. C.  Brady,
     pp. 173-86. Norwood, Mass.: Plimpton Press.
Soil Science Society of  America.  1966.  Pesticides and  their  effects
     on soils and water. ASA  Spec. Publ.  8.
Taiganides,  E. P.   1967.  The animal waste disposal  problem.  In
     Agriculture and the quality of our environment,  ed. N. C.  Brady,
     pp. 385-94. Norwood, Mass.: Plimpton Press.
Thomas, Harold E. 1956.  Changes  in  quantities and  qualities of
     ground and surface waters. In Man's role in changing the face
     of  the  earth,  ed. William L. Thomas, pp. 542-63.  Chicago:
     Univ. Chicago  Press.
U.S. Dept.  of Agriculture. 1966.  Monitoring agricultural  pesticide
     residues. ARS-81-13.
U.S. Dept. of Agriculture. 1969.  Control of agriculture-related pol-
     lution.  A  report to  the  President.   Submitted  by  the Sec.  of
     Agr. and the Dir. of the Office of Sci. and Technol.
U.S. Dept. of  Interior.   1966. Fish,  wildlife and pesticides.  U.S.
     Fish and Wildlife Serv.  Unnumbered pamphlet.
Wadleigh, C. H.  1968. Wastes in relation  to  agricidture  and  for-
     estry.  USDA Misc. Publ.  1065.
Walker, T.  R.   1961.  Ground-water  contamination in the  Rocky
     Mountain  arsenal area. Denver, Colorado.  Eidl.  Geol. Soc. Am.



       £• UTROPHICATION refers to the process of enrichment  of wa-
 ter with nutrients  (Stewart and Rohlich,  1967).  An  obvious effect
 of eutrophication is an increase in the biomass which can be sup-
 ported in a body of water. Although the increase in yield of  a crop
 after fertilization is desirable in terrestrial situations, the effects  of
 eutrophication  of  waters  are   often  undesirable.  Generally the
 aesthetic  value of  a lake is  lowered through  excessive  growth  of
 aquatic weeds  and algae and  production of  floating algal  scums
 which are a nuisance to those who use the water for recreational pur-
 poses. Other effects  include undesirable  odors and tastes, and im-
 pairment of water treatment operations—for example, through clog-
 ging  of filters  by  algae.
     It should be recognized  that  lake eutrophication is  a natural
 process of  lake maturation.  Precipitation and natural  drainage
 contribute  nutrients  which  support  and  enhance the  growth  of
 phytoplankton and  littoral vegetation.  However, the acceleration
 of eutrophication as a result of man's activities in altering the land-
 scape  through  agricultural development,  urbanization, and  waste
 discharge is  of major concern.
     While lake eutrophication involves enrichment with  nutrients,
 the stage or  rate of lake eutrophication is not controlled solely by
 the quantities  of nutrients present or entering  the receiving body
 of water. The interrelationships of climatic, physical, chemical, and
 biological factors which affect lake metabolism are highly complex.
 As illustrated by Rawson (Fig. 23.1),  the morphology of the  basin,
 geological characteristics  of  the  area, temperature, nutrient  input,
 and many other factors influence the metabolism of a lake. Because
 of the  complex interrelationships involved,  establishing reliable
measurements  of lake eutrophication rate and stage has been  a
major problem (Fruh et al., 1966).
    Interest  in control of eutrophication  has  focused on limiting
     D. E. ARMSTRONG is Assistant Professor of Water  Chemistry, Uni-
     versity of Wisconsin.  G.  A. ROHLICH is Director, Water Resources
     Center, and Professor of Sanitary Engineering, University of Wis-


                           CHAPTER 23 / POLLUTION AND EUTROPHICATION / 315
                            .Geographic Location_

                              ' Altitude

                of Substrate
                Shape of Basin
Primary Nutritive /^Drainage Area  Depth
                    Light pene-
Heat Penetration
and Stratifica-
                                    Oxygen  Penetra-
                                    tion and utili-
                                 of Littoral
Nature of •*— Inflow of
 Bottom    Allochth.
Deposits    Materials
 Seasonal Cyclo
Circulation . Stag-
 nation , Growing
                        'Trophic  Nature  of the  Lake'
                     Amount, composition and distribution of plants
                       and animals.  Also rates of circulation.
                              "l-l  J   1' 'I  "
        FIG. 23.1.   Chart suggesting the interrelations of factors affecting  the
        metabolism of a  lake.  (Rawson, 1939.)

  the  amounts of nutrients entering the water.  The success of this
  approach depends on whether the available nutrient supply can  be
  reduced to  the extent  that  growth  of  aquatic  plants is  limited.
  Nutrients which have received the most  attention are nitrogen and
  phosphorus  because,  following  carbon,  they are  required in  the
  greatest amounts  for the production  of green plants.
       Importantly, the amounts of nitrogen and phosphorus  available
  to aquatic plants in  lakes depend not only on the amounts entering
  the body of water but also on the chemical,  biochemical, and physi-
  cal processes occurring within the lake  as  shown in Figure  23.2
  (Armstrong et  al., 1969).  The  available nitrogen and  phosphorus
  pool (mainly  the dissolved inorganic  nitrogen and phosphorus com-
  partment) is  regulated  by a number  of interrelated processes. For
  example, uptake or release of available  nutrients by the bottom sedi-
  ments may occur, depending on sediment properties and environ-
  mental conditions.  Microorganisms  may compete with  plants for
  available nutrients.  It should  be emphasized  that both quantities in
  compartments and rates of interchange  among  compartments  are
  important.  For  example, rapid exchange of nutrients  between  the
  sediments  and  water  might  supply  sufficient  quantities  for  plant
  growth even  at  low  concentrations of  nutrients  in  the  lake water.
       The forms and chemistry of nitrogen and chemistry of nitrogen
  and  phosphorus in soils have been discussed previously and will be

      FIG. 23.2.  Major components of the nitrogen and phosphorus  cycles
      in lakes.  (Armstrong et al., 1969.)
 reviewed only briefly here (see review by Biggar and Corey, 1969).
     Most of  the nitrogen in soils (perhaps more than  95% of the
 total soil nitrogen) is organic. Much of the organic nitrogen (about
 50%) is present  in amino  form.  The main inorganic forms are
 nitrate and  ammonium; nitrite  is usually present only  in  small
 amounts, though a  small portion of the total soil nitrogen, nitrate,
 and ammonium is of primary importance because it is  in  the  form
 of nitrogen utilized by plants. Quantities of ammonium and nitrate
 depend mainly on the processes of organic nitrogen mineralization
 and inorganic nitrogen immobilization, and soil  organic nitrogen or
 organic  matter contents provide a  good indication of  the nitrogen
 fertility  of the  soil.
     Phosphorus in soils exists in inorganic and  organic forms. The
 inorganic phosphorus content varies from  about 25  to  97% of the
 total and is in the range of  50 to 75%  for many soils.   Total phos-
 phorus ranges from  100  to 2,000 ppm and is often about 1,000 ppm.
 Dissolved  inorganic phosphorus is the form directly  available to
 plants but organic  phosphorus is available through  conversion to
 inorganic  phosphorus.   The  amount  of dissolved  inorganic  phos-
 phorus in the soil  solution is low,  usually about 0.01  to 0.1 ppm,
 due to adsorption of phosphorus by the iron, aluminum,  and calcium
components of the  soil.
     Water reaching the soil surface  is disposed of by  (1) surface
runoff.  (2) groundwater runoff (interflow), (3) deep  percolation,

                       CHAPTER 23 / POLLUTION AND EUTROPHICATION / 317

 (4) storage, and (5) evaporation and transpiration (Biggar and Corey,
 1969).  Of  these, the first three—namely, surface runoff,  ground-
 water seepage,  and percolation  to perched  water tables  or deeper
 aquifers—contribute to eutrophication by transporting  nutrients to
 streams and lakes.  Surface runoff may  directly enter streams  and
 subsequently lakes.  Some of the water  that enters  the soil drains
 downslope and may reappear at a lower elevation as surface water
 or  seepage.  Water  percolating  to the groundwater may transport
 nutrients to rivers and lakes which receive a major portion  of their
 water from groundwater flow. According  to Biggar and Corey (1969),
 irrigation, which involves  a recycling of  water derived from runoff,
 seepage, and percolation,  often  increases the amounts  of nutrients
 transported to  lakes and streams by these waters.
     The  amounts of nutrients transported in agricultural drainage
 are determined in part by the chemical forms of the nutrients  and
 the processes controlling their retention  in  the  soil.  Runoff  water
 carries  nutrients in both  dissolved   and particulate forms,  while
 water percolating through the soil generally carries only dissolved
 forms.  Because inorganic phosphorus  is retained  more  strongly
 than  inorganic  nitrogen  by  soil particles,  the  forms  of nitrogen
 and phosphorus  transported differ appreciably  for runoff and  per-
 colate waters. Ammonium and particularly nitrate are quite soluble
 and tend  to move  downward in  the soil with percolating wrater,
 thereby lowering the amounts at the  soil surface.  Since runoff  wa-
 ters tend to transport forms located near the soil surface, ammonium
 and nitrate are carried in runoff waters in dissolved  or  paticulate
 form to a lesser  extent than are the  more insoluble  nutrients.  Due
 to the low anion exchange capacity of soils and the high solubility
 of  nitrate,  the  downward movement of nitrate  with percolating
 waters is quite rapid. Thus, the extent to which nitrate is  leached
 depends to a  large  extent  on  the  quantity of  water percolating
 through the soil and the degree to which  nitrate levels are in excess
 of plant and microbial needs. Although  ammonium is soluble,  the
 downward movement of  ammonium  is  retarded by  retention at
 cation exchange sites. Furthermore, conversion of ammonium to
 nitrate in soils through nitrification is generally quite  rapid.
     Inorganic phosphorus tends to be strongly retained by  soil  par-
 ticles, and phosphorus received  by the soil as commercial fertilizer,
 plant residue, and manure tends  to remain at the soil surface, there-
 by  enhancing the possibility of  transport by runoff in particulate
 and soluble forms.  Biggar and  Corey (1969) have  suggested  that
 due to the low mobility of phosphorus in soils,  application  of phos-
 phorus  to the soil surface will tend to saturate the  phosphorus ad-
 sorption sites and cause the concentration of phosphorus in  solution
 near the soil surface to be relatively  high.  Some  of  the phosphorus
 in solution  would tend  to move downward  but would be rapidly
 adsorbed at the undersaturated adsorption sites beneath the  surface.
 However,  at the surface, phosphorus in solution would be maintained
 at a relatively high concentration and the phosphorus concentration
 in runoff  waters in  contact with these surfaces might be relatively
high.  Whether the dissolved  phosphorus would remain in  solution


 would  depend  on the phosphorus adsorption capacity of  the  sus-
 pended soil  particles  and  stream sediments in  contact  with the
 runoff  water.
      Both organic nitrogen and organic phosphorus as well as inor-
 ganic phosphorus are of low mobility in soil and are likely trans-
 ported  to a large extent  in particulate form in runoff waters. How-
 ever, because the amount  of organic nitrogen  in soil is  high as
 compared to inorganic nitrogen, quantities of soluble organic nitro-
 gen transported may be  significant relative to amounts of inorganic
 nitrogen.  Particulate forms  are generally of less  interest  than dis-
 solved  forms regarding their effects  on the receiving  water due to
 the lower plant availability of these  forms and  the possibility  that
 the particulate material will settle to  the bottom of streams or lakes.
 However, it  should  be recognized that  the  new  environment, for
 example an  anaerobic  lake  bottom, may  markedly  increase  the
 mobility of  nutrients  contained in  these particles.  On  the other
 hand, eroded soil particles transported to streams  or  lakes  may de-
 crease  the available  nutrient  supply in  the  water.  For  example,
 phosphorus-deficient soil particles entering a lake may remove phos-
 phorus  from solution by  adsorption and transport  the  adsorbed
 phosphorus  to  the lake  bottom.
     In summary, it  is generally expected that inorganic nitrogen is
 transported mainly as nitrate  by percolating waters,  although the
 amounts of ammonium and nitrate carried in runoff waters may be
 highly  significant  in terms of  the receiving water.  Similarly, the
 largest  amount  of phosphorus  is  likely  transported  in particulate
 form in runoff  waters, but  the  amount of dissolved phosphorus in
 runoff water  may be of equal  or  greater importance even  though
 lower in quantity.  Obviously these statements are highly generalized
 and will not apply in many situations. An important example is the
 situation in which the soil is frozen.  In  this case, soluble  and  par-
 ticulate  forms of both nitrogen and phosphorus would  be carried in
 surface  runoff.
     Because  concern over the  quality and the nutrient content of
 agricultural  drainage has  developed  only recently, relatively  few

 TABLE 23.1.  Nitrogen  and  phosphorus content of waters in surface and sub-
            surface drains and in shallow wells.
Constituent Surface Drain Subsurface Drain
Dissolved P . . . .
Total P 	
Dissolved P . . . .
Total P 	

Irrigation Season
Nonirrigatioji Season
Shallow Well
Source:  Sylvester and Seabloom (1962) as reported by Biggar and Corey

                        CHAPTER 23 / POLLUTION AND EUTROPHICATION  / 319

 TABLE 23.2.  Constituents in  runoff  waters from a  1.45-acre winter wheat
             field near Coshocron, Ohio.

                                     In  Runoff
      Constituent                Range        Average     In  Rainfall
 Suspended solid;	     5-2074         313.0         11.7
 Total N 	   2.2-12.7           9.0          1.17
 Inorganic N  	   0.2-8.2            5.0          0.86
 Total hydrolyzable P	  0.08-1.07           0.6          0.03

 Source:  Weibel et al. (1966)  as reported by  Biggar and Corey (1969).

 investigations  have involved quantitative  evaluation  of the  factors
 controlling  the  amounts  of nitrogen  and  phosphorus  reaching
 streams and lakes from agricultural sources.   However, some  data
 are  available which  are useful in considering  the amounts of nu-
 trients  transported in this manner.
      Sylvester and Seabloom (1962) studied the  amounts of nutrients
 carried to surface  drains, subsurface drains,  and  shallow wells for
 irrigated  and nonirrigated situations in the  Yakima  Basin (Table
 23.1). Surface drains should reflect surface runoff, subsurface drains
 the  nutrients  leached  to  shallow depths, and  shallow  wells the
 nutrients carried in  percolating water to  the  groundwater.  As ex-
 pected, the amount of phosphorus carried to the  groundwater  was
 small, although amounts appearing in shallow  drains were compar-
 able to the amounts in subsurface drains. The quantities of nitrate ap-
 pearing in subsurface drains and  shallow wells reflected the mobility
 of nitrate in percolating waters.  Recycling  of the water through
 irrigation  tended  to  increase the   amounts  of  nutrients  in  the
     Weibel  et al.  (1966) measured  the nutrient  concentrations in
 runoff waters from a small wheat field  in Ohio (Table 23.2).   It is
 of interest to note the relatively high concentrations reported for
 runoff,  the range in amounts of suspended solids,  and the relatively
 large amounts of  materials in  the rainfall.
     Further indication of the importance of runoff is obtained from
 the results of Duley and Miller (1923) shown in Table 23.3.  Annual

 TABLE 23.3.   Annual nitrogen and phosphorus  content of runoff  and eroded
             material from a shelby loam soil of 3.6% slope.

                                           Pounds per Acre
Cropping System                    Total N       NOr-N      Total P
Not cultivated 	
Spaded 8" deep 	
Bluegrass sod 	
Rotation — corn,
wheat, clover 	
Corn annually 	



Source:  Duley and Miller (1923) as reported by Biggar and Corey (1969).


 TABLE 23.4.  Estimated amounts of nitrogen and phosphorus in  agricultural

                             Pounds per Acre of Land per Year
 Drainage Area         Inorganic N  Organic N  Inorganic P   Organic P
 Lake Monona  	     4.4         1.6        0.06
 Lake Waubesa  	     4.9         1.8        0.10        0.29
 Lake Kegonsa  	     6.4         1.8        0.10        0.31

 Source: Sawyer (1947).

 amounts transported ranged from 0.6 to  99 Ib/acre of total nitro-
 gen, 0.02  to 1.38 Ib/acre of NO3-N, and 0.1 to 48 Ib/acre of total
 phosphorus. Of particular interest is the marked effect of the crop-
 ping system and the low amount of  NO3-N transported relative
 to total nitrogen.
     From  another  point of view,  Sawyer (1947) estimated  the
 amounts of nitrogen and phosphorus carried from the  watershed
 to three Wisconsin lakes based  on analysis of one tributary to each
 lake (Table 23.4). The estimated pounds  of nutrients  lost per acre
 of land per year  were from 4.4  to  6.9 for inorganic nitrogen,  1.6 to
 1.8 for organic nitrogen,  0.06 to 0.10 for inorganic phosphorus, and
 0.29 to 0.31 for organic phosphorus.  Estimates of nutrient  losses
 for harvested areas of the United States (Table 23.5) reported  by
 Lipman and Conybeare (1936) were 4 to 6 times  greater for nitrogen
 and 4 to 40 times greater for phosphorus than the values estimated
 by Sawyer (Table 23.4).  Leaching estimates  were based on lysimeter
 and river  analysis,  while erosion  estimates  involved  amounts  of
 eroded material lost at various locations and the nutrient content of
 the  soil in the corresponding region.

     Although eutrophication of surface waters  through transport
of nutrients from surrounding lands is a natural process, primary
concern is focused on whether the activities  of man are increasing
the amounts of nutrients transported in agricultural drainage as well
as from  other sources  and whether  practices can be  implemented
which will lessen the nutrient influx and thereby preserve the quality
of our waters.
     The relative contribution of agriculture  to the nutrient budget
of a lake depends on  types  of activities occurring in  the  drainage
basin of the lake.  For example, lakes located in  rural areas may be
influenced primarily by agricultural drainage, while the  effect of
extensive  urban development in the drainage  basin may be to lower
the relative importance of agricultural drainage.  Furthermore, the
types of agricultural practices and  activities also influence the qual-
ity  of agicultural drainage,  and consequently the relative role of
agriculture in eutrophication. For example, animal feeding and soil

                       CHAPTER 23 / POLLUTION AND EUTROPHICATION / 321

 TABLE  23.5.  Loss of plant nutrients from harvested crop areas in the U.S.A.,

Intertilled crops
Annual crops not intertilled
Biennial and perennial crops

17 1
48 1
11 1
24 2

per Acre

21 0

4 9

10 6

per Year
39 1
280 7
37 6
65 0

141 1

Source:  Lipman and Conybeare (1936)  as reported by Biggar and Corey

management practices can have  a marked effect on  the quality of
agricultural drainage.
     An  indication of the effect  that  certain agricultural  activities
and management practices can have on the quality of agricultural
drainage is shown in  the  following  examples.
     Animal wastes are  one of the largest  sources  of agricultural
wastes (Loehr, 1969),  and concern has been focused  on the impact
of these  wastes on water quality, particularly the amount of nitrogen
transported in runoff  and percolate waters  from  animal feedlots
which represent a concentrated source of these wastes. The  survey
of well waters in Missouri  conducted by  Smith  (1964)  seemed to
show a relation between  animal population and  the  nitrate content
of the  groundwater.
     Stewart et al. (1967)  compared the nitrogen and phosphorus
contents of the surface groundwater beneath feedlots to that beneath
nearby irrigated fields (Table 23.6). Their results  showr that concen-

TABLE 23.6.  Concentrations of  constituents in  surface of  groundwater  be-
            neath four feedlots and  adjacent irrigated fields.

                                       MG/L in Water of
              Depth to Water                     Total
                   Table     NO,--N  NH4+-N  dissolved P  Organic C

Feedlot .
Feedlot .

field . .

field . .

field . .

field . .







Source:  Stewart et al. (1967).


 trations  of  nitrate, ammonium,  phosphorus,  and  organic  carbon
 were generally higher beneath the feedlots. However, because of the
 greater  area  occupied  by irrigated lands,  it was  suggested that for
 this  area irrigated  lands were contributing more  nitrate  to the
 groundwater than were the feedlots.
     Although attention is usually focused on runoff and percolates,
 the  recent results of  Hutchinson and Viets (1969)  indicate  that
 volatilization of ammonia from feedlots can  cause transport of sig-
 nificant quantities of nitrogen to nearby surface waters (Table 23.7).
 Depending on feedlot size and  distance from  the  feedlot, about  4  to
 35 kilograms of NH3-N/hectare  (one-half of  the values obtained
 by adsorption of ammonia in acid traps) were transported to nearby
 surface waters. These  amounts were much larger than the quantity
 of NH4-N contained in precipitation (the precipitation values shown
 in Table 23.7 are for a  3-month  period).
     The results  of  Weidner et al. (1969) recorded  in  Table  23.8
 indicate  the  effect that  soil management and  crop  rotations  can
 have on the amounts of nutrients carried in runoff waters. Improved
 management reduced nitrogen in runoff by  about 63%  and phos-
 phorus by  70%.  These  values were  estimated from correlations
 between  the  quality parameters   and  total  solids  in the  runoff,
 and it is  seen that the  main effect  of improved management was  to
 reduce the total solids transported in the runoff.  Improved manage-
 ment  primarily  involved contour  tillage, liming of  the soil,  and
 increased fertilization.
     The data obtained  by Johnson  et al. (1965) suggest the im-
 portance of fertilizer and cropping practices  on the  quality of agri-

 TABLE  23.7.   Absorption of ammonia volatilized from cattle feedlots.

                                 Ammonia-N  (kilograms per hectare)
     Site Description              Weekly       Annual    Precipitation
Control — no feedlots or
irrigated fields nearby ....
Small feedlots within 0.8 to
4 km 	
0.2 km east of 800-unit feedlot
and 0.6 km west of about
800-unit feedlot 	
0.5 km southwest of 9,000-unit
feedlot (shore of Clark Lake)
2 km northwest of 90,000-unit
feedlot (shore of Seeley
2 km east of 90,000-unit
0.4 km west of 90,000-unit




Source:  Hutchinson and Viets (1969).
* Absorption in 0.01 N FLSCX; absorption by lake water estimated to be one-
half of these.

                        CHAPTER 23  / POLLUTION AND EUTROPHICATION  / 323

TABLE 23.8.   Estimated annual amounts of constituents in runoff from  rural
             land  as  affected by management practice  (prevailing or im-
             proved) and cover crop.

                                     Pounds per Acre
                          of Constituent in Runoff for Cover Crop of
Total solids 	
Hydrolyzable P . . .
Total N 	

33 200 0
120 0
1 300 0
237 0

3 660 0
28 0
480 0
88 0

1 730 0
170 0
31 0

480 0
4 0
64 0
11 0



Source:  Weidner et al. (1969).

cultural  drainage  (Table  23.9).  The experiments were  performed
on  soils  described as deep, permeable,  silty  clays, with  tile  drains
located  at depths  of 5.5 to  7 feet.  More nitrogen was contained in
both  tile  drain effluents  and  surface  runoff from fertilized than
from  nonfertilized systems.  However,  phosphorus losses were  low
compared to the phosphorus content of the irrigation water, suggest-
ing a net removal of phosphorus from  the  irrigation  water by  the
soil.  Similarly, in the nonfertilized  system,  less  nitrogen was  lost
than  applied in the irrigation water.
    To evaluate  the importance to eutrophication of agricultural
drainage  relative  to  other nutrient sources,  all nutrient  sources for
the particular water  must be considered.  Estimates have  been made
of the nutrient sources for Lake Mendota,  Wisconsin, the  surface
waters of Wisconsin, and  the  water supplies of the  United  States.
Review  of these estimates is  useful  in  evaluating the contribution
of agricultural  drainage  to the nitrogen and phosphorus status of
natural  waters.
TABLE 23.9.   Nitrogen and phosphorus balance for tile-drained soils under
             different cropping and fertilizer treatments.
                      Pounds Applied
    Pounds Lost









Source:  Johnson et al. (19S5).


      Lake Mendota, Wisconsin

      Lake Mendota, Wisconsin, provides  an example of a lake in-
 fluenced to  a major degree by rural and urban areas.  Because of
 the importance of the lake to the region and concern over its eutroph-
 ic nature, an attempt was made to estimate the amounts of nitrogen
 and phosphorus entering the lake from various sources (Lee et al.,
 1966; Schraufnagel et  al.,  1967).
     Lake Mendota is  approximately 9,730  acres in  surface  area,
 with a maximum depth of  24  meters.  Madison, with a population
 of about 175,000, is the largest  city in the watershed.
     The Lake Mendota watershed  covers about  142,000 acres and
 is described by Schraufnagel et al. (1967) as an area occupied by
 permeable, calcareous,  loamy  glacial deposits,  with  a significant
 covering of loess. Most soil development is in the loess cover, with
 some  development  occurring in the  glacial  till immediately below
 the loess.  Many of the soils were developed under prairie vegetation
 and are  characterized  by an A horizon 8 to 16 inches thick and
 relatively high in organic matter.  Slopes  in  most of the watershed
 are gentle.  Numerous  small, undrained  depressions  occur in the
 uplands, and several large, wet lands containing organic  soils are
 located in the watershed.  Numerous dairy farms occupy  the  area;
 the estimated dairy cow  population is 100 cows per square mile.

     The contribution of rural runoff to the nitrogen and phosphorus
budgets  of Lake Mendota  was estimated  by considering  the  land
use in the watershed and the  amounts  of  nutrients lost from  each
type of land (Lee et  al., 1966; Schraufnagel  et al., 1967).  The dis-
tribution  of land in the watershed  according to  use is  shown in
Table 23.10.  A large portion (102,500 acres  or 73%) of the water-
shed is devoted to  cropland, with smaller areas in woodland (7% ),
pasture (8%), wetland (5%), and urban centers (7%).
     Estimates of  the  amounts  of nitrogen   and  phosphorus  con-
tributed to Lake Mendota from the various types of rural lands are
shown in Table 23.11.  The largest contribution was estimated to

TABLE 23.10.   Estimated land  use in the Lake Mendota watershed.

      Land Use                     Acres      Percent of Watershed
Corn and row crops . . . .
Hay and pasture 	
Pasture and other . . .
Major wetland 	


51 000
11 400

142 000



Source:  Water Subcommittee (1967).

                      CHAPTER 23 / POLLUTION AND EUTROPHICATION / 325

TABLE 23.11.  Estimates of the annual amounts of nitrogen and phosphorus
             contained in runoff waters in the Lake Mendota watershed.

                           Pounds per Acre     Pounds per Watershed

    Land Use            Nitrogen  Phosphorus   Nitrogen  Phosphorus
Cropland and pasture . . .
Manured land 	







45 000

52 200


15 000

20 430

Source:  Water Subcommittee (1967).
be from manured land, accounting for about 87%  of the nitrogen
and  73%  of the phosphorus. Cropland contributed about 27%  of
the phosphorus and  13%  of the  nitrogen.  Although  insufficient
data were available  to estimate the  contribution  of wetlands,  it
was believed that the  amounts of nitrogen and phosphorus received,
particularly from  drained marshes,  would be significant.
     The  contributions from manured  lands were calculated  by
assuming that one-half of  the manure from dairy cattle was ap-
plied to frozen soil and that 3 pounds  of nitrogen and  1 pound of
phosphorus were  lost for  each  10 tons  per acre application  of
manure.  These estimates were based on observation of Midgley and
Dunklee (1945) for a  frozen soil of 8%  slope.  Amounts from  crop-
land and  pasture  were  estimated  from concentrations in runoff
from.a Miami  silt-loam soil with 10%   slope (Eck et al.,  1957) and
assuming 2 inches of runoff per year.   Only water-soluble forms of
nitrogen and phosphorus were considered.  Values for wooded  areas
were obtained from the nitrogen  and phosphorus contents of streams
flowing through these areas (Sylvester,  1960) and were  considered
very rough estimates  as they did not distinguish between amounts
contributed by surface runoff and  base flow.

     The relative importance of rural runoff,  percolate waters, and
other sources is shown from  the  estimates  of the total  nutrient
budget of Lake Mendota recorded in Table 23.12 (Lee et al.,  1966).
Rural runoff was  the largest phosphorus contributor  (42%),  while
groundwater accounted for the major  portion  of  nitrogen  (52%).
However, the quantity of nitrogen contributed by rural  runoff (52,-
000 Ib/yr) was larger than the corresponding quantity of phosphorus
(20,000 Ib/yr).   For  nitrogen, precipitation on the lake surface was
the second largest  contributor (20% ), followed by rural runoff (11%)
and  municipal and industrial wastewaters (10%).  For phosphorus,
municipal  and  industrial  wastewaters  were  the  second  largest
source (36%), followed  by urban runoff (17%).
     The large  amount of soluble nitrogen contributed by ground-
water  shows the  importance  of nitrate transport from  soils  to


 groundwater by water  percolating through the soil.  Estimates for
 the contribution of groundwater included both  that  entering  the
 lake directly  (about 30  cfs) and that reaching the lake through con-
 tributing to the flow of surface tributaries (about 35  cfs).  Concen-
 trations of NO3-N  of  2.5  mg/1  for groundwater entering through
 surface tributaries and  1 mg/1 for direct-entering  groundwater were
 assumed.  A  lower NO3-N  concentration for groundwater  entering
 below the surface was  used because it was assumed that denitrifica-
 tion was  of  greater importance in these  waters than  in  surface
     It is  of interest to compare  the estimates in Table  23.12 with
 values  obtained by measuring flow and nutrient  concentrations  in
 the tributaries  entering Lake Mendota (Rohlich,  1963). Tributary
 measurements, which do not include groundwater entering the lake
 beneath the  surface, indicated that  259,700 pounds  of inorganic
 nitrogen and  343,400 pounds  of total nitrogen entered the lake dur-
 ing the year  October 1948 to  October  1949. This compares with a
 total estimated  nitrogen budget of 478,300 pounds per year in Table
 23.12.  The total phosphorus contribution from tributaries indicated
 from direct measurements was 53,389 pounds per year compared
 with an estimate of 47,000 pounds per year in Table 23.12.
     Briefly, estimates  of contributions from other sources shown
 in Table 23.12 were obtained as follows: Quantities of nitrogen and
 phosphorus in municipal and industrial wastewaters were estimated
 from the individual sources,  including municipal-treated domestic
 wastes  from small villages in the watershed, private domestic waste
 disposal systems, milk and  cheese processing and canning  compa-
 nies, and  a  car wash.  Treated domestic  wastes  from  the  city  of
 Madison are not discharged into Lake Mendota. Urban runoff values
 were estimated from data  obtained for  Cincinnati,  Ohio  (Weibel
 et al., 1964),  with allowances  made for the higher degree of indus-
 trialization of Cincinnati than of Madison.  For precipitation, a value
 of 10 pounds  of nitrogen per acre per  year was used (Shah, 1961);
 the value of 1,300 pounds of phosphorus per year for Lake Mendota

 TABLE 23.12.   Estimated sources  of nutrients for Lake Mendota, Wisconsin.
                                                 Percent of Total
   Nutrient Source
   Pounds per Year
Nitrogen   Phosphorus  Nitrogen  Phosphorus
Municipal and industrial
waste water ....
Urban runoff 	
Rural runoff 	
Precipitation on lake
Nitrogen fixation ....
Marsh drainage




17,000* 10
8,100f 6
20,000f 11
1.300 20
600 52
< 1
. Not estimated



Source:  Nutrient Sources Subcommittee (1966).
* Total of nutrient forms.
f Soluble nutrient forms.

                       CHAPTER 23 / POLLUTION AND EUTROPHICATION / 327

 is an average value derived from  several sources.  The  quantity  of
 nitrogen-fixation was based on a rate of 0.02%  of nitrogen fixed per
 day as reported by Goering (1963) and  the assumption that nitrogen-
 fixation occurs 3 months per year and in the top  3 meters of the
 lake.  Although marsh drainage was not estimated, its contribution
 may be  significant.
     Nutrient Sources for  Waters in Wisconsin

     Using an approach similar to that described for the Lake Men-
 dota watershed,  the  amounts of nitrogen and phosphorus reaching
 surface waters of the state of Wisconsin from various sources were
 estimated  by Schraufnagel et al. (1967).
     The importance of rural sources relative to other sources dif-
 fered somewhat  from the values for Lake Mendota (Table  23.13).
 Rural  sources were  estimated to contribute  54%  of  the  nitrogen.
 Of this, the  largest portion  (42%) came from the  groundwater.
 Rural sources accounted for 30% of the  phosphorus, 21.5%  arising
 from manured land runoff. However, municipal treatment facilities
 were the largest phosphorus contributor (55.7%), while groundwater
 contributed the largest portion of nitrogen, as was the  case for Lake
     Nutrient Sources  for Water Supplies of the United States

     In 1967 a Task Group of the American Water Works Association
prepared a report  on the sources of nitrogen  and phosphorus in

TABLE 23.13.   Estimated amounts of nitrogen and phosphorus reaching Wis-
              consin surface waters.

                             Thousands of
                            Pounds per Year        Percent of Total
       Source             Nitrogen  Phosphorus  Nitrogen  Phosphorus
Municipal treatment
Private sewage
Industrial wastes 	
Rural sources
Manured lands . . .
Other cropland . . .
Forest land 	
Pasture, woodlot,
and other
Urban runoff 	
Precipitation on
water areas 	

Total .




12 558


100 0

2 9

100 0

Source:  Water Subcommittee  (1967).


 TABLE 23.14.   Estimated  amounts  of nutrients  contributed from  various
               sources for water supplies of the U.S.
    Nutrient  Source
     Millions of
   Pounds per Year

Nitrogen   Phosphorus  Nitrogen  Phosphorus
                                                 Percent of Total'
Domestic waste ....
Industrial waste . . .
Rural runoff
Agricultural land
Farm animal waste •
Urban runoff 	

1,100-1,600 200-500
1,500-15,000 120-1,200
400-1 900 150-750
110-1 100 11-170
30-590 3-9


29 0

 Source:  AWWA Task Group 2610-P (1967).
 * Percentages are based on mean value of ranges given.

 water supplies  in  the  United States (McCarty et al., 1967).  Their
 estimates  are  shown  in Table 23.14.  These estimates are for all
 water supplies,  including groundwater.  Thus the  contribution by
 rural runoff includes drainage to the groundwater as well as surface
 runoff. It should be noted that the percentages shown in Table 23.14
 were calculated from the means  of the ranges  of  values  reported
 by the Task Group and they may differ appreciably from the actual
 average  contribution for each source.  Consequently  the percentages
 are useful only for very rough  approximations.  Futhermore, in the
 manner  calculated, the percentages total  100, even though farm ani-
 mal waste and industrial waste contributions of phosphorus were not
     The values in this table suggest  that  agricultural land  is  an
 important contributor of nitrogen  and phosphorus to water. About
 60%  of  the nitrogen and 42% of the phosphorus were  estimated to
 come from agricultural land.  To  arrive  at these  figures  it was as-
 sumed that the 308 million acres of  cultivated land in the United
 States contributed 5 to 50 pounds of nitrogen per acre per year or a
 total  of  1,500  to 15,000 million  pounds of nitrogen per year. As
 estimated, phosphorus contribution of 0.4 to 4 pounds  per acre per
 year gave the total estimated amount of 120 to 1,200 million pounds
 per year.
     It should be emphasized that the nutrient budget estimations
 that have  been  discussed were based on data  obtained on a  small
 scale in most cases, and extrapolation of these localized evaluations
 to an entire watershed or larger area gives estimations of a rather low
 reliability.  More precise estimations based on more extensive evalua-
 tion of representative watersheds would certainly be useful  in plan-
 ning  management  programs  to control the influx of nitrogen and
 phosphorus into water supplies. Nutrient sources are numerous and
 generalizations as to which source is the most important  cannot be
made. However, these estimations indicate that the  contribution of
 agriculture is significant.  The  challenge is  that this  contribution
 should be reduced by improved and more efficient agricultural man-
agement  practices.

                      CHAPTER 23 / POLLUTION AND EUTROPHICATION / 329


Armstrong, D. E., Spyridakis,  D. E., and Lee, G. F.  1969.  Cycling
     of nitrogen and phosphorus in natural waters with  particular
     reference to  the  Great  Lakes.  Presented  at the  ACS Symp.
     on the Chemistry of the Great Lakes, Minneapolis, Minn.
Biggar, J. W., and Corey, R.  B.  1969.  Agricultural drainage and
     eutrophication.  In Eutrophication:  causes, consequences, cor-
     rectives.  Proc.  Intern. Eutrophication Symp.,  Madison,  Wis.
     Wash., B.C.:  Natl. Acad. Sci.
Duley, F. L., and Miller,  M.  F.  1923.  Erosion  and surface runoff
     under different soil conditions.  Mo. Agr. Exp. Sta. Res. Bull. 63.
Eck, P.,  Jackson, M. L.,  and Bay, C. E.  1957.  Annual report AES
     Project 791 (Phase 5).
Fruh, E. C.,  Stewart, K. M., Lee, G. F., and  Rohlich, G.  A.  1966.
     Measurements of  eutrophication and trends.  J. Water  Pollu-
     tion Control Federation 38:1237-58.
Goering,  J. J.  1963.   Studies of nitrogen-fixation in natural fresh
     waters.  Ph.D. thesis, Zoology Dept., Univ. of Wis.
Hutchinson, G. L., and Viets, F.  G., Jr.  1969.  Nitrogen enrichment
     of surface water by adsorption of ammonia volatilized from cat-
     tle feedlots.  Science 166:514-15.
Johnson, W.  R.,  Illihadich,  F., Daum, R.  M., and Pillsbury, A.  F.
     1965.  Nitrogen and phosphorus in tile drain effluent.  Soil Sci.
     Soc. Am. Proc. 29:287-89.
Lee, G. F., chairman, Nutrient Sources Subcommittee.  1967.  Report
     on the nutrient sources  of  Lake Mendota. Water Chemistry
     Program, Univ. of  Wis., Madison. (Mimeo.)
Lipman,  J. G., and Conybeare, A.  B.  1936. Preliminary note on the
     inventory and balance  sheet of plant nutrients in the  United
     States.  N. J. Agr.  Exp. Sta. Bull. 607.
Loehr, R. C.  1969. Animal wastes—a  national problem.  /.  Sanit.
     Eng. Div. Am. Soc. Civil Engrs. 95:189-221.
McCarty, P. L., chairman Task Group 2610-P.  1967.  Sources of ni-
     trogen and phosphorus in water supplies. /. Am. Water Works
     Assoc. 59:344-66.
Midgley,  A. R., and Dunklee, D. E.  1945. Fertility runoff losses from
     manure  spread during the ivinter.  Univ.  of Vt. and State Agr.
     College Agr. Exp.  Sta. Bull. 523.
Rawson,  D. C.  1939.  Some physical and chemical factors in the
     metabolism of lakes. AAAS Bull. 10:9-26.
Rohlich,  G.  A.   1963.  Origin and quantities  of plant nutrients  in
     Lake Mendota. In Limnology in North America, ed. D. C. Frey.
     Madison: Univ. of Wis. Press.
Sawyer, C. N. 1947.  Fertilization of lakes by agricultural and urban
     drainage. /. Neiv  Engl. Water Works Assoc.  61:109-27.
Schraufnagel, F. H., chairman. Working  Group on  Control Tech-
     niques and  Research on Water Fertilization.  1967.  Excessive
     water  fertilization.  Report  to  Water Subcommittee,  Nat. Re-
     sources Committee of State Agencies, Wis. (Mimeo.)
Shah, K.  S.  1961. Sulphus and nitrogen brought down  in  precipita-
     tion in Wisconsin.  Master's thesis, Soils Dept. Univ. of Wis.,
Smith, G. E.  1964. Nitrate problems in plants and water supplies in
     Missouri. 92nd Ann. Meeting,  Am. Public Health Assoc., New
     York City.
Stewart,  B. A., Viets,  F.  G., Jr., Hutchinson,  G.  L., and Kemper,


     W.  D.  1967.  Nitrate  and other water pollutants under fields
     and feedlots. Environ. Sci.  Technol.  1:736-39.
Stewart, K. M., and Rohlich, G. A.  1967. Eutrophication—a review.
     Publ. 34, State Water Quality Control Bd., Calif.
Sylvester, R. O.  1960.  Limnological aspects of recreational  lakes.
     Public Health Serv. Publ. 1167.
Sylvester, R. O., and  Seabloom, R.  W.  1962. A study on the char-
     acter and significance of irrigation  return flows in the Yakima
     River Basin. A report from  the Univ. of Wash.
Weibel, S. R., Anderson, R. J., and Woodward, R.  L.  1964.  Urban
     land runoff as a factor in stream pollution. /. Water Pollution
     Control Federation 36:914-24.
Weibel, S. R., Weidner, R. B., Cohen, J.  M., and Christiansen, A.  G.
     1966. Pesticides and other contaminants in rainfall and runoff.
    J. Am. Water Works Assoc. 58:1075-84.
Weidner, R. B., Christiansen, A. G., Weibel, S. R., and Robeck, G.  G.
     1969.  Rural runoff as a factor  in  stream pollution.  J.  Water
    Pollution Control Federation 41:377-84.


       ^ECREATIONAL use of surface waters involves the employment
of leisure time for enjoyment of fishing, boating, swimming, and the
esthetic values of water.  Pollution is  the addition of  material to
water which produces results undesirable to man, including death of
organisms, impairment of metabolic life processes,  or the production
of nuisance odors and algal scums.
    Man's full recreational enjoyment of water demands the presence
and diversity of animals and  plants.  The ecology of these living
organisms, and the impact of agricultural pollutants on them, is best
understood with reference to the aquatic community.

    The aquatic community is the interdependent group of plants
and animals living in a lake, pond, or stream.  Interdependence is
most easily seen in food-procuring activities, where each organism
functions as a food producer or as a consumer.  This complex com-
munity is dependent on photosynthesis in the same way that all ag-
ricultural production is ultimately dependent on food synthesis by
green plants.  The process of photosynthesis converts sunlight energy
to chemical energy which is incorporated into carbohydrates,  fats,
and proteins.  Thus, algae and rooted green plants are the producers
in the aquatic community and comprise the principal source of  food
for the dependent group of consumer animals.  In streams and rivers
rooted plants  and algae are less abundant than in lakes. Plant ma-
terials produced on the land and washed into streams and lakes are
an additional food source for consumers.
    Figure 24.1 shows that ingestion  of green  plants by  animals
initiates the transfer of sunlight energy  to one  or more levels  of
     ROBERT S. CAMPBELL is Professor of Zoology, University of Missouri.
     JAMES R. WHITLEY is Supervisor, Water Quality Investigations, Mis-
     souri Department of Conservation.


     FIG. 24.1.  A simplified food chain involving one producer link and
     four consumer links.  Arrows indicate direction of flow of energy.

consumers, and demonstrates the dependence of each consumer level
on lower levels of consumers  and ultimately on producers. At death
organisms are mineralized by decomposer bacteria and nutrients are
released to be incorporated by producers. Any factor which adversely
affects the environment of  this complex community may  affect
directly all levels (producer,  consumer,  decomposer) in the  com-
munity. If only one level is directly affected, all other levels will be
affected indirectly because of their interdependence.  Thus, environ-
mental pollution, however slight,  may have far-reaching effects on
the  entire  aquatic community.  For example, any  agricultural prac-
tice which increases soil  erosion and turbidity of water will interfere
directly with  the photosynthetic  process  and indirectly with  the
poundage of fish produced in  that body of water.
    We are concerned in this chapter with the aquatic communities
of streams and lakes.  Agricultural pollutants  that have a profound
impact on the aquatic community  include (1) pesticides, (2) irriga-
tion return water, (3) eroded soil, and (4) agricultural fertilizers  and
animal wastes.

    Trace levels of pesticides in water may be concentrated in the
tissues of aquatic organisms. If these organisms are in turn eaten,

                     CHAPTER 24 / POLLUTANTS AND RECREATIONAL USES / 333

 pesticides are further concentrated in the consuming animals.  Thus,
 in one food chain, pesticides may become progressively concentrated
 in animal tissues at successive  levels,  so  in the third  and  fourth
 consumer levels the concentration may  exceed the  concentration in
 the  water by several thousandfold.  This phenomenon of biological
 magnification  occurs with the chlorinated  hydrocarbon insecticides
 because they are selectively absorbed into the oils, fats, and waxes in
 the living organisms in the aquatic environment. Most surface waters
 in the United States now contain DDT  and its related  compounds
 (American Chemical Society,  1969).  An example of this type of
 pesticide magnification is the death of fish-eating birds resulting from
 the use of DDD to control gnats in Clear Lake, California (Hunt and
 Bischoff, I960; Rudd, 1964).
        DDD was applied in 1949, 1954, and 1957, with near-complete
 control of the gnat.  Prior to 1949, more than 1,000 pairs of western
 grebes nested at the lake but apparently did not breed subsequent to
 treatment. Grebes did continue to visit the lake annually. There was
 a die-off in 1954,  1955, and 1957, attributed to high levels of DDD in
 the tissues.  Inspection of the aquatic  food  chain showed that DDD
 levels in  tissues were progressively  greater  at successive consumer
 levels (Table 24.1).
     A prrblem closely related  to biological  magnification of chlori-
 nated hydrocarbons is the development of resistance to pesticides by
 organisms. The development of resistance  is a well-known  obstacle
 in the control  of insect pests with insecticides.  Vinson et al. (1963)
 reported resistance in fish to chlorinated hydrocarbons.  The  ability
 of nontarget  organisms  to become  resistant would seem  to be
 beneficial.  However, resistance  can be  distastrous  to fish popula-
TABLE 24.1.   Biological magnification, Clear  Lake aquatic food chain.  Con-
             centrations are  maximal.  Values for  vertebrates are for vis-
             ceral fat.
Food Chain
of DDD
of DDD
in Excess of
That in Water

fPredaceous birds

 Carnivorous fish
   (largemouth bass)

   small fish

 Animal plankton


80,000 x

85,000 x

   500 x
Producer level

Plankton algae



265 x

Source:  Modified from Rudd (1964).


 tions.  Ferguson  (1967) reports that some fish  populations  from
 heavily treated areas can tolerate up to 1,500 times the dose of some
 insecticides  that  is lethal to nonresident fishes.   Ferguson states,
 "Resistant fishes are able to  tolerate massive body burdens of  these
 compounds in their tissues, and these residues constitute the source
 of concern regarding  the ecological significance of resistance,"  He
 concludes, "Our findings  indicate that  although  selection  of  a re
 sistant fishery may permit exposed populations to survive, it may
 ultimately produce a  biological  product dangerous to consumers of
 all sorts, including man himself,"
     A critical review  of the literature on the effects of pesticides on
 fishes (Johnson, 1968) emphasizes the following points:  (1) Spraying
 of streams has, in some instances, destroyed most of the aquatic in-
 sects.  (2) Pesticides do alter the composition of aquatic communities,
 This can involve reduction in game fish, elimination of predators with
 subsequent increase in prey species, arid reduction in members of the
 zooplankton  and  bottom-dwelling invertebrates. (3) The degree of
 toxicity may be dependent on the position of the plant or animal in
 the food chain—the fourth level carnivore may be affected more than
 the first-level consumer, due to biological concentration of pesticides.
 (4) Most studies with  fish have  concerned acute toxicity where the
 effect  is measurable by  death.  When fish survived, the implication
 was that the  pesticide was not toxic at the level tested; the possibility
 of damage through long-term exposure at sublethal concentrations
 was not answered.  (5) Acute  toxicity is mainly injurious to the  nerv-
 ous  system of fish.  (6) There are reports of damage in fish to the
 liver, gonads, blood, gills, and interference with normal physiological
     A typical example of fish loss from  exposure to DDT, anplied in
 concentrations of 0.5  to  1 pound per acre,  is cited from Cone and
 Springer (1958):  "Large  numbers  of  dead  trout, whitefish,  and
 suckers, including many young-of-the-year, were noted three months
 after the spraying along a 100-mile stretch of the river.  Great reduc-
 tions in numbers of aquatic invertebrates again took ulace.  This loss
 of food appears to have been  the chief cause of the fish die-off."
     Burdick et al. (1964)  compared fry  survival from esgs gathered
 from 12 lakes receiving varying amounts of DDT from the watershed.
 The authors  concluded  that  fry  mortality was induced when con-
 centrations of DDT in the egg exceeded approximately 3 prm. During
 the st'Tdy period when the watershed of Lake George, New York, was
 treated with  amounts of DDT ranging from 0.30 to 0.57 pounds per
 acre, concentration of  the  pesticide ranged from 4  to 15 ppm in egg
 tissue and 112 to 515 ppm in egg oil.
     In recent  months coho  salmon from Lake Michigan were re-
moved  from  the market by the Food and Drnp- Administration be-
cause  they contained  excessive  residues of DDT.  Some countries
and  the states of Arizona  and Michigan have banned  the use of
DDT, and others are considering similar action.
     The major concern  of conservationists regarding  chlorinated
hydrocarbon  pesticides is  the problem  of persistence.  The severe


 ecological effects  resulting from pesticide levels which  are  hardly
 measurable in the aquatic environment suggest that there is no safe
 level of application of these persistent chemicals which is consistent
 with economic agricultural use. These long-lasting compounds spread
 worldwide throughout the environment and concentrate in dangerous
 amounts  through the food chain. There are alternatives to the use of
 persistent pesticides in agriculture—namely, organic phosphates and
 carbamates,  and more intensive employment of biological  controls.

     Irrigation was the greatest  single use of water in the United
 States in 1960, accounting for 135 out of a total use of 322 billion
 gallons per day (U.S.  Bureau of the Census, 1962).  Water quality
 changes resulting from irrigation include temperature increase and
 total salinity  increase.  The  more serious effects from temperature
 elevation are reduction in the dissolved oxygen supply and increased
 toxicity of polluting substances.  An effect on fish of salinity increase,
 caused by irrigation return water, was described for  the San Joaquin
 River, California, by Radtke and Turner (1967).  Concentrations  of
 dissolved substances  in  excess  of 350  ppm  blocked the  upstream
 spawning  migration of striped bass.  However, at lesser concentra-
 tions of 100 to 350 ppm, upstream migration occurred as indicated by
 a 4- to 12-fold increase in gill-net catch.

     Photosynthesis in water is restricted to that upper euphotic zone
which receives 1% or more of incident sunlight.  For example,  the
euphotic zone in Lake Erie, 1939-40,  varied in thickness  from 32
feet when turbidity was 5 ppm to 3 feet when turbidity was  115 ppm
(Chandler, 1942).  This reduction of light penetration by suspended
soil  reduces total photosynthesis  and hence  total production  within
the aquatic community.  Chandler and Weeks  (1945) proposed that
increased turbidity in 1942 in Lake Erie appeared to have resulted in
a 19% reduction of spring phytoplankton from the  1941 level.  But-
ler (1964) wrote that primary productivity in  central Oklahoma farm
ponds varied inversely with  turbidity.   Summer photosynthetic rate
in one clear pond was three times that  of a turbid pond.
     The effects of turbidity on fish and other consumer organisms
are varied, but the overall result is one of reduction in total produc-
tion.  According to Trautman (1957), "Studies made since 1925 have
proved that since then, if not before, soil suspended in water  has been
the universal pollutant in Ohio, and the one which has mast drastical-
ly affected the fish fauna.  Clayey soils, suspended in water, prohibited
the proper penetration of light, thereby preventing development of
the aquatic vegetation, of the food of fishes, of  fish  egers and of fry."
     These views are supported by Cordone and Kelley (1961):  "There


 is  abundant evidence that sediment is detrimental to aquatic life in
 salmon and trout streams. The adult fishes themselves can apparent-
 ly  stand normal high concentrations without harm, but deposition of
 sediments on the bottom will reduce the survival of eggs and alevins,
 reduce aquatic insect fauna, and destroy needed shelter.  There can
 scarcely be any doubt that prolonged turbidity of any great degree is
 also harmful."
     Whether  the physical  contact of  suspended  solids  is directly
 detrimental to adult  fishes is not resolved.  Laboratory studies on 16
 species  of  freshwater fishes  suggested that  "the  direct effect of
 montmorillonite clay turbidity is not a lethal condition in the life of
 juvenile to  adult fishes at turbidities found in nature" (Wallen, 1951).
 Wallen reported that most individuals survived for a week or longer
 exposures to 100,000 ppm suspended clay, a value at least  lOx great-
 er  than expected turbidities  in natural waters.  On the  other hand,
 Herbert and Merkens (1961) concluded that continual  abrasion by
 suspended solids in concentrations of 90 to 810 ppm in experimental
 tanks may have induced gill thickening which  was observed in some
 trout but not in others.  They suggest that such gill alteration  may
 make fish more susceptible to  other stresses in  the environment  and
 thus reduce survival chances.
     Effects of turbidity  on bass and sunfish, measured over 2 grow-
 ing seasons in  12 ponds in Illinois where turbidity was approximately
 25  ppm in  the clearer ponds and in excess of  100  ppm in the most
 turbid,  are  described by  Buck  (1956):  (1) The average  total weight
 of fish in the clear ponds was 5.5 times greater than in muddy ponds
 at the end of the second  growing season; (2) growth rate in length of
 first-year bass  was three times greater in  the  clear ponds than in
 muddy  ponds;  (3) the weight  increase in bass at the end of the
 second  growing  season was  5.5 times  greater in  the clear ponds;
 (4) bass reproduction was suppressed in the more turbid ponds.  He
 found  similar results in studies on 14 hatchery ponds  and 2  large
 reservoirs.  Swingle (1949) also reported the failure of largemouth
 bass to  spawn  in  ponds receiving a  large inflow of highly turbid
     Aneler success for  most  game species is  improved  in clearer
 water.  "The clear reservoir  attracted  more anglers,  yielded greater
 returns  per unit of fishing effort, as well as more desirable srjecies,
 and was immeasurably more  appealing in the aesthetic sense" (Buck,
 1956).  Catch  success for game fish  is  directly related to water
 clarity.   In Little Dixie Lake, Missouri, in  1969, extended summer
 rains restricted water visibility to  a depth of 8 to 16 inches, June
 through  July.  In August, water visibility increased to a  depth of 37
inches.  The catch of bass immediately increased 5-fold (J. L. Choate.
personal communication).  Similarly,  more bass and bluegill  were
 taken by anglers in Fork Lake, Illinois, during  periods of increased
water transparency (Bennett et  al., 1940). Angler use of the Meramec
River watershed  (Missouri) dropped one-third when  the water  flow
was above normal and muddy, resulting in an estimated annual eco-
nomic loss of $60,775 to  the residents (Brown, 1945).



     The slow process of aging of lakes and ponds is accompanied by
 a gradual increase in nutrients  with concomitant increased produc-
 tion of animal and plant life. Associated with this is a reduction in
 dissolved oxygen in deeper  water  because of accumulated organic
 matter, and a loss in water transparency due to blooms of algae and
 animal plankton.  The term applied to aging is eutrophication  and is
 defined  as  the  intentional  or  unintentional  enrichment  of  water
 (Hasler, 1947).  Serious aspects of eutrophication include impairment
 of esthetic qualities by unsightly nuisance algae and dense growth of
 rooted plants and a hastening of lake extinction, since the accumu-
 lated organic matter and eroded soils ultimately fill the lake  basin.
 Agricultural  pollutants which hasten eutrophication  include inor-
 ganic fertilizers  and animal wastes.
     It is currently thought  that nitrogen and phosphorus are the
 elements most  responsible  for lake  eutrophication  (Mackenthun,
 1968).  Large quantities of nitrogen and phosphorus are contributed
 to  surface  waters  by  agricultural  drainage  (Sawyer, 1947; Task
 Group Report,  1967;'Mackenthun,  1968).  It  is shown (Table 24.2)
 that concentrations  in agricultural drainage  and  irrigation  return
 water  are several times greater than  in uncontaminated lakes and
 streams, and as great or greater than in eutrophic lakes.
     Concentrations of  total  phosphorus less than 0.01 ppm usually
 limit biological  activity, whereas nuisance algal blooms may  be ex-
 pected when total phosphorus exceeds 0.05 to 0.1 ppm.  The relation-
 ship of nitrogen and phosphorus enrichment to eutrophication is dis-
 cussed by Armstrong and Rohlich (see Chapter 23).
     It has been suggested that there is a relationship between weight
 of the fish population and water fertility (Moyle, 1956; Table 24.2).
 The relationship of the standing  crop of fish in pounds per acre to to-
 tal phosphorus in ppm was 40 toO.02, 90 to 0.034, 150 to 0.058. and
 370 to 0.126.  However, increase in standing crop is accompanied by
 a change in  species composition.  For example, in Minnesota lakes
 (Moyle, 1956), as the standing crop of fish increased from 40  to 370
 pounds per  acre, the  structure of the fish  population changed from
 one involving lake trout in the 40 pounds per acre lakes to one in-
 cluding yellow  perch, walleye, northern  pike,  bass, and bluegill in
 lakes cf intermediate poundage; and finally, in the 370 pound per
 acre lakes, to one wrhere two-thirds of the standing crop  were un-
 desirable fish such as carp.
     Marked  biological  changes associated with eutrophication  in-
 clude loss cf esthetic values associated with loss of  water clarity and
 development of  algal  blooms, and maior changes in fish fauna and
 fish food  organisms.  Such changes have  been described  for Lake
 Erie (Beetcn, 1965) and for lakes Zurichsee, Switzerland, and Men-
 dota, Wisconsin (Hasler,  1947).  More general aspects of biological
problems  in recreational lakes are  described by Mackenthun et al.
(1964). An  annotated bibliography on nitrogen and phosphorus in
water was compiled by Mackenthun (1965).

TABLE  24.2.  Concentrations  of  nitrogen and  phosphorus  in milligrams per liter  (ppm) in different aquatic  communities, and
             the relationship of nitrogen  and phosphorus levels to biological activity.
   (as P)
  (as P)
  (as N)
 (as N)
Uncontaminated surface water ...      ...       0.01-0.03         ...          ...
Streams, forested area, little
     habitation  or  land use  	    0.007         0.069         0.130       0.204
Eutrophic Green Lake, Seattle  . . .    0.016         0.076         0.084       0.340
Sewage  	     1-13        3.5-9.0         7-40       18-50

Seepage water from agricultural
     soils, Illinois  	   0.2-0.7
Surface irrigation return flow ....    0.162         0.251         1.250       1.455
Cattle  feedlot wastes  	   16.3              ...        0.1-11
Limiting factor to biological
     activity  	      ...        <0.01
Nuisance algal  blooms expected
     when values exceed	    0.01        0.05-0.10         0.3

Fish production: 40 Ib/a	      ...         0.020
                370 Ib/a	      ...         0.126
                                                      Mackenthun (1968)

                                                      Sylvester (1961)
                                                      Sylvester (1961)
                                                      Bartsch  (1961)
                                                      Task Group Report (1967)

                                                      Engelbrecht and Morgan (1961)
                                                      Sylvester (1981)
                                                      Miner et al. (1966)

                                                      Sawyer et al. (1945)

                                                      Sawyer  (1947)
                                                      Mackenthun (1968)
                                                      Moyle (1956)
                                                      Moyle (1956)
 Note: The values of N and P in relationship to animal and plant  production  should be considered only  as  indicative of
 general relations since there is much variation among lakes and many factors affect production.


 TABLE 24.3.   Wastes of hogs and cattle in  Missouri expressed  as  human
             population equivalents.
Hogs and pigs 	
Cattle and calves . . . .

Number in
. . . 4,320 000
. . . 4,257 000
. . . 4,748,000

Pounds BOD
per Day
0 17

4 320 000

Source:  Modified from Ray (1965).
* USDA Statistical Reporting Service.

     Animal wastes which  enter surface waters  have such  a high
oxygen demand that they rapidly exhaust the dissolved oxygen. Ray
(1965)  expressed animal  waste  in terms  of "human  population
equivalents" by dividing the oxygen requirements (pounds BOD per
day) of animal waste by 0.17, the value for human wastes.  The pop-
ulation equivalents  calculated for hogs and  cattle on farms in Mis-
souri,  1969, are shown in Table 24.3.  Clearly the organic load of
animal wastes represents a  potential oxygen  demand on  receiving
waters in excess of that imposed  by human waste.  The concentra-
tion of animals in feedlots with uncontrolled drainage  results in
exaggerated surface water  degradation  at  the locations  of those
drainages.  Dissolved oxygen concentration should be  above  5 ppm
for a diversified warm-water fauna (Federal Water Pollution Control
Administration, 1968).  Lower  concentrations  adversely  affect the
respiratory  rate and general metabolism; prolonged concentrations
as low as 2 ppm are often fatal to fish.
     Cross and Brasch (1969) described a change in land use pattern
in the Neosho River watershed, Kansas, from seasonal  grazing to
year-round maintenance of cattle, with  many concentrated in feed-
lots. Associated with this  change was a loss  of 5 species of fishes
and a decline in abundance of at least 20 species.  Numerous fish
kills were attributed to  pollution from cattle feedlots whose  wastes
drained into streams.
     Smith  and Miner (1964) considered animal feedlot  runoff  a
significant source of water pollution in Kansas. They described run-
off water quality as follows:  (1) very high organic content, (2) con-
centrations  of  ammonia frequently  in excess of 10  ppm, and (3)
heavy bacterial populations.  The existence  of pollution was usually
indicated by fish kills  which they attributed  to ammonia and lo\v
dissolved oxygen.
     With  the  recent  adoption of wrater  quality standards  by the
states, minimal limits for the  addition of agricultural pollutants to
state waters were set by state law and are backed by federal enforce-
     The Federal Water Pollution Control Act of 1948  (Public Law


 660) formed the basis for federal-state cooperation and for enforce-
 ment of federal regulations on interstate waters through the attorney
 general. With the adoption of the Federal Water Quality Act of 1965
 (Public Law 234)  Congress  authorized the  states and the  federal
 government to establish water quality standards for interstate waters.
 After holding public hearings the states adopted  standards and sub-
 mitted them for review by the secretary of the interior.  As  of May
 1969, there was whole or partial acceptance of water quality stand-
 ards  by all 50 states.
      The development of standards considers the  uses to be made of
 the water in question, the assignment of specific water quality criteria
 to protect the water use, and plans for implementation and enforce-
 ment.  Water quality  criteria differ from state to state  and for dif-
 ferent waters within a state.
      A standard reference for water quality criteria which will pro-
 tect recreational and other uses of surface water is the Report of the
 National Technical  Advisory Committee to the Secretary  of the In-
 terior  (Federal   Water Pollution  Control  Administration,   1968).
 Criteria adopted by the state of  Missouri  are cited  as examples of
 their application to agricultural pollutants:

         All tributary streams and  all municipal, industrial, agricultural,
    and mining effluents  shall not  create  conditions  in the stream which
    will adversely affect the present water uses or ths future  water uses
    as they become current.
         Pesticides. Substances toxic  to man,  fish, and  wildlife or detri-
    mental  to agricultural, mining, industrial, recreational,  navigational,
    or other legitimate uses shall be limited to  nontoxic or nondetrimental
    concentrations in  the streams.
         Irrigation Return Water.  Effluents  shall not elevate  or  depress
    the average cross-sectional temperature of  the stream more than 5°F.
    The stream temperature shall not exceed 90°F due to effluents.
         Eroded  Soil.  There shall be  no turbidity of other  than natural
    origin  that will  cause substantial visible  contrast with  the  natural
    appearance of the stream or with its legitimate uses.  There shall be
    no noticeable man-made deposits of solids  either  organic or inorganic
    in nature on the stream bed.
         Animal Wastes. Dissolved  oxygen in the  stream shall  not be less
    than 5 ppm at any time due to effluents or surface runoff.
         The intent of enforcement of water  quality  standards (and the
    specific  water quality criteria) is the orderly development and improve-
    ment of the  nation's water resources, guaranteeing their long-time
    preservation for industrial, municipal,  and  agricultural uses, for recre-
    ation, and for esthetic enjoyment.

     Unquestionably many agricultural pollutants  affect recreation
through  alteration  of water  quality and degradation of fish and
aquatic life.  The more serious polluting agents we judge to be eroded
soil, nutrients, and  pesticides. We sense  there is an awareness and
appreciation of the problem among those concerned with agriculture.

                    CHAPTER 24 / POLLUTANTS AND RECREATIONAL USES  / 341

 While the problems relating to agricultural pollution are complex,
 and the solutions will not easily be attained, it seems reasonable that
 in many  instances alternative  procedures can be developed.  Pollu-
 tion control measures are available (e.g., pesticides) which will allow
 continuation of agricultural production  and enhance and  protect
 water quality and recreation. While these procedures may be costly to
 apply, the expenditure  should be judged in light of its contribution
 toward the preservation  of man's environment.  Especially  in  the
 instance of pesticide use, protection of water quality may be requisite
 to protection of the health of man from unknown long-term effects of
 pesticides. Reduction and control of  agricultural pollutants are es-
 sential to develop and maintain a high-quality environment.  Quality
 of life and quality of environment are synonymous.

American Chemical Society.  1969.  Cleaning our environment:  the
     chemical basis for action. Am. Chem. Soc.. Wash., B.C.
Bartsch, A.  F.  1961.  Induced eutrophication—a  growing water re-
     source  problem. In Algae and metropolitan wastes, pp. 6-9. U.S.
     Dept. of Health, Education and Welfare.
Beeton, A. M.  1965. Eutrophication of the St. Lawrence Great Lakes.
     Limnol. Oceanog. 10:240-54.
Bennett, C.  W., Thompson. D. H., and Parr, S. A. 1940.  A second
     year of fisheries investigations at Fork Lake, 1939. Lake Man-
     agement Kept. 4.  III. Nat. Hist. Sari;., Biol. Notes 14:1-24.
Brown, C. B.  1945. Floods and fishing. Land 4:78-79.
Buck,  D. H. 1956.  Effects of turbidity on fish and fishing.  Trans.
     21st North Am. Wildlife  Conf., pp. 249-60.
Burdick, G.  E., Harris,  E. J., Dean, H.  J., Walker. T. M., Skea, J., and
     Colby,  D.  1964.  The accumulation of DDT in  lake  trout and
     the effect  on reproduction. Trans. Am.  Fisheries Soc. 93:127—
Butler, J. L. 1964.  Interaction of effects by environmental factors
     on primary productivity  in ponds  and microecosystems  Ph.D.
     thesis,  Okl'a. State Univ. Graduate School.
Chandler, D. C.  1942. Limnological studies  of western  Lake Erie.
     II.  Light  penetration and its  relation  to  turbidity.  Ecology
Chandler, D. C.,  and  Weeks, O.  B.   1945.  Limnological studies of
     western Lake Erie.  V.  Relation  of limnological and meteor-
     ological conditions to  the production of  phytoplankton in  1942.
     Eco/. Monographs 15:435—57.
Cope. O. B., and Springer, P.  F. 1958.  Mass control  of insects:  the
     effects  on  fish and wildlife. Bull. Entomol. Soc. Am. 4:52-56.
Cordone, A.  J., and Kelley, D.  W.   1961. The influences of inorganic
     sediment  on  the  aquatic life of  streams.   Calif.  Fish  Game
Cross,  F. B., and Braasch, M.  1969.  Qualitative changes in the fish-
     fauna of the upper Neosho River system, 1952-1967.  Trans.
     Kans. Acad.  Sci. 71:350-60.
Engelbrecht, R. S., and Morgan,  J.  J. 1961.  Land  drainage as a
     source  of  phosphorus in  Illinois  surface  waters.  In Algae and


     metropolitan wastes, pp.  74-79.  U.S. Dept. of Health,  Educa-
     tion and Welfare.
 Federal Water Pollution Control Administration.  1968.  Water qual-
     ity criteria.  Report of the Nat. Tech. Advisory Committee to the
     Sec. of the Interior.
 Ferguson, D. E.  1967.  The  ecological consequences  of  pesticide
     resistance in fishes. Trans. 32nd North Am. Wildlife Nat.  Re-
     sources Conf., pp. 103-7.
 Ferguson, D. E., Culley, D. D., Cotton, W. D., and Dodds, R. P.  1964.
     Resistance to  chlorinated  hydrocarbon  insecticides  in three
     species of freshwater fish.  BioScience 14:43-44.
 Hasler,  A.  D.  1947. Eutrophication of lakes by  domestic sewage.
     Ecology 28:383-95.
 Herbert, D. W. M., and Merkens, J. C. 1961.  The effect of suspended
     mineral solids  on  the survival of trout.  Intern. J. Air Water
     Pollution 5:46-55.
 Hunt, E. G., and  Bischoff, A.  I.  1960.  Inimical  effects on wildlife
     of periodic DDD applications to Clear Lake.  Calif. Fish Game
 Jamison, V. C., Smith, D. D.,  and Thornton, J. F. 1968.  Soil and
     water research on a clay pan soil.  USDA, ARS Tech. Bull. 1379.
 Johnson, D. W.  1968.   Pesticides and  fishes—a  review of selected
     literature. Trans. Am. Fisheries Soc. 97:398-424.
 McKee, J. E., and  Wolf, H. W.  1963. Water quality criteria. 2nd ed.
     Calif. State Water Quality Control Bd. Publ. 3-A.
 Mackenthun, K. M.  1965.  Nitrogen and phosphorus in water.  An
     annotated selected bibliography of their biological effects. Pub-
     lic Health Serv. Publ. 1305.
 	.  1968.  The phosphorus problem.  /. Am. Water Works Assoc.
 Mackenthun, K. M., Ingram, W. M., and Forges, R.  1964. Limno-
     logical aspects of recreational lakes. Public Health Serv. Publ.
 Miner, J. R., Lipper, R. I., Fina, L. R., and Funk, J. W.  1966.  Cattle
     feedlot runoff—its  nature and variation. /.  Water  Pollution
     Control Federation 38:1582-91.
 Moyle, J. B. 1956.  Relationships between the chemistry of Minne-
     sota surface waters and wildlife management.  /. Wildlife Man-
     agement 20:303-20.
 Piest, R. F., and Spomer, R. G.  1968. Sheet and gully erosion in  the
     Missouri  Valley loessal region.  Trans. Am.  Soc. Agr.  Engrs.
 Radtke, L. D., and Turner, J. L. 1967.   High concentration of total
     dissolved  solids block  spawning  migration  of  striped  bass,
     Roccus saxatilis, in the San  Joaquin River, California.  Trans.
     Am. Fisheries Soc. 96:405-7.
Ray, A. D.  1965.  Pollution from  industrial wastes and sewage.  In
     Water Forum, pp. 31-36.  Spec. Rept. 55, College of Agr., Univ.
     of Mo., Columbia.
Rudd, R. L.  1964.  Pesticides  and the  living  landscape. Madison:
     Univ. of Wis.  Press.
Sawyer, C. N.  1947. Fertilization  of lakes by agricultural and urban
     drainage.  /. New Engl. Water Works Assoc.  6.1:109-27.
Sawyer, C. N., Lackey, J. B., and Lenz, R. T.  1945.  An investigation
     of the odor nuisances occurring in  the Madison lakes, particu-


     larly Monona,  Waubesca and Kegonsa from July 1942 to July
     1944. Kept, of the Governor's Committee, Madison, Wis.
Smith, S. M., and Miner, J. R.  1964.  Stream pollution from feedlot
     runoff. Trans. 14th  Ann. Conf.  Sanit. Eng.  Univ. of  Kans.
     Publ., Bull, of Eng. and Architecture 52:18-25.
Swingle, H. S.  1949.  Some  recent developments in  pond  manage-
     ment. Trans. 14th North Am. Wildlife Conf., pp. 295-310.
Sylvester, R. O.  1961. Nutrient  content  of drainage  water from
     forested, urban and agricultural areas.  In Algae  and metropoli-
     tan ivastes, pp. 80-87.  U.S. Dept. of Health,  Education and
Task Group Report.  1967. Sources of nitrogen and  phosphorus  in
     water supplies. /. Am. Water Works Assoc. 59:344-66.
Trautman, M. B.  1957. The fishes of Ohio.  Columbus: Ohio State
     Univ. Press.
U.S. Bureau of the  Census.  1962.  Statistical abstract of the United
Vinson, S. B., Boyd, C. E., and Ferguson, D. E.  1963. Resistance to
     DDT in the  mosquito  fish,  Gambusia affinis.   Science  139:
Wallen, I. E.  1951. The  direct effect of turbidity on fishes.  Bull.
     Okla. Agr. Mech. College 48:1-27.
The White House.  1965.  Restoring the quality of our environment.
     Rept.  of the Environ. Pollution  Panel, President's Sci.  Advis.



      • HE state of Iowa has been a national leader in water pollu-
 tion control for nearly half a century.  Its first stream control law,
 passed in 1923, gave the State Department of Health regulatory and
 enforcement authority. At the time the 1923  law was passed, almost
 200 municipal sewage treatment plants were  already in operation
 (Iowa  Water  Pollution Control Commission, 1969).  These plants
 were in the smaller towns and served 350,000  persons, or  30%  of
 the population connected to municipal sewers.
     The pollution control  law has been revised twice, the latest re-
 vision being enacted in 1965. This legislation created the Iowa Water
 Pollution Control Commission as the policy-making body in Iowa's
 water pollution activities.   Present stream water quality regulations,
 in effect, necessitate secondary treatment (85  to  90% removal  of
 BOD) on all interior streams.1 Plant construction has  steadily pro-
 gressed so  that as  of January 1,  1969,  there  were 510  municipal
 water pollution control plants in operation  or under construction, and
 the population served by treatment had increased to 99.3% of the
 sewered population.  Municipalities not presently treating their wastes
 are smaller communities which now have plants in the planning  or
 construction stage.   One hundred  percent of the medium size and
 larger communities had sewage treatment facilities at the beginning
 of 1969. This record of water pollution control ranks Iowa with the
 most progressive states in the nation.
     Of the industries,  Iowa's meat-packing plants represent the
largest potential source of industrial water pollution.  Every meat-
 packing plant in the  state has a treatment plant in operation or under
 construction (Iowa Water Pollution Control Commission, 1969), and
 this represents some 3.5 million population equivalent being treated.

     E.  ROBERT BAUMANN is Professor, Department of Civil Engineering,
     Iowa State University.  SHELDON KELMAN is Assistant Professor, De-
     partment  of Civil Engineering, Iowa State University.
     1.  In  October 1969, the FWPCA adopted national regulations requir-
     ing Iowa to provide secondary treatment also for all wastes discharg-
     ing to  both the Mississippi and Missouri rivers.



With the exception of packing plants on border streams, all packing-
plant wastes receive at least secondary treatment.
     Although this municipal and industrial waste treatment record
is impressive, much work remains before  the quality of the water in
our  streams is adequately protected.  Some cities have grown to  the
point where their treatment facilities are undersized and/or obsolete.
Most secondary treatment facilities do not provide sufficient treat-
ment efficiency in the wintertime (only 65 to 75%  removal of BOD)
due  to the effects of cold weather on the efficiency of biological treat-
ment on trickling filters.  Recently, some  authorities have been call-
ing for still higher levels of wastewater treatment to alleviate water
pollution  problems. Iowa's  interior  streams are  characterized in
much  of the state by extremely low  minimum flows and  relatively
high summer temperatures.  At Ames, for example, the design flow
in the Skunk River (7-day low flow with  a frequency of occurrence
of once in 10 years) is  only 0.1 cfs while the Ames waste  discharge
currently approximates  5.0 to 6.0 cfs.  Even when low flow augmen-
tation is  available from  the proposed Ames reservoir,  the design
stream flow will be increased to only  30 to 40 cfs.  Under these con-
ditions—typical of those for many of Iowa's cities and industries—
maintenance  of  current water quality standards  in Iowa streams
will  require better waste treatment. The  treatment needs  indicated

i.   increased carbonaceous BOD removal
"2.   increased oxidation of the ammonia  in the treated waste to ni-
3.   increased phosphate removal

The  increased removal of carbonaceous BOD is required to maintain
the oxygen level in the  stream to protect  fish life.  The oxidation of
ammonia to nitrate is required to reduce  the ammonia levels in  the
stream below levels toxic to fish. Unfortunately, the increased oxida-
tion  of ammonia  to nitrate, together with the availability of -phos-
phates in the stream, results in significant algal growths. Such algal
growths can have  a detrimental effect on water quality because:

1.   they increase the carbonaceous BOD  in the stream and may re-
     sult in significantly lowered DO levels at night
2.   they increase the  turbidity and suspended  solids load in  the
     stream and can impart tastes and odor to the water

Before increased treatment requirements (advanced waste or  tertiary
treatments) are imposed on cities and industries, it would appear de-
sirable to consider first whether such wastes are the more significant
contributors of carbonaceous BOD, nitrates, and phosphates to Iowa
streams. Agriculture which  contributes  carbonaceous  BOD  from
animal wastes, plant residues, etc., and phosphates and nitrates from
the above sources  and  from  fertilizer  applications may  contribute
such significant quantities of similar pollutants on an  uncontrolled
area basis as  to negate any  desirable effect of increased municipal
and/or industrial waste  treatment.


     This chapter is designed to explore the relative effects of both
 municipal-industrial and agricultural wastes as they affect treatment
 requirements which may be imposed on municipal use of the stream
 for receiving treated wastes.

     Each treatment process achieves certain end results. For many
years all Iowa municipal and industrial wastes have received primary
treatment. Primary treatment is commonly a settling process which
removes floating and settleable  material, including a portion of the
suspended solids  and its  associated organic carbon.   The  organic
carbon is measured by the amount of oxygen bacteria required  to
oxidise it in  a fixed time period under aerobic conditions.  Thus the
organic carbon removal is  typically described as a  reduction in bio-
chemical oxygen demand or reduction in BOD.- (measured in 5 days
at 20° C). Primary treatment will remove  over 95% of the floating
and settleable solids, from  60 to 70% of suspended solids,  and from
30 to 40% of BOD5. Since much of the organic carbon is present  in
solution or in colloidal suspension, most (60 to 70%) is not remova-
ble by sedimentation.
     All Iowa  municipal-industrial  wastes will  ultimately  require
secondary treatment  prior  to discharge to  Iowa's  surface  waters.
Secondary treatment normally employs a biological process utilizing
bacteria and other simple  forms of life to convert much of the re-
maining  suspended and soluble organic  carbon  to biological cell
protoplasm and energy.  These  process units are then followed by
sedimentation tanks to remove the cells produced.  Typical domestic
effluents after secondary treatment contain only 10 to  20% of their
original  suspended solids and BOD and involve pollutant removal
efficiencies of 80 to 90%  under ideal conditions.
     Primary and secondary treatments (complete treatment) remove
only part of the wastes in wastewater.  The organic material present
is partially oxidized and partially converted into settleable cell pro-
toplasm. This reaction can be represented empirically as2

    COHNSP + O.	> protoplasm + CO2 + H=O + NCV + NOr
                            + NH.,+ + SOr' + POr" + organic P

Thus, it can  be  seen that although secondary treatment removes 80
to 90%  of the organic carbon and reduces the oxygen demand on
the stream which receives the treated wastes, these wastes will con-
tain significant amounts  of  NO2, NO3,  and NH3 as  well as phospho-
rous compounds, substances that are also applied to land in the form
of common fertilizers. These compounds have become increasingly im-
portant since in recent years the widespread use of phosphate builders
in detergents has more than doubled the concentration of phosphorus
     2.  Note:  C = carbon; O = oxygen; H = hydrogen; N= nitrogen;
     S = sulfur; P = phosphorus.


in municipal wastewaters. In addition, significant concentrations of
nonbiologically degradable organics will remain in the wastevvater.
     Nitrogen and phosphorus in municipal and industrial wastewa-
ters add to the concentrations of these fertilizer elements in surface
waters.  Such nutrient enrichment of water is termed eutrophication
and is of concern since it stimulates algal growth. This  process can
be represented as

  POr" + NO3- + CO, + sunlight 	> COHNSP (Algae protoplasm)

Algae thus  defeat the purpose of secondary treatment  by creating
more organic carbon (which we have just removed by primary and
secondary treatment) in surface waters.
     In the past 10 years attention has been focused on the problems
created by such  eutrophication.  Many studies have  been made of
the various methods intended to solve  this problem by  achieving a
higher  degree of wastewater treatment. These methods employ an
additional treatment step, usually termed tertiary treatment.  Ter-
tiary treatment is designed to achieve either a better degree  of sus-
perded solids and BOD removal and/or the removal of nonbiologi-
cally degradable organic and inorganic compounds, especially nitro-
gen and phosphorous compounds.  Tertiary treatment, now economi-
cally and technically feasible, can result in BOD and phosphate re-
movals of above  98%  and the effluent quality would be as good or
better than that of the  receiving stream.
     Among the processes which can be used singly or in combination
to achieve these  results are

\. lime or  alum  precipitation of phosphates
2.  air  stripping  of  ammonia at high pH levels
3.  activated carbon adsorption of dissolved organics
4. pulsed adsorption beds (PAB) for increased biological removal of
   dissolved organics
5. sand or diatomaceous earth filtration for removal of residual bi-
   ological  cells  and  suspended organics
6. ion exchange  for removal of specific cations and anions

     All of these processes are expensive relative to present  treatment
methods.  To put  such  a program into effect on a statewide basis for
all municipal and industrial wastewater will cost many millions of
dollars. Adding typical tertiary processes, such as lime precipitation
for  phosphorus removal and air stripping of ammonia,  could  triple
the cost of wastewater treatment for  a typical  Iowa city.

     To understand why there is a concern over the addition of ni-
trates and phosphates to surface water, it is first necessary to consider
its effect on water uses.  The presence of pyrophosphates and/or tri-


 phosphates has been shown to interfere with the efficiency of potable
 water treatment  processes,  including  the coagulation-flocculation-
 sedimentation  process and the lime-softening  process. These prob-
 lems become noticeable  (Task Group  Report, 1966)  at combined
 triphosphate and  pyrophosphate levels of 0.3 mg/1, measured as P.
 Shrobe (1967) found  that approximately  half the phosphates  dis-
 charged in treated wastewater are  orthophosphates and this ratio
 continues several  miles downstream from the treatment plant. Most
 of the remaining phosphate will be in the condensed forms discussed
 above.  Since typical Iowa rivers have  been found to  contain peak
 orthophosphate levels  approaching  3 mg/1 (Dept. of Civil Engineer-
 ing, Engineering Research Institute, 1969), it  can be assumed that
 similar levels of condensed  phosphates will also  be  present below
 sewage discharges, creating potable  water treatment difficulties.
     Two of the  principal nitrogen compounds, nitrates and am-
 monia, also cause problems.  The United States Public Health Service
 drinking water standards  limit the  concentration of nitrates in  po-
 table water to 45 mg/1 as nitrate. This limit is based on the fact that
 higher levels cause methemoglobinemia in infants.  In general, ni-
 trate levels in surface waters are  well below this limit; however,
 there are times when Iowa rivers contain more than 45 mg/1 nitrates
 (Engineering Research Institute,  1969).  Ammonia causes problems
 by increasing significantly the chlorine  demand  of water if an  un-
 combined or "free" chlorine residual is  required.  Ammonia is also
 detrimental to stream water quality  since it is toxic to fish. The Iowa
 water quality standards for fishing  streams limit the  allowable am-
 monia concentration  to  2 mg/1.
     The major water quality problem created by nitrogen and phos-
 phorus, however, is concerned with their stimulation of algae growth.
 Algae create problems in potable water treatment by clogging filters
 and  causing undesirable tastes and odors. These  problems were esti-
 mated in 1967 to  affect as much as  56# of total municipal  surface
 water supplies  in the United States (Task Group  Report, 1966).
     Algae  interfere with recreational use of rivers and lakes by  col-
 oring the water green and forming unsightly floating mats. Fish may
 be affected when large numbers of algae are present by lowering  sig-
 nificantly the DO  during the night.  This algae oxygen demand can
 deplete  the  dissolved  oxygen sufficiently to lead to fish kills. Fish
 may also find it hard to feed if the algae color the water and obscure
 their vision.
     The conservationist pressures to solve  the problems  created by
 eutrophication are becoming  greater every year. However, before we
 move toward requiring tertiary treatment of municipal and industrial
wastes to control nutrient discharges in agricultural regions, we need
to determine whether even 100% tertiary treatment of these  waste-
waters will correct the  problems of eutrophication or even contribute
 to the correction.  To accomplish this, we need to determine  the  rel-
ative nutrient contributions from various sources.



     Several potential sources of nitrogen and phosphorus which can
enter the surface waters of Iowa are readily recognized.  For years,
industrial and municipal wastes have been pointed out as the major
contributors of N and P. In recent years, attention has been focused
on the potential contribution  from surface runoff.  Corey et al. (1967)
have listed the following available nitrogen sources for plant growth:
soil  organic  matter, animal manure, legume fixation, commercial
fertilizer, and fertilizer naturally present in precipitation. The same
sources,  of course, are potential contributors of nitrogen to runoff.
     The average daily  N and P contribution of each person connected
to a sewer is well established.  Given sewered population data, the
quantities of nitrogen and phosphorus discharged in domestic waste-
water can readily be computed (an  example will be discussed  later
in this chapter). Other sources of these nutrients may equal or  even
exceed the quantities in domestic wastewater.  Industrial wastes, es-
pecially those from packinghouses, contain large quantities of nitro-
gen  and phosphorus, since meat is a protein containing high con-
centrations  of both  N and  P.
     Animal  wastes  form  an increasing source of such  nutrients.
Both wild and domestic animals are significant  sources of fertilizer
elements.  When  storm runoff occurs,  large  quantities of animal
wastes are washed into streams.  In Iowa, for example, we have  a
domestic population of 2.75 million.  This state is noted, however, for
its production of pork and beef. Approximately 6,100,000 swine and
3,300,000 beef animals are on feed  at any one time.  Nearly 46,000
cattle feedlots  are recorded  in Iowa.  The  daily waste from these
animals is equivalent to the dailv waste from a human population of
65 to 90 million people. Naturally,  not  all this  waste finds its way
into  our streams.  But  when it rains—as it does in Iowa  20  to 30
times per year with intensities of 1 inch/hour or more—several  days'
accumulation of these  wastes  may  find their  way into our surface
waters.  Feedlot regulations are now designed to control runoff  from
feedlots that feed  over 100 head of cattle.  Such animal wastes, to-
gether with dairy and poultry wastes, can be significant contributors
of carbonaceous and nitrogenous BOD. ammonia,  and nitrogen to
our  surface  waters.
     Another  increasing  source of  the  nutrients  which  enhance
eutrophication  is that  derived  from row  crop  agriculture.   Nitrogen
fertilizer use in the  United  States  has increased from 2.15 million
tons  in 1957 to 6.56 million  tons in 1968 (Sulphur Institute. 1969).
The  1957 figure represents 76 rc  of  total  United States consumption
of nitrogen at that time (Sauchelli, 1961).  During approximately the
same time period, phosphorus  use in fertilizers has risen from 0.99
million tons  in 1958 (Van Wazer. 1961) to 2 02 million  tons P in
1968 (Sulphur Institute, 1969). The 1958 figure represents 70%  of
that  year's total United States consumption of phosphorus.  Other
large uses of phosphorus include use  in  detergents (13.3%) and in
animal  feeds (8.4%).  Reportedly,  commercial  fertilizers  are the


 source of only 10 to 20% of the gross nitrogen available for agricul-
 ture (Corey et al., 1967;  Willrich, 1969).
     The United States as a whole is in a nutrient mining phase; i.e.,
 more nutrients are removed in the harvested crops than are replaced
 by all nutrient inputs, including commercial fertilizers.  This is us-
 ually true for most Iowa crops.  Corn can be an exception, however.
 High rates of nitrogen application in excess of 150 to 200 pounds per
 acre may cause a nitrogen accumulation situation rather than a min-
 ing situation depending on the amount of corn grain and stalks that
 are harvested.  If only the grain is harvested, about 1  to 1.2 pounds
 of nitrogen are required  per bushel of corn. If the whole plant is
 harvested for silage, about 2 pounds of nitrogen  are removed for
 each bushel of corn produced.
     TVA statistics (National Fertilizer Development  Center,  1968)
 indicate that the average 1968 Iowa application rate of nitrogen on
 corn  was 120 pounds N  per acre.  The  USDA Statistical Reporting
 Service stated the average Iowa corn yield for  1968 was 93 bushels
 per acre.

     For the past 2 years, the Sanitary Engineering Section of the
Engineering Research Institute at Iowa State University  has  been
making surface water quality studies between Boone  and Des Moines
as a part of a "Preimpoundment Survey of Water Quality in the Des
Moines River above Saylorville Reservoir."  This study is supported
by the Rock Island District, Corps of Engineers.  Among the param-
eters measured weekly are stream flow, BOD, COD, suspended solids,
turbidity, the various forms  of nitrogen, phosphates,  and  the  algal
     Total  algal counts have been extremely high as  shown by the
algal count data for 1968. For example, in the first 6 months of 1968,
the  phytoplankton count averaged  96,000  cells/ml,  ranging  from
17,000 to 281,000 cells/ml.  These values appear to be 5 to 10 

must be made of the nonsewered rural population, most of which use
septic tanks. At the present time, roughly half the Iowa population
of 2,783,000 can be classified as being rural (Iowa Natural Resources
Council,  1953).  Part of the wastes from this population eventually
reaches a stream by surface runoff or by illegal connections to drain
tiles, or enters the groundwater and reaches a stream during low flow
periods.  Since the  rural  population is rather  evenly  distributed
throughout the state, the ratio of the basin's rural population to the
state's rural population should be roughly equal to the ratio of the
two  areas.  The estimated rural population calculated on this basis is
135,600.  Some of the nutrients in the rural wastes are removed by
plant uptake or lost by denitrification, soil  adsorption, etc.,  and we
can  assume that only 50% reaches a  stream.  The  effective  total
population contributing wastewater to the Des Moines River is thus
171,500.  An average value for nitrogen in  wastewater is about 10
pounds per capita per year (Task Group Report, 1967).  At  this rate
we could expect roughly 4,700 pounds nitrogen per day or 860 tons
of nitrogen per year from domestic wastewater.  Similarly an average
value for phosphorus in wastewater is 3 pounds per capita per year,
which would result in a discharge of 1,410 pounds phosphorus per
day  or 256 tons per year in domestic wastewater in the basin. These
estimates assume typical low removal efficiencies of these constituents
in wastewater treatment.   In  the case  of nitrogen, the form  may
merely be changed,  i.e., to ammonia or to  nitrate.  In the  case  of
phosphorus, it may be converted  from organic phosphates to ortho-
     Major sources of industrial wastes in the basin are an anhydrous
ammonia plant and several packing plants.  Information from the
records of the Iowa State Department of Health indicates that ap-
proximately 2,300,000 pounds of live weight  of beef and hogs are
killed each weekday at the packing plants in the basin above Boone.
These plants will have losses of about 1 pound N and 0.1  pound P
per  1,000 pounds live weight killed, based on a U.S. Department  of
Health,  Education  and Welfare  publication (1954)  and based on
sampling  experience  of the authors.  These losses would thus  total
about 2,300 pounds N and 230 pounds P each weekday or 300 tons
N and 30 tons P yearly. The fertilizer plant contributes comparatively
little N and no P to the Des Moines River.  Most of these nutrients will
be discharged in treated wastewater to the Des Moines River.
     The possible losses of nutrients from growing corn, the dominant
crop in the 5,490 mi2 basin, can be estimated based on crop patterns.
The  harvested area of corn in  1968 in Iowa is believed to be about
10,200,000 acres (USDA Economic Research Service  and Statistical
Reporting Service,  1964), and  if the watershed is  assumed to have a
proportionate acreage in corn  the watershed corn acreage would  be
1,000,000 acres.  Timmons et al. (1968) have shown  that with  3
inches of runoff 20.3 pounds per acre of nitrogen was lost annually
from continuous corn plots. If this value is accepted for discussion
purposes,  then corn  acreage  in  the basin  could have  contributed
approximately 10,000 tons of nitrogen  annually  to the Des Moines
River above Boone.


     Timmons et al. (1968) also found that phosphorus losses with 3
 inches of runoff from continuous corn plots were 0.2 pounds per acre.
 If this value is accepted, then the 1,000,000 acres of cornfields in the
 watershed could lose 100 tons of phosphorus annually to  the river.
 Since corn is the predominant row crop in Iowa and  receives 95%
 of the applied commercial fertilizer (USDA, ERS and Stat. Rep. Serv.,
 1964), no  estimate was made  of the  nutrient contributions  from
 other crops or noncultivated land.
     A rough  estimate can be made of nutrient contribution  to the
 Des Moines River from animal wastes.  Loehr  (1969) has  presented
 data indicating hogs produce wastes with 0.05 pound N  and 0.03
 pound P2O5  per 100 pounds animal weight per day.  Beef cattle
 produce  wastes with  0.40 pound N and 0.12 pound PoO-,  per 1,000
 pounds animal weight per day.  Since we have 6,100,000 swine and
 3,300,000 beef cattle on feed in Iowa at any one time, it is possible
 to calculate their pollution potential. In making this estimate  it was
 assumed that the distribution  of animals was  uniform thrcughout
 the state, the average animal was half-grown, and 25%  of the ani-
 mal waste nutrients (a guess)  were  lost in runoff. Based on these
 assumptions,  the quantities  of  nutrients lost from  feedlots in the
 basin each year could well approximate 3,600 tons N annd 500 tons
 P. Additional quantities of nutrients are lost from poultry and dairy-
 ing operations but are not included in this estimate.
     The N and P estimated from these sources are tabulated in Table
 25.1.  These estimates can now be compared to the gross amounts
 determined  from actual stream data collected between  1967 and
 1969. Analyses for the various forms of  nitrogen and orthophosphate
 were  made  weekly from  samples  collected from  the  flowing  river
 water. No analyses were made of bottom sediments, but during high
 flow periods the water was highly turbid and  contained large quanti-
 ties of sediment. The analyses were performed  according to the pro-
 cedures outlined in "Standard Methods"  (American Public Health As-
 sociation, 1965).  Since the test for orthophosphate is  performed in
 an acid medium, part of the phosphate  adsorbed on  sediment is de-
 sorbed and  detected  by this method.
    The concentration of N and P in the Des Moines River  at Boone,
 based on these weekly  analyses, is shown in Figure  25.1, together
 with river flows. Similarly, Figure 25.2 shows the pounds per day of
N  and P in the river at Boone.
    During the first year of the study (1967—1968) rainfall was below

TABLE 25.1.   Estimated sources of nitroqen and phosphorus in the Des Moines
            River basin above Boone, 1968.

                   Ton N/Yr   Ton P/Yr    Lb N/Day    Lb P/Day
Domestic wastewater
Packinghouse wastes .
Animal \vastes 	
Agricultural losses . .
2 700
    Total  	   14,760        886        81,350       4,825

                        -DISCHARGE, cu ft/sec
                        -PHOSPHORUS,  mg/1	

                        -NITROGEN,  mg/l	
               FIG. 25.1.  Concentration  of  nutrients  and  flow  in  the  Des  Maine?
               River at Boone, Iowa.
  >! 54,0
                        -DISCHARGE, cu ft/sec

                        -PHOSPHORUS,  Ib/day -

                        -NITROGEN, Ib/day —
                                       WAI ' Af« 'MAY' JUN ' JUL ' AUG ' SEP ' OCT 'NOV' etc aiii'm 'MAI ' A/I
               FIG.  25.2.  Pounds of nutrients  per  day  and flow  in  the  Des Moines
               River at Boone, Iowa.

                    -DISCHARGE, cu ft/sec

                    -CHLOROPHYLL A, mg/1	
           J'Jl ' AUG ' 2f ' OCI 'MOV' DEC I JAN I Fll ' MAI '< AM 'MAY1 JUN ' JU. ' AUG ' SEP ' OC7 'MOV1 Of C I JAN 'HI ' M*l ' AM '
                 1967       I                1968               I         1969
            FIG. 25.3.   Chlorophyll A, turbidity, and  flow in the Des Moines River
            at Boone, Iowa.

       average, totaling 26.8 inches at Boone and  only 18.4 inches at the
       Des Moines airport.  As a result, runoff and river  flows during  this
       period were exceptionally low.  Nitrogen levels were lower than the
       estimated wastewater contributions for most of this year. Under these
       conditions of flow and  low turbidity, high concentrations  of chloro-
       phyll A,  a measure of algal activity, were found, as shown in Figure
       25.3. Figure 25.4 is a plot of the nitrogen  and chlorophyll data dur-
       ing winter to summer at this low flow period. The dashed lines at
       6,350 pounds nitrogen per day represent the amount of nitrogen ex-
       pected in the river water from  domestic and industrial wastewater.
            Several researchers  (Willrich, 1969)  have reported  drain  tile
       concentrations of 15 to 25 mg/1 of nitrate N and O.I mg/1 of P.  On
       this basis, we might conclude that groundwater entering the river
       during the low flow  periods might approximate these same levels of
       N and  P. Howrever, we have no real data to indicate the groundwater
       contributions of  N and P from this source,  and have attributed all of
       the N  and P in  the  river during low flow  periods  to municipal and
       industrial wastes.  The uptake  of nitrogen  by algae and  attached
       plants  can account for the fact  that the quantities of nitrogen ob-
       served  were lower than the estimated wastewater  contributions for
       extended periods.
            The second  year of the study (1968—1969) was a wet period dur-
       ing which rainfall totaled 37.8 inches at Boone.  During the second
       year, when runoff and the river flows were relatively low,  the nitrogen
       content of the river water was approximately at the level estimated to





f~\ II OPl^DIIVl 1 A , i i /! -— 	
L-nLUKUrli rLL A my/I — 	 • • •
	 NITROGEN, Ib/day 	

(6,350 Ib N/day)
i i1,
' wj s
FIG. 25.4.  Nitrogen, chlorophyll A, and  flow  in the Des Moines River
during a dry period.

                           DISCHARGE , cu ft/sec -

                          -CHLOROPHYLL  A, mg/1-

                          •NITROGEN, Ib/day	
                     AND INDUSTRIAL
                      WASTE WATER
                      JAN   I FEB  ' MAR I APR  I MAY  ' JUN '

FIG. 25.5.   Nitrogen, chlorophyll A, end flow in the Des Moines  River
during a wet period.


result from the continuous domestic and industrial wastewater con-
tribution. At times of high runoff and river flow,  the nitrogen con-
tent was correspondingly high.  Figure  25.5 shows the  relationship
between nitrogen content, flow,  and chlorophyll A during winter to
summer of the wet year. It is  apparent that during low flow periods
mestic and industrial wastewater is the principal source  of nitrogen.
During high runoff periods, large quantities of nitrogen are entering
the stream over extended periods of time; but  due  to high turbidity,
algal growth is  low.
     By assuming that the product of the weekly analysis  for nitrogen
and flow represented  the average weekly nitrogen load, it was possi-
ble to compute the weekly and  annual nitrogen  load to the river.  The
values for the weeks of incomplete  data were estimated.  The annual
nitrogen totals for the first and second years were 2,207 and 24,230
tons, respectively. During the first year, the first  2 weeks and  the
last week were exceptionally wet.  When these 3 weeks were elimi-
nated, the river nitrogen content during the remaining 49 continuous
weeks was 647 tons.  The nitrogen total during these 49 dry weeks
was 60%  of the estimated wastewater nitrogen. The nitrogen total
for the second wet year was 164%  of the estimated  combined waste-
water, animal waste, and agricultural loss  contributions.  Apparently
during a wet year the additional nitrogen derived from runoff is equal
to about 31 % of all the nitrogen contained in the fertilizer and animal
wastes generated that year in  the  basin. Annual gross nitrogen in-
puts originate from  many sources, including  mineralization of  or-
ganic matter, animal waste, commercial  fertilizers,  and that received
from the atmosphere by legume fixation,  soil  absorption, precipita-
tion, and dust sedimentation. The exact quantities derived from each
source have not been and cannot  be determined from  the data in
this study.
     BOD loads in the river also increased dramatically during periods
of peak runoff. During the dry first year the average BOD load was
28,100 pounds per day, the equivalent of untreated wastes from a
population of 165,000. During the  second wet year the average BOD
load in the stream was 127,000 pounds per day, the equivalent of
untreated wastes from a population of 750,000. The peak BOD dur-
ing the  second year,  experienced on March 26, was 916,000 pounds
per day. The carbonaceous BOD  (subtracting  the  oxidation of ni-
trogenous compounds) was equivalent to  untreated wastes from a
population  of 4,200,000.  These values  demonstrate the effect  of
runoff  on stream quality in a  watershed where the total population
is  238,000.
     Figures 25.1 and 25.2 also  show the concentration  and pounds
per day of phosphorus in the river at Boone.  Only orthophosphate in
the flowing water was measured and no analysis  was made of or-
ganic phosphate, phosphorus  in bottom deposits,  etc.  Figure 25.6
shows winter and summer data during this low flow period.  The
dashed line at 1,575 pounds P per dav represents the amount of phos-
phorus expected from domestic and industrial wastewater sources.
During this first year of the study, the dry conditions resulted in little
runoff.  Much of the phosphorus in  the wastewater either precipitated





Pill OPODIIVI 1 A in.- A
\~l\\.\Jf(\Jl 1 \ YLL A, Ilig/l
	 PHOSPHORUS, Ib/day 	

\ \
1 r
1 A ^ I A
A/ • • \
_^^ /^ _^ii
FIG.  25.6.   Phosphorus, chlorophyll  A,  and  flow in  the  Des Moines
River during a  dry period.


                              -DISCHARGE, cu ft/sec 	

                              •CHLOROPHYLL A, mg/1-

                              •PHOSPHORUS, !b/day	
                         DOMESTIC AND
                          JAN   I FEB  I MAR I  APR I  MAY  I JUN
    FIG.  25.7.  Phosphorus, chlorophyll  A,  and  flow in  the  Des  Moines
    River during a wet period.



  out or was utilized by algae (which were present in unusually  high
  concentrations), resulting in low levels in the stream.
      Figure  25.7 illustrates the river phosphorus content during the
  winter and spring of the year (1968-1969). During this wet period
  when runoff and river flows were high, phosphorus levels rose to as
  high as 19 times the level expected from wastewater alone. At low
  river flows the phosphorus levels were again well below the  domestic
  and industrial wastewater level.  Apparently the phosphorus is asso-
  ciated with channel scour and bottom sediments.  It is interesting to
  note that the phosphorus levels both increased  and decreased faster
  than the river flows,  indicating  that the phosphorus was bound  to
  sediment  particles.
     As with the nitrogen data, it was assumed that the product  of
  the weekly analysis for phosphorus and flow represented the average
 weekly phosphorus load in the river. The values for the weeks of in-
 complete data were estimated. The annual phosphorus totals for the
 first and second years were 50 tons and 1,653 tons, respectively.  The
 total for the first dry year is 18% of the  estimated domestic phos-
 phorus alone. The total for the second wet year  is nearly 6 times the
 estimated wastewater phosphorus alone and 186% of the estimated
 combined contributions from wastewater,  animal wastes, and agri-
 cultural losses.  The  additional  phosphorus in the  stream  derived
 from runoff during a wet year is equal to about 6% of all the phos-
 phorus contained in the applied  fertilizer and animal wastes gener-
 ated in  the  basin.
     Based on the assumptions made in this  analysis, it would appear
 that treating domestic and industrial wastewater to remove the nu-
 trients  will benefit the receiving stream only  during  dry  weather
 flows,  given present  inputs  from all  other  sources.  During  wet
 weather most of the nutrients in the water will originate from sources
 other than domestic and industrial wastewaters.

     The protection of the quality of water in Iowa streams requires
that attention be directed at the various contributors of the significant
pollutants.  Attention is currently being directed  at municipal  and
industrial waste  discharges, since  these enter streams  through  a
point source and are easily controlled. All such  wastes must be given
secondary treatment prior to discharge to Iowa's streams. As more
stringent treatment requirements are demanded in the future, there
is some question as to whether nutrient removals from municipal and
industrial wastes will be sufficient to protect the stream.
     This  study indicated that  during periods of dry weather when
light and turbidity conditions are favorable for phytoplankton growth,
the principal source of the N and P required to support such growth
is derived from municipal and industrial wastewater discharges.  Re-
moval of N  and P from such wastewater discharges will help reduce
phytoplankton growth.


     In periods of high stream flow, when turbidity  levels are high
enough to be unfavorable to phytoplankton growth, runoff from urban
and rural lands and channel erosion are probably the principal con-
tributors of N and P to the stream. Removal of N and P from mu-
nicipal and industrial wastes  during  these periods will not reduce
nutrient levels significantly.   However, these are not the  periods
when eutrophication is a problem in flowing streams. In those situa-
tions where the stream flow is impounded,  the  runoff sources pre-
dominate and the  clarified water in the reservoir will support large
phytoplankton blooms.  Under the latter  conditions, tertiary treat-
ment of municipal and industrial wastes will be  of less benefit until
runoff contributions of N and P  are  also controlled.

American Public Health Association.  1965.  Standard Methods  for
     the Examination of Water and Wastewater.  12th ed. New York.
Corey, R. B., Hasler,  A. D., Lee, G. F., Schraufnagel, F. H.,  Wirth,
     T. L.  Jan.  1967.  Excessive water fertilization.  Rept. to Water
     Subcommittee, Nat. Resources  Committee  of State Agencies,
     Madison, Wis.
Dept. of  Civil Engineering.  Oct. 1966-Sept. 1967.  Annual rept.
     Coralville project. Univ. of Iowa, Iowa City.
Engineering Research Institute.  Fiscal year 1968-69.  Preimpound-
     ment  water quality study,  Saylorville reservoir,  Des  Moines
     River, Iowa. Sanit. Eng. Sec., Iowa State Univ., Ames.
Iowa Natural Resources Council. 1953.  An inventory of water re-
     sources and water problems, Des Moines River Basin.  Bull. 1.
Iowa Water Pollution  Control Commission.  Apr.  1969.  Statement in
     support of the Iowa water quality standards  and plan for imple-
     mentation and enforcement, Mississippi River Basin.
Loehr, Raymond C.   1969.  Animal wastes—a national problem. /.
     Sanit. Eng. Div.  Am. Soc. Civil Engrs. 95(SA2): 189.
National Fertilizer Development Center.  1968.  Fertilizer summary
     data 1968.  TVA, Muscle Shoals, Ala.
Sauchelli, V., ed. 1960.  Chemistry and technology of  fertilizers.
     Am. Chem. Soc.  Monograph Ser. 148. New York:  Reinhold.
Shobe, William R. 1967. A study of diatom communities in a hard-
     water stream.   Unpublished Ph.D.  thesis,  Iowa  State  Univ.,
Sulphur  Institute. 1969.  Potential plant nutrient consumption in
     North America. Tech. BuU. 16, Wash., D.C.
Task Group 2610-P Report. 1966.  Nutrient-associated problems in
     water quality  and  treatment.  ].  Am.  Water Works  Assoc.
     58(10):  1337.
	  1967.  Sources of nitrogen and phosphorus in water supplies.
     J. Am. Water Works Assoc. 59 (3): 344.
Timmons, D. R., Burwell, R. E., and Holt, R. F.  1968.  Minnesota
     science.  Univ. of Minn. Agr. Exp. Sta. 24 (4).
USDA Economic Research Service and Statistical Reporting Service.
     1964. Fertilizer use in the United States.  1964 estimates, Sta-
     tistical Bull. 408.

362 / PART 5 / ACJRiCyiTUAAL POllUT ifcfl

U.S.  Dept.  of Health, Educaiion and Welfare.  1954.  An  i,'sdus'.riai
     waste guide to the 'meat industry. Publ. 386.
	.  1962.  Plankton 'population dynamics.  Nail.  Water  Onaliiy
     Network Suppl. 2, Public Health Serv.  Publ, 663.
Van  Wazer, V.  R.  1361. Phosphorus and its compounds,  ii. Tech-
     nology, biological functions, and applications. New York. Inter
     science Publishers,
Willrich, Ted L,  1969,  Personal communication.  Agr,  Er.g  E>:i ,
     Iowa State Univ., Ames.

                 POLLUTED AND  CLEAN WATER



     IN view of the preceding extensive technical discussions, it is
unnecessary  to emphasize the "iceberg"  character of  agriculture's
contribution to the pollution load carried by our waters.  The highly
publicized forms  of agricultural pollution,  such as the wastes from
concentrated feedlot operations, are analogous to the tip of an  ice-
berg—they signal the presence of a much larger mass of pollution
that exists just below the surface of visibility.  The law's present
concern with  agricultural pollution reflects and reinforces this  dis-
tinction between overt instances of pollution from identifiable point
sources and subtle, broad-gauge pollution from materials carried to
waterways through surface runoff and underground drainage from
agricultural lands. Thus far, the law recognizes and enforces some
duties in respect to agricultural pollution from point sources and has
created regulatory schemes to control this type of pollution.  Little
legal attention, however, has been devoted to the larger problem of
nonpoint pollution from land runoff containing animal and vegetable
wastes, agricultural chemicals, and silt.  Because this  book brings
together experts from  so  many disciplines, I will assume that  my
specific responsibility as  a representative of the legal  community
is twofold: (1) to describe what the law now requires and permits in
the area of water pollution from agricultural sources, and (2) to sug-
gest how the law might constrain or promote the implementation of
pollution policy changes affecting agricultural production.

     Under the common law doctrine of riparian rights, which was
uniformly adopted in the eastern United States, rights and duties  re-
lating to water use were incident to the ownership of  land on the
banks of a watercourse. Each riparian  owner was said to have a
right to use the water that flowed by for any beneficial purpose so long
as his use did not unreasonably interfere with the use of  the com-
mon watercourse by another riparian owner.  Stated in water quality
terms, the right of each riparian owner was said to be that of having

     N. WILLIAM HINES is Professor of Law, University of Iowa.


 the water flow by his land "unimpaired in quality."  This was inter-
 preted to mean that he had a right to receive the water in a quality
 state reasonably suitable for the use he  wished to make of it.  If
 some upstream riparian user was diminishing  the water quality be-
 low this level,  his pollution was actionable and the injured riparian
 could sue for damages or could seek to enjoin the polluting activity.
 The  technical form  of action under which such suits were brought
 is called nuisance; therefore, you frequently see this private law ap-
 proach referred to as the "nuisance theory."
     The essence of  private nuisance is an interference with a prop-
 erty owner's use and enjoyment of his  land, and under the riparian
 theory water rights are an incident to the ownership of riparian land.
 A nuisance may be either private or public, depending upon whether
 it harms only a few persons or affects the interests  of the general
 public.  For example,  stream pollution  that damages only isolated
 downstream users is a private nuisance, but if the pollution causes a
 fish kill, it is a  public nuisance.  If the nuisance is public, it subjects
 the polluter to criminal punishment, and actions to abate it may be
 brought by public officials (see Prosser, Laiv of Torts, 605-23 [1964]).
     In many states  the  nuisance concept has been legislatively  en-
 dorsed  and the procedures for remedying  the  situation specified by
 statute.  For example, Iowa Code provides:

    657.1  Nuisance—what constitutes—action to abate.
           Whatever is injurious or offensive  to the senses, or an obstruc-
           tion to the free use of property, so as  essentially to interfere
           with  the comfortable enjoyment of life or  property, is a  nui-
           sance, and a  civil  action by  ordinary proceedings may be
           brought  to  enjoin and abate the same and to recover damages
           sustained on  account thereof.

    657.2  The following are nuisances:
           (4) The  corrupting or rendering unwholesome or impure  the
              water of any river, stream, or pond, or unlawfully diverting
              the same from its natural course or state, to the injury or
              prejudice of others.

Relatively few reported cases can be found in which agricultural  pol-
lution has been attacked as either a private or public nuisance.  In
fact, no  higher  court case can be found where  a nuisance suit was
brought  to remedy an  injury  caused  by  water-borne  agricultural
chemicals or silt.  A look at a recent private nuisance action arising
in Kansas as the result of pollution caused by feedlot wastes should
serve to illustrate the application of the nuisance law.
     Some of you may be familiar with  the litigation which  reached
the Kansas Supreme  Court in 1968 under  the title Atkinson v. Her-
ington Cattle Company (200 Kan. 298, 436 P.2d 816). To my know-
ledge it is the only feedlot pollution case to be decided by a state su-
preme court in the modern era. Under the facts  alleged, farmer Cecil
Atkinson's water supply for  his  Grade A dairy  operation was Level
Creek and a well 400 feet distant from the creek.  The Herington
feedlots, on which were  fed as  many as  7,500 cattle, drained  into
Level Creek  1\4 miles upstream from Atkinson's farm, causing  the

                                     CHAPTER 26 / LEGAL ASPECTS  / 367

water, as it passed through Atkinson's property, to have a foul ma-
nure  odor and a dark yellow-brown color. The water in Atkinson's
well had many of the same properties and was grossly unfit for hu-
man  or  animal consumption.  Atkinson sued Herington and Swift
and Company jointly, claiming the latter party was a joint venturer
in the feedlot enterprise, the cattle being supplied by Swift and fed
by Herington on a contract basis.  Evidence was submitted by a  host
of expert witnesses,  including bacteriologists, chemists, geologists,
and sanitary engineers. Veterinarians testified that Atkinson's cows
had died  from nitrate poisoning.  The trial court awarded Atkinson
$29,060  actual damages and $7,500 punitive damages against  both
defendants jointly.  Punitive damages are awarded in cases where the
defendant is deserving of punishment for his willful and malicious
invasion of another person's rights.  The purpose of such an award
is to  make  an example of the defendant  and thereby deter others
from  the commission of like wrongs.
    On  appeal, the  Supreme  Court sustained  the actual  damage
award, but  denied  the punitive damages. The court  said  that al-
though there was conflict over the details of how the water  became
polluted  and the precise physical effects of the pollution, ample evi-
dence existed to support  the lower court's finding that Herington
had unreasonably polluted  Level Creek and that Atkinson's damages
resulted from that pollution.  The  court's statement of its ruling was
as follows:  "Runoff becomes a harmful substance when it combines
bacteria and chemicals in  such an amount as to produce excessive
pollution  resulting  in injury."  On the punitive  damage issue, the
court found the evidence  inadequate to support the  awarding  of
such  exemplary damages because  the evidence showed that  Hering-
ton took immediate, although as it developed ineffective, steps to try
to remedy the situation upon receiving the first complaint from At-
    Because in this  case the injured party recovered substantial
damages from the agricultural polluter, it should not be inferred that
such is always the result.  Nuisance cases are generally hard to win
for a variety of factors, including the difficulties of proving the source
and effect of the alleged pollution, the possibility that a complaint is
not made quickly enough  to protect  the right  asserted,  and  the  will-
ingness of the courts  to engage in  a balancing process which  pits the
social utility of the polluter's economic  activity against the personal
loss of the complainant. Additionally, courts of law are not  particu-
larly well suited to the determination of the typical pollution  suit.
The perceived inefficiency  in placing direct reliance on the  judicial
system to control pollution is  the reason we presently assign the
major responsibility in this area to public pollution control agencies.

     Public regulation  of pollution generally is carried on by control
agencies established by local,  state, interstate, and federal govern-
mental units.  Traditionally, pollution control  by state level agencies


 has been  the mainstay of water  quality regulation,  Nearly every
 state has a pollution control law administered by a separate agency
 of a special division within a larger agency (see Hines, Nor Any Drop
 to Drink: Public Regulation of Water Quality Part I: State Pollution
 Control Programs, 52 Iowa L. Rev. 186 [1966]).  The  concentration
 of control activity at the  state level continues to  hold  true today;
 however, Congress's enactment of the Water Quality Act of 1965 inter-
 jected the federal government very directly into the business of  pol-
 lution control on most of the nation's waterways. The 1965 Act re
 quired the establishment of federal water quality standards for all of
 the interstate waters in the country. As defined by federal authori-
 ties, nearly all larger streams, rivers, and lakes  are interstate waters.
 The 1965  Act authorized the individual states  to develop standards
 for the waters within their jurisdiction,  but required  that these stand-
 ards be acceptable to the federal government.
     If you read the newspapers, you  know  that the establishment
 of federal standards has led to a number of serious disputes between
 state pollution control  agencies and  the federal officials  responsible
 for approving the standards adopted  by the states.  Although a num-
 ber of  the states in the Missouri and Mississippi basin  have  en-
 countered  problems in obtaining approval  of their  standards, Iowa
 has been the chief antagonist  of the FWPCA's requirements for ac-
 cepting state standards. This  dispute  came  to a head when Secre-
 tary of Interior Hickel announced on October 29, 1969, that he was
 imposing federal standards on Iowa, the first action  of its  kind under
 the 1965 Act.
     This is probably not a good occasion to air in detail  the dispute
 between the Iowa  Pollution  Control  Commission and  the  FWPCA;
 however, it is worthy of note that the major issue is the requirement
 of secondary treatment for all sewage  discharged into  Iowa's 27 in-
 terstate streams. Other standards imposed relate to water tempera-
 ture,  phenols, and continuous disinfection of  all municipal waste.
 Only  the standard  limiting  phenol  levels  to one  part  per billion
 would appear remotely related to agricultural pollution. That stand-
 ard speaks in terms of phenols "from  other  than natural  sources,"
 so arguably phenols produced by decomposition  of vegetative agricul-
 tural wastes  may not be covered by the standard because, except in
 a rare case, it would be impossible to distinguish these  from phenols
 produced by decomposition of natural vegetation. On a more general
level, it is  worthy of note that the federal  Guidelines for Establish-
ing Water Quality Standards make  no specific mention of regulation
 to control agricultural pollution, nor has the FWPCA's application of
 these standards demonstrated  any  immediate concern  for  problems
of agricultural pollution, except in a research capacity.
    It thus seems  a safe conclusion that the current  furor around
the country over the establishment of  water quality standards  has
very Little to do with agricultural pollution.  Municipal and industrial
pollution are the immediate targets of the federal-state effort to peel
back the flood of pollutants.  Only when these obvious  point sources
of pollution are brought under control  is attention likely to shift to
cleaning up  agriculture's insidious wastes. Given the present rate

                                      CHAPTER 26 / LEGAL ASPECTS / 369

 of  success  in  controlling municipal and industrial wastes, it  will
 probably be some time before the spotlight shifts to agricultural pol-
 lution.  Two notable exceptions exist to the current disinterest in ag-
 ricultural pollution.  Pollution control agencies have taken a direct
 interest in  the regulation of feedlot wastes and chemical pesticides
 used in agriculture and  have attracted widespread attention.  Both
 of these areas deserve special comment.  The pollutional effects of
 chemical fertilizers and soil erosion have received much less atten-
 tion. No regulatory  interest has been  addressed to  the enormous
 volume of animal and  crop wastes periodically washed into streams
 by surface runoff from agricultural lands.  Some  experts suggest that
 shock loads of these organic materials that reach watercourses as  the
 result of heavy rains or rapid thaws pose the most serious agricultural
 pollution problems (see Morris, Pollution Problems in Iowa, Paper
 presented to the Iowa Academy of Sciences, April 18, 1969).
     Regulation of Feedlot Wastes

     At the outset  it should be noted that pollution resulting from
concentrated  feedlot operations  is a  type  of agricultural pollution
that is fundamentally different from the great bulk of agriculture's
contribution to the pollution load of  our waters.  Feedlot pollution
emanates from a readily identifiable source—it is point pollution—
and it is susceptible to the same types of treatment procedures as are
applied to municipal wastes and organic industrial wastes. The simi-
larity  of feedlot pollution to municipal and  industrial  wastes, phis
its severe pollutional effects (a feedlot  of 20,000 cattle is said to pro-
duce  a waste with a population equivalence of a city of over 300,000
people), no doubt explains the prevalence of efforts to regulate feed-
lot  wastes  in  the midwestern states where this type of  agricultural
practice is  popular.
     Several states have  enacted legislation regulating  the  waste
discharge practices of feedlots.  The  Kansas statute,  for example,
provides that any feedlot operator feeding more than 300 cattle, 100
hogs,  or 500  sheep must register with the health department and
provide pollution control faculties if needed to prevent pollution run-
off  from the premises (Kan. Stat. Ann.  47-1505).  Arizona requires
feedlot operations with more than 500 cattle  to obtain  a license and
co provide  reasonable  methods to dispose of  excrement  and control
drainage (Ariz. Rev. Stat. Ann.§§ 24-391-397).  In most other states,
the pollution control agencies are working to  promulgate regulations
relating to feedlot  wastes under their general power to make rules
and regulations necessary to perform their regulatory function. Con-
siderable discussion has centered on what a  good regulation should
provide (see Matthews, A Recommended Procedure  for Developing a
Model Feedlot Regulation, proceedings of Animal Waste  Manage-
ment Conference,  Cornell Univ., Ithaca, N.Y., 1969).
    Iowa's  recently promulgated regulation covering cattle feedlots
is worthy of note.  In  1968 hearings conducted around the state on
proposed feedlot regulations provoked  considerable interest in agri-


 cultural  circles. The 1969 session of the Iowa legislature  amended
 the pollution control law to require registration of all Livestock and
 poultry operations where a potential for water pollution exists. Un-
 der the amendment the Commission cannot  require waste disposal
 facilities unless it is determined that the registrants are in fact pol-
 luting water or may reasonably be believed to  threaten pollution. The
 thrust of the regulations recently issued under this amendment is to
 specify the feedlot situations in  which a pollution potential exists,
 and therefore  in which registration is required. The regulations re
 quire registration of cattle feedlots  which  confine  more than  1,000
 cattle or which contribute effluent to a  watercourse draining more
 than 3,200 acres above the lot, which watercourse is less distant than
 2 feet per head of cattle, or from which the runoff flows into an un-
 derground conduit or drainage well.  If the control agency determines
 that the registered feedlot is, or reasonably may be, a source of pollu
 don, then the feedlot is required to obtain  a  permit for disposal  of
 wastewater. Permits are granted on a showing of adequate  water
 pollution control facilities constructed in accordance with plans arid
 specifications approved  by the agency. The regulation specifies ter-
 races or retention ponds sufficient  to contain a surface runoff of  3
 inches  as  the minimum pollution control facility permissible.
     Iowa's regulations  seem  to  meet most of the objectives sug-
 gested by Matthew iri  his recommended model feedlot regulations
 One facet of the regulations that is not clear  relates to the enforce-
 ment of the registration and permit requirements.  What penalty is
 incurred  by failure  to  register a  feedlot required to be registered
 or failing  to obtain a permit under circumstances  where the regula-
 tions would require a permit?  If the feedlot  is actually  creating  a
 situation  of pollution, the Commission can issue an abatement order
 without relying on any violation of the regulations.  If only a  potential
 for pollution exists, it is not certain what steps, if any, the  Commis-
 sion can take to compel compliance with  the registration and permit
 requirements.  Presumably the permit required by the regulations is  a
 permit of the type covered by Iowa Code 455B.25, which  makes the
 construction of disposal systems or use of  a new waste outlet with-
 out permit an unlawful  act.  The feedlot  operator who does nothing
 about his wastes would  seemingly not be guilty of an unlawful act
 under this section.  Iowa Code 455B.24 concerning contempt citations
 for failure to obey orders of the Commission has been interpreted as
 referring to abatement orders based on proof of pollution. Orders are
 not ordinarily issued on  a showing of pollution potential. The injunc-
 tion power granted the agency under Iowa Code 455B.23 applies to
 situations in which a person is placing wastes in a location  where
 they will probably  cause pollution.  Under this provision the Commis-
 sion could apparently enjoin the waste disposal activity of a feedlot
operator if it appeared likely to cause pollution. This brief excursion
through the  enforcement section of the Iowa law proves nothing
more than the need to make sure regulatory schemes fit the enforce
ment pattern of the statute under which they operate.  If registration
is the key to controlling  feedlot pollution, it appears desirable to put
some teeth in the  registration requirement by  indicating the penalty
for failure to comply,

                                      CHAPTER 26 / LEGAL ASPECTS / 371

      Pesticide  Regulation

      Pesticides have been in the headlines for much of the last year;
 not so much as water pollutants as total environment pollutants.  The
 so-called  hard pesticides, DDT  and  the other chlorinated  hydro-
 carbons, have  received the lion's share of attention, with 2-4,D  and
 related herbicides taking a secondary position. The toxic effect of the
 existing chemical biocides on man is still hotly debated; less disputed
 are the obvious incursions that have been made on the food  chains
 and eco-systems of lower animals.
     Because they enter the environment at a multitude of contact
 points, the  most effective  method for controlling pesticide  pollution
 seems to be to regulate their initial use. Almost every state has some
 form  of pesticide registration law that requires the filing of  an in-
 gredient statement, the label, and  directions for use (see Iowa  Code §
 206.4). Many states have additional provisions regulating the use of
 pesticides by commercial applicators (see  Iowa Code §  206.5).  These
 latter statutes were primarily aimed  at assuring technical competency
 on the part of persons who  applied chemicals for  hire; they  do not
 reach the individual applying chemicals to his own property.
    In sharp contrast to this traditional approach of minimum regu-
 lation are the actions  taken by several states recently  in prohibiting
 the use  or  banning  the sale  of  certain pesticides thought to be
 dangerous.  Arizona declared a  1-year moratorium on the use of DDT
 and DDD  in January 1969.  In  April,  the Michigan  Agricultural
 Commission decided to ban the sale in the state of all products con-
 taining DDT. In  August, Wisconsin  created a Pesticide Review Board
 for the purpose of governing the use of pesticides.  Actions outlawing
 or phasing out DDT were taken by one house of  the legislature in
 California, Wisconsin, and Illinois (New York Times, Apr. 30, 1969, p.
 43; July 19,  1969, p. 30; Nov.  4, 1969, p. 6).  This bustling of state
 regulatory  activity presaged the  announcement by HEW Secretary
 Finch on November 13,  1969,  that  the federal  government had de-
 cided  to halt DDT use in this country within the next 2 years.
    To appreciate the nature and  extent of the federal power  in  this
 area it is necessary to understand the federal regulation of pesticides
 used in agricultural production. The Food and Drug Administration
 within HEW is authorized by Congress to safeguard the safety  and
 quality  of food products and  drugs distributed in interstate com-
merce.  If you remember your high school civics, you may recall that
 the federal government has no general police  power, but must regu-
late through  the specific enumerated powers granted by the  Constitu-
 tion.  Thus purely intrastate marketing of  agricultural products is not
 subject to the FDA requirements.  Since the Pesticide Chemicals in or
on  Raw Agricultural Commodities  Act of  1954,  pesticide  residues
have been one  of the elements  of food quality the FDA has been re-
sponsible for regulating.  The FDA, therefore, sets the  levels of toler-
ance of pesticide  residue that will be permitted on foodstuffs market-
ed in interstate commerce.
    Responsibility for the regulation of agricultural chemicals mar-
keted  in interstate commerce rests with the USDA under the Federal
Insecticide. Fungicide and Rodenticide Act (7 U.S.C. §  135).  This act


 makes it a criminal offense to sell any "economic poison" which has
 not been fully and accurately registered with the USDA. The secre-
 tary of agriculture determines what chemicals are economic poisons,
 and chemical biocides are uniformly so classified.  Any party seeking
 to register a chemical to be used on food crops must indicate the
 crops on which the chemical is to be used, the quantity to be used for
 each crop, and describe the exact procedure of use. Additionally, test
 data must be provided to show the safety  of  specific residues of the
 chemical in or on foodstuffs.  If no residue should be  left under cor-
 rect application procedures, the product is registered as a "no-residue"
 chemical.  If a residue subsequently  shows  up,  the  registrant has
 violated the act.
     This is  what happened in the great  cranberry snafu  of 1958.
 The herbicide  amenotraezole was registered  on a no-residue basis,
 then when  a residue  appeared because of improper use, FDA pan-
 icked the buying public by announcing the confiscation of 300,000
 pounds of cranberries.  Testing ultimately showed that relatively few
 of the cranberries  were contaminated and cranberry growers  were
 reimbursed $8.5 million for their losses. A similar problem arose in
 1963 concerning  endrin residue  on brussels sprouts.  Endrin  had
 been registered on a no-residue basis at a  time when  the testing de-
 vices could detect residues of no smaller amounts than 0.1 ppm. Im-
 proved testing  techniques enabled inspectors to find a 0.03-ppm resi-
 due and the products were pulled from the market.
     If the chemical will leave a residue, the product will  not be
 registered until a residue tolerance level has  been set by FDA.  The
 tolerance level is set on the  basis of information submitted to FDA
 by the applicant,  showing the expected amounts of residue, the effect
 of such  a residue on test animals, the pattern of normal use of the
 foodstuff, and a workable method  of analysis  for enforcing  the toler-
 ance level. If USDA is satisfied with the tolerance level set by FDA.
 the chemical is then authorized.
     Thus working in  concert, as  they apparently plan  to do, HEW
 and USDA can, through reduction of the permissible tolerance levels
 of DDT, eliminate its use in conjunction with agricultural  products.
 Critics of the federal government's past activities in pesticide control
 have asserted that one of the major weaknesses has been the lack of
 coordination  between  FDA and HEW (see Note, Agricultural Pesti-
 cides:  The Need For Improved Control Legislation, 52 Minn. L. Rev.
 1242 [1968]).  Perhaps the joint plan to phase out DDT signals a new
 era in effective cooperation between these  two agencies.
     Lawyers active in environmental defense  litigation claim that
 the most difficult problem in pesticide regulation is uncovering the
 scientific facts  about the relative toxicity of the different chemicals.
They suggest that this difficulty could be  substantially cured if the
regulatory agencies held public hearings where experts could be pro-
duced and  cross-examined on  the questions concerning the danger
of chemicals to various forms of  life.  The lawyers argue  that the
adversary process used in our courts is peculiarly suitable for testing
the reliability of the evidence presented by proponents and opponents
of controversial chemicals.  The one  experiment with this method be-

                                      CHAPTER 26 / LEGAL ASPECTS / 373

 fore a special hearing board in Wisconsin suggests that there is merit
 in the argument for adversary proceedings before agencies charged
 with protecting environmental quality. Until recently, such a require-
 ment would have been meaningless because only the chemical  in-
 dustry would have been represented.  Today, however, a number of
 citizen groups are ready and able to present the  case for the public
 interest in a  safe and  wholesome environment (see  Foster, Counsel
 for the Concerned Conference on Law  and the Environment,  Sept.
 11-12, 1969, Warrenton, Va.).
     Looking  to the future, the agricultural producer should expect
 much closer state and federal regulation of both the chemicals availa-
 ble to him  and his  procedures in applying them. Special pesticide-
 regulating agencies  are likely to be created by many states.  From a
 purely legal standpoint, given the range of uncertainty  concerning
 the long-range effects of pesticides on the environment,  almost any
 type of restrictive  state regulation is likely to be sustained as a  valid
 exercise of the state's police power. However, chemical biocides play
 such a major role in modern commercial  agriculture that  it is  in-
 conceivable that many  of the  other important chemicals will be  dealt
 with as harshly as DDT. Much more likely is regulation designed to
 encourage selective use of chemicals and substitution of softer chem-
 icals or biologic techniques for the more toxic  pesticides.  Because
 pesticides entering water directly from agricultural land are believed
 chiefly to travel adsorbed to  sediment particles  washed away by soil
 erosion, more careful attention to land-use practices is likely to be
 required by law.
     Chemical  Fertilizers

     It is frequently asserted that the increasing levels of nitrates
and phosphates in midwestern  waters are  caused by residues  from
chemical fertilizers carried to watercourses by runoff and percolation
of precipitation falling on agricultural cropland.  Some experts dis-
pute  this explanation, claiming  that  agricultural  fertilizers make
only a very  minor contribution to the  current high levels of nitrates
and phosphates compared  to the amounts contributed  by natural
sources and by organic effluents discharged by municipalities and
industries (see  Smith, Fertilizer Nutrients in Water Supplies, in Agri-
culture and the Quality  of Environment,  1967).   One point  upon
which considerable  agreement exists is that to the extent chemical
nutrients reach watercourses, they are principally transported  there
aboard soil particles lost through erosion.  Thus if chemical fertilizers
are proved to be a source of nitrate pollution, methods for controlling
silt pollution, discussed below, should have a double-barreled effect in
reducing the amount of chemicals washed into watercourses.
     At present the use of chemical fertilizers is subject to no  regu-
lation at either the  local  or national level.  The legal disinterest in
chemical fertilizers is typified by the Iowa Pesticide Statute which ex-
pressly provides that in products where pesticides and fertilizers  are
mixed, the fertilizer is to be treated as  an inert ingredient (Iowa Code


 § 206.4).  Looking to the future, if it is proved that substantial quan-
 tities of phosphates and nitrates from chemical fertilizers are wash-
 ing into streams, it should be anticipated that fertilizer use itself will
 be regulated.  Limits could be placed on the volume and  strength of
 chemical  fertilizers that can be applied  at one time or  during one
 season, and controls may be adopted governing the timing and meth-
 ods of  fertilizer application.  Such  regulations would not be different
 in kind from the controls now being exercised over pesticide use in
 some parts of the country.  There seems  little doubt that whether
 such controls  were necessary  to prevent nitrate poisoning  or to re-
 duce the  nutrient load carried by our waters,  and thereby  reduce pol-
 lution caused by nuisance aquatic  growths, it would be a valid exer-
 cise of the state's police power  (see Williamson v. Lee Optical, 348
 U.S. 483 [1955]; Brackett v.  City  of Des Moines,  246 Iowa 249, 67
 N.W.2d 542 [1954]).

     Recent studies on water  quality in the Mississippi River show
 that sediment from the over  16  million acres of Iowa  agricultural
 land which drains into the river will reduce  the recreational value of
 the river more rapidly than either municipal or industrial pollution.
 Soil erosion is  acknowledged  to  be the single  largest pollutant  of
 nearly all mid-continent streams  draining land  intensively used for
 agricultural production. Also, as noted earlier, soil erosion is thought
 to be the principal vehicle for transporting agricultural pesticides and
 chemicals from the  site of their application  to our waterways.  Only
 recently, however, has soil erosion been thought of in  terms of a
 water pollution  problem (see  Browning,  Agricultural  Pollution.