THE IMPACT OF AQUATIC PLANTS AND THEIR MANAGEMENT
TECHNIQUES ON THE AQUATIC RESOURCES OF THE
UNITED STATES: AN OVERVIEW
by
Jerome V. Shireman, William T. Haller
Daniel E. Canfield, and Vernon T. Vandiver
Aquatic Plants Research Center
118 Newins-Ziegler Hall
University of Florida
Gainesville, Florida 32611
Grant No. R-805497-02
Project Officer
Gerald E. Walsh
Environmental Protection Agency
Environmental Research Laboratory
Gulf Breeze, Florida 32561
ENVIRONMENTAL RESEARCH LABORATORY
OFFICE OF RESEARCH AND DEVELOPMENT
U.S. ENVIRONMENTAL PROTECTION A3ENCY
GULF BREEZE, FLORIDA 32561
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DISCLAIMER
This report has been reviewed by the Environmental Research Laboratory,
U.S. Environmental Protection Agency, Gulf Breeze, Florida, and approved for
publication. Approval does not signify that the contents necessarily relfect
the views and policies of the U.S. Environmental Protection Agency, nor does
mention of trade names or comroerical products constitute endorsement or recom-
mendation for use.
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FOREWORD
The protection of aquatic ecosystems from damage caused by toxic organic
pollutants and other pest control agents requires that control regulations be
formulated on a sound scientific basis. Accurate information describing
treatment-response relationships for organisms and ecosystems under varying
conditions is required. The EPA Environmental Research Laboratory, Gulf
Breeze, contributes to this information through research programs aimed at
determining:
.the effects of toxic organic pollutants on individual species and
communities of organisms;
.the effects of toxic organics on ecosystem processes and components.
Infestation of aquatic ecosystems by native and exotic plants has in-
creased dramatically in the past 10 years. Herbicides, herbivorous fishes
and other organisms that attack plants, and mechanical harvesting are often
ineffective in control of nuisance plants, especially for extended periods
of time. In addition, secondary effects of control may cause undesireable
physical, chemical, and biological characteristics of the aquatic system.
This report reviews methods of weed control and their secondary effects. It
suggests research needed for weed control that have minimal effect on the
quality of water and its use for human needs.
Henry f
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PREFACE
Aquatic systems in the United States have been studied for a long time,
and there is a voluminous quantity of information on the impact of aquatic
plants and their management techniques on the aquatic environment. It is
beyond the scope of this paper to summarize each study. We have, therefore,
attempted to select those studies which have documented various impacts of
aquatic plants or their management techniques on the aquatic environment.
Inclusion of any particular study in this review paper does not constitute
an endorsement of the results or conclusions. We have not examined data for
errors and have made no judgment as to quality of the studies. We report
only what is currently accepted in the published scientific literature.
Because results and conclusions from different studies often conflict,
readers of this paper should not accept the results or conclusions of any
particular study until additional scientific research clarifies the dis-
crepancies.
Throughout this paper there are many subjects which we have discussed in
the most general of terms. In these cases, we have not provided factual
information because information is lacking and most studies lack data to
support the generalizations made. In other instances, there may be volumi-
nous amounts of physical and biotic data available. However, because there
was often no rationale for collection of the data or appreciation of inter-
relationships between physical and biotic factors, most of these data have
not been analyzed. Without our own analysis, all we could possibly do is
list the data in tables which would not provide much additional insight
beyond our general description of the reported results.
As an additional comment, this review paper has made us aware of the
large amounts of information now available. We hope our review of the
literature on the impact of aquatic plants and their management techniques on
water quality and the aquatic environment will be of use to water resource
managers. We must, however, caution all water resource managers about
attempting to use results from studies cited in this paper to manage aquatic
systems that differ from the studied systems. There is an urgent need for
studies on large numbers of aquatic systems in order to document quantita-
tive patterns in the behavior of aquatic systems. Once these quantitative
empirical relationships have been obtained, water resource managers will be
able to manage our nation's aquatic resources with greater confidence.
Until these empirical relationships are available, we recommend caution
before a particular management strategy is adopted.
IV
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Contents
Foreword ...... iii
Preface iv
Abstract vi
Figures vii
Tables xi
Introduction 1
Problem Aquatic Plants 2
Control Methods in the Sunbelt 46
Water Level Fluctuations 51
Mechanical Control 53
Biological Control 54
Impact of Control Methods 55
References 110
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ABSTRACT
This paper provides an assessment of nuisance aquatic plants and the
problems associated with their control in the United States. Major emphasis
is given to the Sun Belt states where aquatic plant control is critical due
to introduction of exotic plants and extended growing seasons. The impact
of aquatic plants (algae, non-native, and native plants) and their management
techniques are discussed as they pertain to water quality and aquatic life.
Herbicide residue data, both in the soil and water, and herbicide toxicity to
aquatic organisms are presented and discussed. The chronic or long-term
effects of herbicide on aquatic organisms have not been fully investigated.
Current information indicated that non-fatal physiological changes might
cause mortality in test organisms if the organisms are exposed repeatedly.
Effects of vegetation removal are discussed. Major research needs are iden-
tified for development of environmentally safe aquatic plant management pro-
grams.
VI
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FIGURES
Number Page
1 Relationship between chlorophyll ji concentrations and measured
total phosphorus concentration for the EPA-NES natural and
artificial lakes. From Canfield, 1979 3
2 Relationship between summer levels of chlorophyll ji and measured
total phosphorus concentration for 143 lakes. From Jones and
Bachmann, 1976 5
3 Distribution of water hyacinths in the United States, 1978 6
4 Known distribution of hydrilla in Florida in 1960. The total in-
festation in both Crystal River and the Miami River was approxi-
mately 10 ha 6
5 Known distribution of hydrilla in 1978. Larger dots represent
dense infestations totaling some 40,000 ha, and smaller dots
indicate hydrilla common in the flora of an additional 200,000
ha of Florida's fresh water 7
6 Distribution of hydrilla in the United States, 1978 7
7 Distribution of Eurasian watermilfoil in the United States, 1978. 8
8 Relationship between mean Secchi disc transparencies for July
and August and the mean July-August chlorphyll & concentra-
tions for 16 lakes. From Bachmann and Jones, 1974 10
9 Double logarithmic plot of Secchi disk depths against average
chlorophyll zi concentrations. Data for non-Iowa lakes were
taken from the literature (Bachmann and Jones, 1976, Dillon
and Rigler, 1975, Oglesby and Schaffner, 1975) and from reports
of lake self-help projects in Ontario and Michigan. The addi-
tion of the constant 0.03 to the chlorophyll values prevents the
calculated Secchi disk values from approaching infinity as the
chlorophyll levels approach zero. From Jones and Bachman, 1978. 12
10 Light penetration through water and plant communities in Rodman
Reservoir. Light measurements in the hydrilla community were
made through small (20 cm2) openings in the canopy. From
Mailer and Sutton, 1975 13
VII
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Number
11 Mean depth and the average standing crop of plankton in twenty
lakes. From Rawson, 1955
12 Mean depth and the average weight of bottom fauna in twelve
lakes. From Rawson, 1955 .................. 18
13 Mean depth and the long-term average commerical fish production
in twelve lakes. From Rawson, 1955 .............. 18
14 Upper: Changes in dissolved oxygen in the littoral and open water
areas over a diel period in eutrophic Winona Lake, Indiana,
9 August 1922. (From data of Scott, 1924.) Lower: Vertical
stratification of oxygen within the littoral zone of Parvin
Lake, Colorado, 9 July 1955, in a luxuriant stand of the sub-
mersed macrophyte Elodea. (Generated from data of Buscemi,
1958.) From Wetzel, 1975 ................... 21
15 Relationship between the maximum chlorophyll a_ and minimum
Secchi disc transparencies in various prairie ponds (Erickson-
Elphinstone district of southwestern Manitoba). From Barica,
1975 .............................. 22
16 Diurnal fluctuations in free CO2» HCCvj", O2, and pH measured
in the surface 5 cm of water over a Hydrilla verticillata mat.
Data were collected on October 14 and 15, 1975 at Lake Killar-
ney, Florida. Figures in parentheses refer to the water tem-
perature (C) at the time of sampling. From Van et al., 1976. . 24
17 Depth-time diagram of isopleths of pH in hypereutrophic Wintergreen
Lake, Michigan, 1971-1972. Opaque area = ice cover to scale.
From Wetzel, 1975 ....................... 25
18 Depth-time distribution of isopleths of sodium concentrations
(mg Na+ I"*) of Lawrence Lake, 1972 (a), potassium concentra-
tions (mg IT" I"1) of hypereutrophic Wintergreen Lake, Michigan,
1971-1972 (b), chloride concentrations (mg CI~ I"1) of
eutrophic Little Crooked Lake, Whitley County, Indiana, 1964 (c),
and magnesium concentrations (mg Mg"1"* I"1) of hardwater
Lawrence Lake, Michigan, 1972 (d). Opaque areas = ice-cover to
scale. From Wetzel, 1975 ................... 26
19 Depth-time distribution of isopleths of calcium concentrations
(mg Ca"1"1" I"1) of hypereutrophic Wintergreen Lake, Kalamazoo
County, Michigan, 1971-72. Opaque area = ice cover to scale.
From Wetzel, 1975 ....................... 28
20 Depth-time diagram of the concentrations of ammonia (mg NH3~N
I"1), Rotsee, Switzerland, 1969-70. (Redrawn from Stadelmann,
1971). From Wetzel, 1975 ....... .' ............ 29
VIM
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Number Page
21 Comparison of in_ situ acetylene reduction throughout the day of
21 August 1968 by samples from various depths in Lake Mendota.
The ranges between replicates are indicated, and the mean
values are plotted. Numbers within brackets represent the depth
in meters at which the samples were collected. Samples were
incubated at the surface in a pail of water. From Rusness and
Burris, 1970 30
22 Variations in nitrogen fixation with depth (A) in Lake Windermere,
and (B) in Esthwaite Water, England, 30 August 1966. (Modified
from Home and Fogg, 1970.) From Wetzel, 1975 33
23 The periodicity of the diatom algae Asterionella formosa, Fragi-
laria crotonensis, and Tabellaria flocculosa in relation to
fluctuations in the concentration of dissolved silica, 0.5 m
in Lake Windermere, England, 1945-1960. From Lund, 1964. ... 33
24 Correlation between growing season mean photosynthesis (per unit
volume euphotic zone) and mean total phosphorus concentration
for 38 north temperate lakes. From Smith, 1979 39
25 Correlation between growing season mean photosynthesis (per unit
volume euphotic zone) and mean chlorophyll concentration for 49
North American lakes. Limits shown are 95% confidence limits
for individual points around regression line. - Lake Wash-
inton: - Tuttle Creek Reservoir; - European lakes.
From Smith, 1979. 40
26 Lake mean zooplankton abundance versus mean chlorophyll «i
for data from the literature and this study. Values from
Patalas (1971) include mean depth data from Brumskill and
Schindler (1971) and chlorophyll & data from Armstrong and
Schindler (1971). From Noonan, 1979 45
27 A drawdown scheme which will provide hydrilla control in North
Florida 52
28 Comparison of 2,4-D residual in water and the number of gallons
of 2,4-D applied monthly in the St. Johns River, Florida.
Open bars represent cumulative gallons of 2,4-D, closed bars
represent 2,4-D residual in ppb, and ( ) represents that dis-
tance upstream from mouth in river miles. From Joyce and
Sikka, 1977 60
29 Comparison of 2,4-D residual in water and the number of gallons
of 2,4-D applied daily in the St. Johns River, Florida. Open
bars represent number of gallons of 2,4-D applied, closed bars
represent 2,4-D residual in ppb, and ( ) represents distance
from mouth in river miles. From Joyce and Sikka, 1977 61
IX
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Number Page
30 Endothall residues in water and the top 1-inch of hydrosoil of a
treated farm pond, with time. The bars represent the range of
duplicate values. From Sikka and Rice, 1973 71
31 Endothall residues in water and hydrosoil of r ^uaria treated with
2 and 4 ppm of the herbicide. The bars represent the range of
duplicate values. From Sikka and Rice, 1973 71
32 Chart of the major metabolic and degradative routes of dalapon:
the starred acids will be in equilibrium with their anions with
the position of the equilibrium depending on pH and the specific
cations in a particular environment and the compounds in brackets
will be transient. From Kenaga, 1973 76
33 The effect of size of sunfish on time of response to sarin (10 ppb)
and oxygen consumption. From Weiss and Botts, 1957. ..... 94
34 The effect of time in laboratory stock tanks on the time of res-
ponse of sunfish to sarin (10 ppb) and oxygen consumption.
From Weiss and Botts, 1957 95
35 Effects of herbicide applications and the destruction of submerged
plants likely to be of consequence in determining faunal
changes. From Brooker and Edwards, 1975 101
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TABLES
Number Page
1 Classification of English lakes on the basis of their physical
characters and rooted vegetation. From Pearsall, 1929. ... 4
2 Percentile absorption of light of different wavelengths by one
meter of lake water, settled of particulate matter, of
several Wisconsin lakes of progressively greater concentra-
tions of organic color. From Wetzel, 1975 14
3 Transpiration for water hyacinth in relation to evaporation.
From Penfound and Earle, 1948 16
4 Comparisons of pH, temperature, dissolved CL, and dissolved 002
in hyacinth and "open water" areas. March - October is
the growing season of Eichhornia in Gainesville, Florida.
From Ultsch, 1973 21
5 Carbon dioxide: vertical variation 1928. From Costing, 1933. 24
6 Odor, tastes, and tongue sensations associated with algae
in water. From The Practice of Water Pollution Biology,
1969 34
7 Comparison of rates of primary production of phytoplankton in
selected fresh waters of varying fertility and representative
estimations of annual above-ground biomass and productivity of
aquatic macrophytes. From Wetzel, 1975 36
8 Examples of annual net productivity of phytoplankton, littoral
algae, and macrophytes of several lakes in which productivity
estimates of attached algae were made on natural substrata.
From Wetzel, 1975 41
9 Effect of dense aquatic plant growth on abundance of phytoplank-
ton. From Hasler and Jones, 1949 42
10 Number of organisms per kilogram of different types of aquatic
plants. From Andrews and Hasler, 1943 43
11 Residues of the dirnethylamine salt of 2,4-D in water (mg/l);
hydrosoil (mg/kg), and fish (mg/kg) from ponds in Florida and
Georgia treated with 2.24, 4.48, and 8.96 kg 2,4-D per hectare.
From Schultz and Gangstad, 1976 59
XI
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Number
Page
12 Physical, chemical and toxicological properties of some aquatic
herbicides. From Newbold, 1975 ............... 62
13 Analyses of water samples, Watts Bar Reservoir, Gordon Branch
Bnbayment. From Smith and Isom, 1967 ..... ....... 63
14 Analyses of water samples, Guntersville Reservoir, Vicinity of
Comer Bridge. From Smith and Isom, 1967 ........... 64
15 2,4-D analyses - Watts Bar and Guntersville Reservoirs. From
Smith and Isom, 1967 ..................... ^
16 Physical, chemical and toxicological properties of some aquatic
herbicides. From Newbold, 1975 ............... 66
17 Details of residue trials 1, 2, and 3, Dade County, Florida,
1966. From Mackenzie, 1969 ................. 67
18 Water samples - diquat residues as related to time after treat-
ment and depth in residue trials 1,2, and 3. From Mackenzie,
1969 ............................. 67
19 Diquat residues in elodea based on dry weight as related to
time after treatment in residue trials 1, 2, and 3. From
Mackenzie, 1969 ........................ 67
20 Bottom soil samples - diquat residues based on dry weight as
related to time after treatment in residue trials 1,2, and
3. From Mackenzie, 1969 ................... 68
21 Florida elodea control - results in residue trials 1 , 2 , and
3 treated with 0.5 ppmw diquat, 1966. From Mackenzie, 1969. . 68
22 Disappearance of endothall from laboratory aquaria containing
various combinations of tap water, lake water, 4 to 6 live
fish, plant debris, and mud. From Hiltibran, 1962 ...... 70
23 Disappearance of endothall from plastic-enclosed test plots
of aquatic vegetation in farm ponds. From Hiltibran,
1962 ............................. 70
24 Toxicity of commonly used aquatic herbicides to bluegill
fingerlings. Mortality is cumulative across the table.
From Haller, unpublished data ................ 73
25 Copper in solution after treatment of Inglis Reservoir.
Samples for stations 1 to 4 are from the area treated with
diquat plus cutrine and stations 5 to 8 from the diquat plus
CSP (T = top; B = bottom).. From Mobley et al., 1971 ...... 74
XII
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Number Page
26 Copper content of hydrilla after treatment of Inglis Reservoir.
Plant samples for stations 1 to 4 are from the area treated
with diquat plus cutrine and stations 5 to 8 from the diquat
plus CSP (T = top; B = bottom). From Mobley et al., 1971. . . 74
27 Loss of herbicide from the chemical reach. From Brooker, 1976. . 77
28 Characteristics of waters which, if combined, will invariably
produce major changes in water quality if treated for
aquatic weed control 78
29 Herbicides, fungicides, defoliants. From Anon., 1973 81
30 Effects of various concentrations of herbicides on small blue-
gills and from four species of fish. From Hiltibran, 1967. . . 85
31 Residues of 2,4-D in fish from Loxahatchee National Wildlife
Refuge. From Schultz and Whitney, 1974 87
32 Diquat: toxicity to aquatic organisms. From Folmar, 1977. ... 89
33 Endothall: toxicity to aquatic organisms. From FoLnar, 1977. . . 91
34 Lethal and tolerant concentrations of 12 toxic compounds tested
on 13 species of Protozoa. From Cairns, 1974 93
35 Dalapon: toxicity to aquatic organisms. From Folmar, 1977. ... 96
36 Toxicity of copper to marine and aquatic life. From Anon., 1973. 99
XIII
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INTRODUCTION
Excessive growth of aquatic plants often seriously interferes with many
domestic, agricultural, industrial, and recreational water uses. For this
reason, a number of diverse management techniques have been developed that
can reduce the growth of aquatic plants (Dunst et al. 1974). In the past,
these techniques were often used without knowledge of their long-term impact
on the aquatic environment, and a technique was considered good if it
controlled aquatic plants. Water, however, is becoming an increasingly
valuable resource throughout the United States and many user groups are now
expressing concern about the impact of aquatic plant management techniques on
the aquatic environment. Because of this concern, there have been increased
demands for development of effective, yet environmentally safe aquatic plant
management programs.
Formulation of aquatic plant management programs for different aquatic
systems is extremely difficult. It is often very difficult to obtain an
accurate assessment of the extent of an aquatic plant problem because of the
wide range of public and private waters affected and the numerous overlapping
governmental jurisdictions responsible for managing aquatic plant problems.
If aquatic plant management programs are not to become mere cosmetic treat-
ment programs, information is needed on the factor or factors that control
aquatic plant growth. This information, however, is often unavailable.
Finally, information is needed on the Impact of various aquatic plant manage-
ment techniques on the aquatic environment. Study of the aquatic environment
is not new (Forbes 1887; Apstein 1896; Kofoid 1903; and others), and exten-
sive literature exists concerning the general limnology of different aquatic
systems, the limnological roles of aquatic plants, and the impact of differ-
ent management techniques on the aquatic environment (Gessner 1955, 1959;
Hutchinson 1957, 1967, 1975; Sculthorpe 1967; Dunst et al. 1974; Wetzel 1975;
and others). However, very few patterns in the behavior of aquatic systems
have been quantitatively documented. Thus it is very difficult to predict
the short-term and long-term impact of aquatic plant management programs.
In this review paper, we attempt to provide a general assessment of the
aquatic plant problem in the United States. We also present information on
what is currently known about the impact of aquatic plants and management
techniques on water quality. Due to the complexity of aquatic systems, we
also discuss the overall impact of aquatic plants and management techniques
on the aquatic environment. We include this information because other
components of the aquatic environment which are affected either directly
or indirectly by aquatic plants can also affect water quality and because
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without this information it is impossible to provide a balanced perspective
of the total impact of aquatic plants and techniques for their .management on
the water resources of the United States. Finally, after reviewing the
information currently available, we identify what we consider to be major
research needs if effective, environmentally safe aquatic plant management
programs are to be developed.
PROBLEM AQUATIC PLANTS
Algae
Algae are found in nearly all waters located throughout the United
States. In 1972, the United States Environmental Protection Agency (EPA)
initiated the National Eutrophication Survey to investigate the nationwide
threat of accelerated cultural eutrophication to freshwater lakes and
reservoirs. This survey collected physical, chemical, and biological data on
more than 800 lakes and reservoirs throughout the contiguous United States.
Figure 1 shows the relationship between algal biomass as measured by
chlorophyll a_ concentrations and total phosphorus concentrations for the
sampled lakes and reservoirs. If chlorophyll ji values above 10 mg/m and
total phosphorus values above 20 mg/m are used to demarcate eutrophic
lakes and reservoirs with algal problems, algal problems occur in many of the
nation's lakes and reservoirs (Figure 1).
Research on natural lakes and reservoirs (Edmondson 1961; Sakamoto 1966;
Vollenweider 1968; Dillon and Rigler 1974a; Schindler 1975; Jones and
Bachmann 1976; Canfield 1979) has shown a strong correlation between algal
biomass as measured by chlorophyll a_ concentrations and total phosphorus
concentrations (Figure 2). This research strongly suggests that phosphorus
is the element most likely limiting algal biomass. Whole-lake experiments by
Schindler (1975) have further shown that reduction in phosphorus input to
lakes will significantly reduce algal biomass. For this reason, there has
been a nationwide effort to reduce phosphorus inputs to lakes and reservoirs.
Nutrient reduction programs have been successful on a number of lakes
(Michalski and Conroy 1973; Schindler 1975), with Lake Washington (Edmondson
1961, 1966, 1969, 1970, 1972a, 1972b) being a clasic example of the benefits
of controlling nutrient inputs. Other lakes with a long history of high
phosphorus inputs, however, have recovered very slowly with reductions in
phorphorus inputs (Ahlgren 1972; Bjork 1972; Larsen et al. 1975). While
Shapiro (1979), Canfield (1979), and others suggest that factors other than
phosphorus, such as zooplankton and sediments, affect algal biomass, reduc-
tion of plant nutrient inputs, especially phosphorus, is an important first
step in management of algal problems in the United States. Additional re-
search on lake biology, internal recycling on plant nutrients, and algal pop-
ulation dynamics, however, will be required in order to determine how to man-
age algal problems in waters where it is impossible to reduce nutrient input.
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100
E
o>
E
at 10
CL
O
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Table 1.
Classification of English lakes on the basis of their physical characters and rooted vegetation. From Pearsall, 1 929.
Lake
Wastwater
Ennerdale
Buttermere
Crummock
Hawes Water
Derwentwater
Bassenthwaite
Coniston
Ullswater
Windermere
Esthwaite
Drainage
System
per cent
cultivable
5.2
5.4
6.0
8.0
7.7
10.0
29.4
21.8
16.6
29.4
45.4
Lake
Shore*
per cent
Rocky
73
66
50
47
25
33
29
27
28
28
12
Relative
transparency
of water
9.0
8.3
8.0
8.0
5.8
5.5
2.2
5.4
5.4
5.5
3.1
Percent
Isoetes
49
35
40
48
5
31
42
34
34
9
2
of submerged
Nitella
36
48
40
26
71
42
3
9
15
40
26
vegetation
Potamogenton^
Vi
Vi
1
2
5
6
3
30
35
38
56
* To a depth of 30 ft.
t Including Naias and Elodea in small quantities.
Native Aquatic Macrophytes
Native aquatic macrophytes are also found in nearly all waters in the
United States and may produce nuisance growths which require control. To
the best of our knowledge, there has been no study similar to the National
Eutrophication Survey which details the extent of native aquatic macrophyte
problems. Reports, however, of either localized or extensive native macro-
phyte problems caused by emergent (Typha sp., Polygonum sp., Eleocharis sp.,
and others), submergent (Potomogeton sp., Utricularia sp., Elodea sp.,
Cabomba sp., and others), floating-leafed (Nelumbo lutea, Nuphar sp. and
others), and floating (Lemna sp., Spriodela sp., Wolffia sp., and Wolffiella
sp.) aquatic plants can be found.
Unlike for algae, there is a general lack of information on the factor
or factors that limit growth of aquatic macrophytes. Early works by Pearsall
(1921, 1922) have suggested that edaphic factors, particularly the coarse-
ness of the bottom, determine distribution of different types of aquatic
plants in English lakes (Table 1). Pearsall also noted that finer bottom
silts generally were associated with higher nutrient concentrations and
generally support abundant macrophyte growth. Although these data would seem
to suggest that plant nutrients control growth and distribution of aquatic
macrophytes and that increased nutrient inputs resulting from human activ-
ities might be responsible for increased native macrophyte growth, there
are very little data to support this hypothesis. There also have been no
studies to determine if reductions in nutrient inputs will cause a corre-
sponding reduction in growth of aquatic macrophytes. Additional research
is urgently needed in these areas.
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1000
100
o
I
Q.
o
CE
o
_l
I
u
10
0.1
LOG CHL_a_ -1 09 + 1 46 LOG TOTAL P
r 095
10 100
MEASURED TOTAL PHOSPHORUS MG/M3
1000
Figure 2. Relationship between summer levels of chlorophyll a_ and measured
total phosphorus concentration for 143 lakes. From Jones and
Bachmann, 1976.
Non-native Aquatic Macrophytes
Water hyacinth (Eichhornia crassipes), and hydrilla (Hydrilla
verticillata), Eurasian watermilfoil (Myriophyllum spicatum) and alligator-
weed (Alternanthera philoxeriodes) are all introduced plant species that have
or are currently causing serious aquatic weed infestations in the United
States. As with native aquatic macrophytes, there has been no national study
to document the extent of the non-native aquatic macrophyte problem. Since
its introduction into Lousisiana in 1908, water hyacinth has become distrib-
uted throughout the South and parts of California (Figure 3). Hydrilla,
which is currently a major aquatic plant problem, has spread since its intro-
duction in 1960 (Figure 4) throughout Florida (Figure 5) where it now covers
over 250,000 ha of Florida's fresh water. Hydrilla is now present in Ala-
bama, Mississippi, Georgia, South Carolina, Louisiana, Texas, California, and
Iowa (Figure 6) and may pose a serious threat to other states. Eurasian
watermilfoil is currently the most widespread aquatic macrophyte that causes
problems in the United States (Figure 7). The area of most rapid coloniza-
tion seems to be in the lower Great Lakes-St. Lawrence River area, but severe
infestations have been found in reservoirs of the Tennessee Valley Authority
(TVA) and some of the lakes and reservoirs in Washington.
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Figure 3. Distribution of water hyacinths in the United States, 1978
Figure 4. Known distribution of hydrilla
in Florida in 1960. The total
infestation in both Crystal River
and the Miami River was approxi-
mately 10 ha.
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Figure 5. Known distribution of hydrilla
in 1978. Larger dots represent
dense infestations totaling some
40,000 ha, and smaller dots indicate
hydrilla common in the flora of an
additional 200,000 ha of Florida's
fresh water.
Figure 6. Distribution of hydrilla in the United States, 19;
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oo
Figure 7. Distribution of Eurasian watermilfoil in the United States, 1978.
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There are practically no data on factors that might limit growth of
these non-native aquatic macrophytes in the United States. Alligator-weed,
which once caused serious problems, is being controlled by the alligator-
weed flea beetle (Agasicles hygrophila Selman and Vogt) which was introduced
into the United States. Water hyacinths are currently being managed at
tolerable levels through the use of the herbicide 2,4-D, but in backwater
areas and little-used canals where it is often ijnpossible or too costly to
spray, water hyacinths tend to cover the surface. Herbicides are also being
used in an attempt to control hydrilla and watermilfoil, but with difficulty
because the plants are submersed. In the case of hydrilla, the plant can
reproduce vegetatively (Haller 1976) and produce underground tubers which are
unaffected by most control methods (Miller et al. 1976). It has been spec-
ulated that nutrient enrichment of the nation's waters due to human activ-
ities might be contributing to the rapid spread of aquatic weeds. While
nutrients no doubt affect growth of these plants, cultural eutrophication
probably is not the major cause of non-native aquatic macrophyte infest-
ations. What we are probably witnessing is the invasion of new territory by
competitively superior plants. For example, Van et al. (1976) showed that
hydrilla had a lower light requirement for photosynthesis than some native
plants. They suggested this provided hydrilla with a competitive advantage
in the uptake of carbon dioxide, a necessary plant nutrient. Further
research will be needed to determine if there are natural factors that can
be managed to limit the growth of non-native aquatic macrophytes.
Impact of Aquatic Plants on the Aquatic Environment
Impact of Aquatic Plants on Physical Characteristics
Optical Properties. A major water quality parameter that is readily
evaluated by the general public and significantly affected by aquatic plants
is water clarity. The development of excess plankton algae can signifi-
cantly reduce water clarity. Recent studies on natural lakes (Edmondson
1972a,b; Bachmann and Jones 1974; Dillon and Rigler 1974a; Oglesby and
Schaffner 1975; Jones and Bachmann 1978) have shown a hyperbolic relationship
between algal biomass as estimated by chlorophyll ti concentrations and water
clarity as measured with Secchi disc (Figure 8). Using data from a large
number of lakes (Figure 9), Jones and Bachmann (1978) showed that water
clarity as measured with a Secchi disc (SD) could be predicted from
chlorophyll a_ (Chi &) concentrations by the following equation:
Log SD = 0.807 - 0.549 Log (Chi a + 0.03) (1)
These relationships (Figures 8 and 9) clearly show that significant improve-
ments in water clarity cannot be expected until chlorophyll a_ concentrations
are reduced below 10 mg/m . With these relationships, water resource man-
agers can predict not only water clarity, but the general public's response
to either increases or reductions in algal concentration.
-------
40
35
30
co
a:
25
20
00
0 15
X
u
CJ
UJ
CO
10
6r
1 -
-- --
10 20 30 40 50
0 50 100 150 200 250 300 350
CHLOROPHYLL a MG. M3
Figure 8. Relationship between mean Secchi disc transparencies for
July and August and the mean July-August chlorophyll _a
concentrations for 16 lakes. From Bachmann and Jones,
1974.
10
-------
Not all aquatic plants, however, reduce water clarity. A number of
researchers have noted that dense growths of aquatic macrophytes seem to
inhibit development of phytoplankton populations and are generally associated
with clear water (Kofoid 1903; Schreiter 1928; Hasler and Jones 1949; Hogetsu
et al. 1960; Stangenberg 1968; Goulder 1969). If aquatic macrophytes ac-
tually inhibit phytoplankon development, excellent water clarity associated
with dense macrophyte growths probably results from reduction in algal
numbers. Irwin (1945) also suggested that aquatic macrophytes enhance
reduction of clay turbidities by reducing charges on the clay particles and
thus improving water clarity. Aquatic macrophytes might also contribute to
improved water clarity by reducing resuspension of bottom sediments as Pond
(1903) and Klugh (1926) suggested. There are, however, no data relating
aquatic macrophyte biomass to water clarity in lakes, thus it is very
difficult to predict what level, if any, that aquatic macrophyte biomass
will contribute to a significant improvement or reduction in water clarity.
Additional research is especially needed to ascertain quantitatively the
relationship between aquatic macrophyte biomass and algal biomass as avail-
able data clearly show algae sigificantly affect water clarity.
Although water clarity is readily evaluated as an index of water quality
and clear water has a greater aesthetic value than turbid water, the general
public also uses water color as an index of water quality. Phytoplankton are
often responsible for imparting dull green, blue-green, or yellowish-brown
colorations to water (Naumann 1922; Hutchinson 1957; Wetzel 1975). Occasion-
ally, populations of the genera Euglena, Haematococcus, or Glenodium may
impart a blood-red coloration to the water (Klausner 1908; Huber-Pestalozzi
1936; Wetzel 1975). To the best of our knowledge, relationships between
algal biomass and water color are based strictly on observational data.
There seem to be very little quantitative data that can be used to determine
the algal biomass at which color becomes noticeable. If quant- itative data
could be obtained, it might be possible to develop an empirical relationship
between algae and water color that would be of use to water resource managers
in predicting response of the general public to increases or reductions in
algae. This type of information is important to water resource managers
because, regardless of what the physical, chemical, or biological limnolog-
ical data may show, the general public must perceive a change or else often
will assume that no change has occurred.
Aquatic plants also can significantly affect other optical properties of
the aquatic environment. The relationship between Secchi disc depth and
algal biomass as measured by chlorophyll a. concentrations clearly shows that
an increase in algal biomass reduces depth of light penetration (Figure 9).
Studies by Beeton (1958), Stepanek (1959), and Tyler (1968) have shown that
the Secchi disc depth represents from 1 to 15% of the surface light trans-
mitted. Studies by Costing (1933), Gessner (1955), Szumiec (1961), and
Mailer and Sutton (1975)* have shown that the amount of light entering various
depths of the aquatic environment can be significantly affected by dense
growths of aquatic plants (Figure 10). Ultsch (1973) observed that dense
mats of water hyacinths can effectively block all light from entering the
aquatic environment.
11
-------
100
LOG SO- 0.807-0.549LOG(CHL a* 0.03)
to
IT
LJ
(-
UJ
o.
UJ
o
X
o
o
UJ
tO
10
* IOWA LAKES
OTHER LAKES
... ..I ,
0.03
Figure 9.
O.I
I 10 100
CHLOROPHYLL fl » 0.03 MG/M3
1000
Double logarithmic plot of Secchi disk depths against average
chlorophyll a_ concentrations. Data for non-Iowa lakes were taken
from the literature (Bachmann and Jones, 1976, Dillon and Rigle'r,
1975, Oglesby and Schaffner, 1975) and from reports of lake self-
help projects in Ontario and Michigan. The addition of the con-
stant 0.03 to the chlorophyll values prevents the calculated
Secchi disk values from approaching infinity as the chlorophyll
levels approach zero. From Jones and Bachmann, 1978.
12
-------
Open water
Vallisneria
A Hydrilla
2.00
100
Percent of Surface Radiation
Figure 10. Light penetration through water and plant communities in
Rodman Reservoir. Light measurements in the hydrilla com-
nunity were made through small (20 cm ) openings in the
canopy. From Haller and Sutton, 1975.
Aquatic plants can directly influence the quantity of light entering the
aquatic environment and also indirectly influence its quality. Studies
(Wetzel 1969a; 1969b; Fogg 1971; Hellebust 1974; and others) have strongly
suggested that dissolved organic compounds are released by aquatic plants.
James and Birge (1938) showed that increasing concentrations of dissolved
organic compounds as measured by water color not only reduce the depth to
which light is transmitted, but shift absorption selectivity (Table 2).
Though most of the dissolved organic compounds studied by James and Birge
(1938) were probably derived plant material produced within the watershed,
their data strongly suggest that aquatic plants can influence not only the
quantity, but also the quality of light entering the aquatic environment.
Because the quantity and quality of light entering water influences photo-
synthesis, and hence production by various aquatic plants, research should be
conducted to determine how quantitative and qualitative changes affect
distribution, abundance, and production of aquatic plants.
13
-------
Table 2.
Percentile absorption of light of different wavelengths by one meter
of lake water, settled of participate matter, of several Wisconsin lakes of
progressively greater concentrations of organic color3 From Wetzel, 1975.
Wavelength
(nm)
800
780
760
740
720
700
685
668
648
630
612.5
597
584
568.5
546
525
504
473
448
435.9
407.8
365
Color Scale
(Pt units)
Distilled
Water
88.9
90.2
91.4
88.5
64.5
45.0
38.0
33.0
28.0
25.0
22.4
17.8
9.8
6.0
4.0
3.0
1.1
1.5
1.7
1.7
2.1
3.6
0
Crystal
Lake
89.9
91.3
93.5
89.3
67.6
50.4
45.2
40.3
37.0
34.4
32.1
27.5
22.0
19.3
19.2
19.8
20.7
21.7
23.8
24.4
28.1
40.0
0
Lake
Mendota
90.5
91.9
92.6
91.5
71.0
49.7
42.2
36.8
31.9
28.9
26.3
22.5
17.6
14.0
13.5
14.1
15.2
21.7
27.8
31.0
44.3
80.0
6
Alelaide
Lake
92.4
93.5
94.5
92.7
78.0
66.3
65.7
65.0
64.5
65.8
66.8
67.0
67.1
67.6
70.9
74.5
81.0
88.6
92.2
95.2
99.0
28
Mary
Lake
91.7
93.0
94.8
93.0
78.0
70.7
71.7
72.3
75.2
77.8
80.3
83.2
85.7
88.5
91.6
94.8
974
99.4
101
Helmet
Lake
93.2
94.5
96.0
96.2
86.9
82.5
86.6
88.0
91.2
94.0
96.0
97.6
98.2
98.6
99.3
264
"Selected data from James and Birge, 1938.
Thermal Properties. Absorption of solar energy by water is influenced
by a mjnber of physical, chemical, and, under certain conditions, biological
characteristics of the water (Wetzel 1975). Because water temperatures may
influence various physical, chemical, and biological characteristics of the
aquatic environment, it is important to understand the effect of aquatic
plants on the thermal properties of different aquatic systems. Costing (1933)
noted that water temperatures were frequently lower in beds of floating or
floating-leafed aquatic macrophytes than in open water areas. Dvorak (1970)
also found consistently lower temperatures within stands of the emergent
Glyceria than in open water. His data showed that water temperatures aver-
aged about 17 C within the Glyceria stands but about 21 C in open water.
Hotchkiss (1941) observed that surface water over submersed aquatic beds
heated very rapidly to relatively high temperatures. He suggested that sur-
face heating resulted because aquatic macrophytes restricted water movements
and thus prevented mixing of warm water with cooler water.
14
-------
Although there are data that indicate aquatic plants can affect thermal
properties of aquatic environments, there are no data that indicate how
important their role is in aquatic systems of differing sizes or depths.
It would be interesting to know how aquatic plants affect the total heat
budget of different aquatic systems. It also has been speculated that
aquatic plants, by affecting temperatures, can affect the behavior of
aquatic organisms such as fish. There are, however, very little data on
which to judge just how important temperature modifications by aquatic plants
are to other aquatic organisms. Because temperature is an important water
quality parameter, studies are needed to determine quantitatively how
different levels of aquatic plants affect the thermal properties of different
aquatic systems.
Water Movements. The movement of water is extremely important in the
aquatic environment as it can transport plant nutrients and organic matter,
add oxygen to water through surface aeration, affect silt deposition and
erosion sites, and either directly or indirectly affect various water quality
parameters. Direct measurements of the impact of aquatic plants on water
movements have been few. Studies have shown that aquatic plants in irri-
gation and drainage canals can impede the flow of water (Stephens et al.
1957; Timmons 1967). Stephens et al. (1957) showed that water hyacinths
reduced flow by 40%, while submersed aquatic macrophytes reduced flow by 97%
in extreme cases. Observations of lakes by various researchers (Costing
1933; Hotchkiss 1941; Seabrook 1962; Dvorak 1970; Unni 1972) have suggested
that aquatic plants, especially aquatic macrophytes, can significantly reduce
water movements. The lack of data, however, makes it very difficult to
determine the magnitude of water movement reductions even though some re-
searchers have suggested plants can reduce water movements to the ppint of
stagnation.
Water flow and circulation can affect nutrient dynamics, the transport
and final deposition of sediments and associated toxicants, and a host of
other parameters. Therefore, detailed studies are needed to quantify the
impact of various aquatic plants on water movements. At the present time,
there are no quantitative data to indicate at what levels various types of
aquatic plants affect water movements or how significant these effects might
be to the functioning of aquatic systems. Quantitative data should be col-
lected from different aquatic systems to determine if these are general be-
havior patterns which would permit the prediction of the impact of different
levels of aquatic plants on water movements.
Water Balance. The extent and rate of evaporative losses from the
aquatic environment are affected by many factors. It has long been recogni-
zed, however, that transportation by floating, floating-leafed, and emergent
aquatic macrophytes (Gessner 1959; Wetzel 1975) can greatly increase evapor-
ative losses. For example, studies by Penfound and Earle (1948), Timmer and
Weldon (1967), and Holm et al. (1969) have shown that evaporative water loss
due to water hyacinths was 3.7 to 7 times greater than evaporative loss from
open water (Table 3).
15
-------
Table 3.
Transpiration for water hyacinth in relation to evaporation. From Penfound
and Earle, 1948.
Total
Date Conditions transpiration. Evaporation, Transpiration,
Milliliters Milliliters Evaporation
June 17-June 19
June 20-June 28
June 20-July 2
JulyS-July 9
July 10-July 20
Total
Clear
Cloudy, rain
Clear
Cloudy, rain
Clear, except
13th, 14th
Clear, 13
days; Rain,
21 days
8,650
27,400
11,700
26,900
31,700
106,350
1,900
10,500
2,800
13,200
4,800
33,200
4.5
2.6
4.2
2.0
6.6
3.2
Annual water loss from irrigation canals due to evapotranspiration by
aquatic plants may be considerable (Timmons 1960). Hotchkiss (1941) sug-
gested that transpiration by aquatic plants was especially important in
drying of shallow aquatic systems, thus enhancing establishment of terres-
trial plants.
Although it is recognized that transpiration by aquatic plants in-
creases the rate of water loss, there is very little information on loss
rates for different types of aquatic plants. Further, to the best of our
knowledge, water loss rates have not been determined for different quantities
of aquatic plants. Research is needed in these areas if water resource
managers are to decide rationally what types of aquatic plants and how many
aquatic plants should be allowed to grow in various aquatic systems. This
information is especially needed in regions of the United States where
irrigation is important or surface water supplies are limited.
Basin Morphometry. Basin morphometry may have important effects on
nearly all major physical, chemical, and biological characteristics of the
aquatic environment (Thienemann 1927; Hutchinson 1938; Rawson 1952, 1955,
1956; Patalas 1961; Hayes and Anthony 1964; Sakamoto 1966). For example,
studies by Rawson (1955) have shown that the average standing crop of plank-
ton (Figure 11), the average weight of bottom fauna (Figure 12), and the
average long-term commercial fish production (Figure 13) increase as mean
depth decreases in lakes. Aquatic plants may have a major long-term indirect
impact on aquatic systems by filling in of basins through precipitation of
calcium carbonate (Welch 1952; Otsuki and Wetzel 1974; Wetzel 1975), entrap-
ment of inflowing sediments (Hotchkiss 1941), and accumulation of their
remains (Wilson 1945; Wetzel 1975). The role aquatic plants play in the
modification of basin morphometry, however, is often overlooked by water
resource managers when they attempt to evaluate 'the impact of aquatic plants
on the aquatic environment.
16
-------
18O
170
160
150
140
130
~ 120
X
> 110
o
Z 100
0
\
1 90
°- BO
t
LU
2 70
U_
O
x 60
(J
iso
rx 40
Q
30
20
10
O
1 1 1 1 1 1 1 1 ' 1 1 1 r~~-T
-
" .
-
-
-
.
* Mendoto
-
_ 3765 0 ^ 8Q
Losi Mountain " H5337
~
-
-
-
-
-
Trout
* Waskesiu
- \
\
\ La Ronge
\ Nesslin
\
\ . Paul
\
Hunipr Sav \« Athabaska
, \ m Minnewanka
* " Amisk\* Cullus
^v Okanagan
* Pyramid \ Great Slave
- **^_ Walerton
Mahgne ^"-^^^
^^^^^^^^ Christie Bav
" McLeod Bav
i i 1 1 1 1 1 1 1 1 1 »«»»il
MEAN DEPTH, METRES (d)
Figure 11. Mean depth and the average standing crop of plankton in twenty
lakes. From Rawson, 1955.
17
-------
251 ; 1 1 1 1 r r- 1 1 1 r
20
o
t
§10
5
tr
f =
69 2.
( Slavel
Nlp'9°n Ontario
Superior
100 200 300 400 50°
MEAN DEPTH OF LAKES. IN FEET (d)
Figure 13. Mean depth and the long-term average commercial fish production
in twelve lakes. From Rawson, 1955.
18
-------
Water resource managers often consider aquatic plants to be detrimental
to an aquatic system because they contribute to the infilling of the basin
and eventually to destruction of the system. Rawson's work, however, strong-
ly suggests that in some aquatic systems, reduction in mean depth resulting
from the presence of aquatic plants may have beneficial results, such as in-
creasing fish production. The infilling of an aquatic basin, however, often
does not occur uniformly across the whole basin. Pond (1903) and KLugh
(1926) suggested that establishment of rooted aquatic macrophytes retards
erosion of shallow water sediments to deeper waters. Wilson (1938),
Hotchkiss (1941),. and Welch (1952) suggested that stablization of bottom
sediments enhanced the buildup of littoral zones. The importance of littoral
zones as sites of nutrient recycling and as nursery areas for various organ-
isms has long been recognized. The special significance of littoral develop-
ment has been discussed by Strtftn (1928), Alsterberg (1930), Strata (1930),
and Fee (1979).
Although aquatic plants may contribute to the eventual destruction of
the basin, it should also be recognized that in some cases aquatic plants are
responsible for the formation and preservation of aquatic systems. Murray
(1910) suggested that aquatic plants, by blocking drainage channels, were
responsible for formation of many aquatic systems in low-lying regions in
the tropics and arctic. Penfound and Earle (1948) and Welch (1952) noted
that organic dams formed by aquatic plants could significantly increase water
elevations in existing aquatic systems. Parker and Cooke (1944) further
suggested that aquatic plants, by blocking drainage channels, were respons-
ible for preservation of Lake Okeechobee, Florida. Because aquatic plants
can reduce the power of waves (Pond 1903; Costing 1933; Welch 1952), they
can reduce shoreline erosion which subsequently reduces accumulation of sedi-
ment in deeper waters.
The impact of aquatic plants on basin morphometry and subsequent direct
and indirect effects on physical, chemical, and biological characteristics of
the aquatic environment are probably not considered by water resource
managers, because many of the changes in basin morphometry occur over long
periods of time. Long-term management of water resources, however, will re-
quire that the impact of aquatic plants on basin morphometry be considered.
At the present time, it is very difficult for water resource managers to
predict the impact of changes on basin morphometry. Except for Rawson's work
on the importance of lake mean depth, there are no general quantitative
relationships that can be used to predict the impact of changes in basin
morphometry on physical, chemical, or biological characteristics of the
aquatic environment. Even Rawson's 1955 work is based on data from only a
few lakes. Additional research will be needed to verify Rawson's relation-
ships. Research should also be conducted to determine the relationships of
volume-depth distributions to characteristics of the aquatic environment.
Finally, research is needed to determine the rate at which plant remains
accumulate and if accumulation rates vary for different types and abundances
of aquatic plants. Without this information, water resource managers will be
unable to predict the rate at which aquatic plants modify basin morphometry.
19
-------
Impact of Aquatic Plants on Chemical Characteristics
Oxygen. Dissolved oxygen, which is essential to all aerobic aquatic
organisms and strongly affects the solubility of various chemicals in the
aquatic environment (Hutchinson 1957, 1967; Wetzel 1975), is an important
water quality parameter. Because aquatic plants, through photosynthesis,
respiration, and decomposition affect oxygen levels, there have been a large
number of studies (Purdy 1916; Thienemann 1928; Rudolfs and Huekelekian 1931;
Juday and Birge 1932; Olson 1932; Tomlinson 1935; Eberly 1959, 1963, 1964;
Wetzel 1966a, 1966b; and many others) which have documented how aquatic
plants affect oxygen content of aquatic systems. Purdy (1916) showed that
aquatic plants, through photosynthesis, caused oxygen concentrations to
increase during the day, but through respiration caused nocturnal reductions.
This diel fluctuation in oxygen concentration (Figure 14) has also been
described for other systems (Olson 1932; Welch 1952; Hutchinson 1957; Wetzel
1975) and seems to be a general phenomenon of aquatic systems.
Verduin (1956) summarized the literature on aquatic plant primary pro-
duction in lakes and suggested that aquatic plants generally contribute
42-57 Ib of oxygen per acre per day. This addition of oxygen to the aquatic
environment has been shown to significantly affect distribution and percent-
age saturation of oxygen in lakes. Woodbury (1941) reported that dense algal
blooms could produce supersaturation in open water of lakes. Juday and Birge
(1932), Eberly (1959, 1963, 1964), and Wetzel (1966a, 1966b) have shown that
algal growth in the region of the thermocline in some lakes can cause in-
creases in oxygen concentration and may result in supersaturation at the
thermocline. Wetzel (1975) has shown that vertical stratification of oxygen
can occur in luxuriant stands of submersed macrophytes (Figure 14). In-
creases in oxygen levels were found near the surface of the macrophyte beds.
Although aquatic plants can increase the oxygen content of water and
generally produce an excess of oxygen through photosynthesis, plants can also
decrease oxygen concentration both directly and indirectly. Plant respira-
tion does not normally reduce oxygen to amounts critical for survival of
aquatic animals, but intense respiration coupled with high water temperature
may stress or kill animals (Olson 1932; Prescott 1939). The major mechan-
ism by which aquatic plants contribute to reduction of oxygen in the aquatic
environment is through decomposition (Tomlinson 1935; Sears 1936; Hutchinson
1936; Moore 1942; Thomas 1960; Ruttner 1963; Wetzel 1975). In highly produc-
tive aquatic systems, plant decomposition can cause anoxia in the hypolim-
nion. Even in unstratified lakes, oxygen concentration may be reduced and
fish killed after the death of large quantities of aquatic plants. Aquatic
plants, however, do not necessarily have to die to reduce oxygen levels.
Floating, floating-leafed and submergent aquatic macrophytes, by reducing
wave action and water circulation, can prevent physical reaeration of the
aquatic environment. Lynch et al. (1947), Ultsch (1973), and Wetzel (1975)
have provided data that indicate that oxygen levels under mats of water hya-
cinth, alligator-weed, and Elodea can be reduced to near anoxia (Figure 14;
Table 4).
20
-------
Table 4.
Comparisons of pH, temperature, dissolved 02 and dissolved C02 in hyacinth and "open water" areas.
March October is the growing season of Eichhornia in Gainesville, Florida. From Ultsch, 1973.
Top Temp. (°C)
Bottom Temp. (°C)
Top pH
Bottom pH
Top 02 (ppm)
Bottom 02 (ppm)
Top C02 (ppm)
Bottom C02 (ppm)
Yearly
21.5
19.0
5.4
5.3
4.2
1.2
31
51
Hyacinths
Nov.-
Feb.
13.5
12.0
5.7
5.6
5.6
2.5
16
26
March-
Oct.
25.5
23.0
5.2
5.1
3.5
0.6
39
63
Yearly
21.5
20.0
5.6
5.5
6.4
4.5
13
29
"Open" water
Nov.-
Feb.
13.0
13.0
5.9
5.8
8.1
7.2
7
8
March-
Oct.
25.5
23.5
5.5
5.4
5.6
3.2
16
40
10
_ 8
d
? 7
LITTORAL
o.o
0 4
X
Q_
UJ
Q
0.8
1.2
1.6 -
2.0 -
0800
1200
1600 . 2000
HOURS
2400
O400
4 6
mg Oj T'
10
Figure 14. Upper: Changes in dissolved oxygen in the littoral and open
water areas over a diel period in eutrophic Winona Lake, Indiana,
9 August 1922. (From data of Scott, 1924). Lower: Vertical
stratification of oxygen within the littoral zone of Parvin
Lake, Colorado, 9 July 1955, in a luxuriant stand of the sub-
mersed macrophyte Elodea. (Generated from data of Buscerai,
1958). From Wetzel, 1975.
21
-------
Despite the large amount of data now available on the relationship of
aquatic plants to oxygen concentration in aquatic systems, there is a general
lack of information on aquatic plant biomass and observed oxygen changes.
For example, it has long been recognized that oxygen depletion in the hypo-
limnion of productive lakes occurs. There is, however, no reliable method to
predict the quantity of aquatic plants needed by aquatic systems to maintain
satisfactory oxygen regimens. Nor is there a reliable method to predict
oxygen depletion in unstratified lakes, although Barica (1975) has provided
data which might make it possible to predict summer fish kills (Figure 15).
Water resource managers need quantitative information that can be used to
predict the relationship of aquatic plant biomass to oxygen if oxygen deple-
tions are to be prevented. Research should be conducted on a large number
of lakes, such as that of Barica (1975), in order to determine general pat-
terns. This type of information could provide water resource managers with
predictive equations.
350
300-
x 250
3.
2
ca 200
.c
Q.
O
CJ
X
CO
150-
100-
50-
SUMMERKILL
5 RISK
* =SUMMERKILL LAKES
0 0.2 0.4 0.6 0.8 1.0 1.2 1.4 1.6 1.8 2.0 m
Min. Secchi disc, transparency(s)
Figure 15. Relationship between the maximum chlorophyll a_ and minimum
Secchi disc transparencies in various prairie ponds (Erickson-
Elphinstone district of southwestern Manitoba). From Barica,
1975.
22
-------
Carbon Dioxide, pH, Alkalinity. Most natural waters are buffered
principally by a carbon dioxide - bicarbonate system. As 002 dissolves in
water it forms carbonic acid: OC^ + HoO -> ^OOg (2). Because carbonic
acid is a weak acid, it dissociates rapidly:
t^COg -» H* + HOOg- (3); HOOg- -> FT + COg (4).
After equilibrium is established the bicarbonate and carbonate ions dissociate:
HOOV + H20 » HgCOg + OH~ (5);
°°3 * H2° "*" HOOo'-vHODo" = OH" (6);
HgCOg * H20 = COg (7).
The total quantity of carbon dioxide that passes into natural water, however,
is not determined only by the pressure and solubility of carbon dioxide, but
by the formation of bicarbonates with alkaline earth metals. For example,
carbonic acid effectively solubilizes calcium carbonate, forming calcium bio-
carboate:
+ CaC03 -> Ca(H003)2 (8); which dissolves in water:
3)2 -> Ca++ + 2HCOo (9).
These reactions influence not only the carbon dioxide content of water but
also the pfl and alkalinity of natural waters.
Because aquatic plants utilize carbon dioxide during photosynthesis and
release it during respiration, they strongly influence carbon dioxide content
pH, and alkalinity of water. In well-buffered natural waters, the effect of
aquatic plants may be minimal, but in poorly buffered systems aquatic plants
may have a major impact. For example, Costing (1933) showed that aquatic
plants could deplete carbon dioxide in the upper water levels of Snail Lake
(Table 5) while plant respiration or decomposition maintained higher carbon
dioxide concentration at depth. Uptake and release of carbon dioxide can
also produce diel variations in carbon dioxide, pH, and alkalinity (Figure 16),
as noted by Unni (1972) and Van et al. (1976), and cause vertical variations.
Oosting (1933) observed sharp pH gradients within aquatic macrophyte beds.
At the surface of the beds he recorded a pfl of 8.9 while only 2.5 feet below
the leaves a significantly lower pH of 6.9 was found. Wetzel (1975) showed
that similar vertical stratifications (Figure 17) in pH could occur also in
the open water areas of hypereutrophic Wintergreen Lake.
Many aquatic organisms can withstand large variations in pH: thus,
changes caused by aquatic plants seldom kill organisms. However, some organ-
isms are extremely sensitive to pH changes and those caused by aquatic plants
could affect these organisms. Carbon dioxide concentration, pH, and alkalin-
ity, however, affect many of man's water uses, and it is extremely important
to understand how aquatic plants affect carbon dioxide concentration, pH, and
alkalinity. Although the chemical reactions are well understood and there is
ample information on changes aquatic plants may cause, water resource managers
have a difficult time predicting chages in carbon dioxide concentration, pH,
and alkalinity that might result from aquatic plant activity. Quantitative
data relating changes in carbon dioxide, pH, and alkalinity to aquatic plant
biomass in aquatic systems are lacking as most studies have been qualitative.
Before water resource managers can predict the effect of aquatic plant biomass
on carbon dioxide, pH, and alkalinity in different aquatic systems, data will
have to be obtained from a large number of aquatic systems.
23
-------
TableB.
Carbon dioxide: vertical variation 1 928.
From Costing, 1933.
Snail Lake (cc./L)
Aug. Aug. Aug. Aug.
4 11 18 25
6 in 25 0 0 0
3 ft
6 ft
9 ft
1 2 ft
15 ft. ...
1 8 ft. . . .
21 ft
22 ft . .
24 ft. . . .
26 ft. . . .
51
51
51
51
51
76
76
. . 1 00
o
o
o
o
1 0
1 5
1 8
20
2 5
o
o
o
0
0
51
1 00
1 80
2 30
o
o
o
o
o
3
5
3 0
3 3
30
25
20
10
o -
10
1600
2000
Day 1
2400
0400
Time
0800
1200 16OO
Day 2
Figure 16. Diurnal fluctuations in free CC^, HCO-, , On, and pH measured
in the surface 5 cm of water over a Hydrilla verticillata mat.
Data were collected on October 14 and 15, 1975, at Lake Killar-
ney, Florida. Figures in parentheses refer to the water tem-
perature (C) at the time of sampling. From Van et al., 1976.
24
-------
3 -
5 -
JUN JUL AUG SEP OCT NOV DEC JAN PEB MAR APR MAY JUN
Figure 17- Depth-time diagram of isopleths of pH in hypereutrophic
Wintergreen Lake, Michigan, 1971-1972. Opaque area= ice
cover to scale. From Wetzel, 1975.
Water Salinity. Salinity of freshwaters is usually dominated by the
cations of magnesium, sodium, potassium, and calcium and the anions of chlo-
ride, sulfate, and carbonate. Concentrations of magnesium, sodium, potass-
ium, and chloride are relatively conservative and undergo minor spatial and
temporal fluctuations due to aquatic plant utilization (Figure 18; Wetzel
1975). Surface water reductions in potassium (Figure 18) may, however, be
due to aquatic plant uptake and subsequent aquatic plant sedimentation
(Barrett 1957;Wetzel 1975). Plant decomposition has also been suspected of
occasionally releasing large amounts of magnesium to the aquatic environment
(Wetzel 1975).
Aquatic plants, however, significantly affect temporal and spatial dis-
tribution of calcium and carbonate concentrations. For example, calcium
(Figure 19) and carbonate concentrations in Wintergreen Lake, Michigan, under-
went marked seasonal fluctuations (Wetzel 1975) that are typical of many
hardwater lakes (Welch 1952; Hutchinson 1957; Wetzel 1975). Such fluctu-
ations result when aquatic plants utilize carbon dioxide during photosyn-
thesis. Uptake of carbon dioxide affects the chemical equilibrium of the
waters by: Ca"1"* + 2HC03~ * 002 + H20 + CaCOg. This process
results in precipitation of calcium carbonate (CaCOo). Otsuki and Wetzel
(1974) observed that aquatic plant photosynthesis, by inducing precipitation
of calcium carbonate, was responsible for decalcification of surface waters.
The loss of calcium also reduced the specific conductance of the surface
waters. Because some calcium carbonate can be resolubilized by carbon
dioxide in deeper waters, precipitation of calcium carbonate from surface
waters can cause an increase in concentration of calcium and carbonate ions
in deeper waters.
25
-------
IT 6
10
I 3
(a)
45
45
JAN FEB MAR APR MAY JUN JUL AUG SEP OCT NOV DEC
1972
(b)
50
60 n 6.0 60 60 5.5
6 5
6.5
35
JUN JUL AUG SEP OCT NOV DEC JAN FEB MAR APR MAY JUN
Figure 18. Depth-time distribution of isopleths of sodium concentrations
(mg Na+ I"1) of Lawrence Lake, 1972 (a), potassium concentrations
(mg K"1" 1~ ) of hypereutrophic Wintergreen Lake, Michigan, 1971-
1972 (b), chloride concentrations (mg Cl~ I"1) of eutrophic Little
Crooked Lake, Whitley County, Indiana, 1964 (c), and magnesium
concentrations (mg Mg I"1) of hardwater Lawrence Lake, Michigan,
1972 (d). Opaque areas=ice-cover to scale. From Wetzel, 1975.
26
-------
01
2
\20l
1 iii,
'4 12 12 14 1616 14 14 14
J 8
Q.
10
12
14
"I ~T
(c)
14 id 1816 14 14 15
16 16
FEB VIAR APR MAY JUN JUL AUG SEP O<" T NO" DEC
o
2
4
I 6
Q.
UJ
Q
g
10
12
' 26
28 26
-
24
26
26 ;
-
26
28 | 26 26 28
I 30
N
28 26
24 24
24 24 24"
(d)
26! , 28
JAN FEB MAR APR MAY JUN JUL AUG SEP OCT NOV PFT
1972
Figure 18. Continued.
27
-------
5
JUN JUL AUG SEP OCT NOV DEC JAN FEB MAR APR MAY JUN
Figure 19. Depth-time distribution of isopleths of calcium concentrations
(mg Ca"1"*" 1~1) of hypereutrophic Wintergreen Lake,
Kalamazoo County, Michigan, 1971 - 72. Opaque area = ice cover
to scale. From Wetzel, 1975.
Past research has documented the possible impacts of aquatic plants on
water salinity. This research has, however, been basically descriptive.
There is a lack of quantitative information on the relationship of aquatic
plant biomass to salinity. For example, there are no data to predict the
effect of aquatic plant biomass on the quantity of calcium carbonate precip-
itated. Future research should relate changes in water salinity to the
quantity of plants present.
Nitrogen. Aquatic plants, through utilization of nitrogen compounds
during growth, can affect temporal and spatial distribution of nitrogen. For
example, Stadelmann (1971) observed that uptake of ammonia and nitrate re-
duced the concentration of these compounds to very low levels in lake sur-
face water. Plant decomposition in the bottom water, however, caused an in-
crease in ammonia concentration (Figure 20). These changes are only a general
pattern, as seasonal changes in nitrogen can very greatly among aquatic
systems of different productivities. For a detailed discussion of general
seasonal patterns in the transition from nutrient poor to nutrient rich
aquatic systems, consult Hutchinson (1957) or Wetzel (1975).
28
-------
15
MAMJJ ASONDJ FM
1969 1970
Figure 20. Depth-time diagram of the concentrations of ammonia (mg
1~*), Rotsee, Switzerland, 1969-70. (Redrawn from Stadelmann,
1971). From Wetzel, 1975.
Blue-green algae, because of their ability to fix atmospheric nitrogen
(Burris et al. 1943; Home and Fogg 1970; Rusness and Burris 1970; Home et
al. 1972), significantly affect nitrogen budgets of aquatic systems and thus
affect their productivity. Wetzel (1975) noted that nitrogen fixation in
waters without blue-green algae is insignificant, but as blue-green algal
populations increase, nitrogen fixation becomes significant. If fixation of
nitrogen becomes great enough, algal blooms may occur in waters that were
otherwise nitrogen limited. Within a given water body, Rusness and Bums
(1970) noted diel variations in nitrogen fixation with fixation being great-
est at midday and lowest at night (Figure 21). Wetzel (1975), using data
from Home and Fogg (1970). showed that nitrogen fixation may vary with
depth and that the pattern of fixation varies among lakes (Figure 22).
Billaud (1968), Home and Fogg (1970), and Toetz (1973) noted that nitrogen
fixation rates decline abruptly with decreases in blue-green algal pop-
ulations and are greatly reduced during the winter in northern lakes.
Although research has documented the effects of aquatic plants on the
nitrogen cycle of the aquatic environment, there is still a lack of quanti-
tative data. Before water resource managers will be able to predict the
total impact of aquatic plants on the nitrogen cycle of aquatic systems,
relationships now known will have to be related to aquatic plant biomass.
Because nitrogen is often believed to be a limiting plant nutrient, add-
itional research is needed to determine if nitrogen fixation by blue-green
algae supplies a quantitatively important part of the nitrogen budget of
aquatic systems. This information is especially important to water resource
managers who are charged with reducing nutrient inputs to aquatic systems
in order to control aquatic plant growth. If blue-green algae can fix
sufficient atmospheric nitrogen to support this growth and eventually the
growth of other aquatic plants, reduction of. nitrogen inputs resulting from
human activities will have little impact on aquatic plant growth.
29
-------
IUJ 7.5
CO ^ ^ 5'°
uj UJ O
_j :ȣ 00
23 2.5
Xz 40
J*0 30
CO \
^Z 20
ii
c 10
UJ
Q_
CO
x
CO
100
80
60
40
-------
Phosphorus. Studies on natural lakes (Sakamoto 1966; Dillon and Rigler
1974b; Jones and Bachmann 1976; Canfield 1979; and many others) have shown a
strong correlation between total phosphorus concentrations and algal biomass
(Figures 1 and 2), thus suggesting phosphorus is the element limiting algal
biomass. In fact, phosphorus may be the element limiting all aquatic plant
growth. For this reason, many studies have examined phosphorus dynamics in
lakes.
Studies with radioactive phosphorus (Hutchinson and Bowen 1947, 1950;
Coffin et al. 1948; Hayes et al. 1952; Hayes and Phillips 1958) have shown
that aquatic macrophytes, attached algae, and phytoplankton rapidly assim-
ilate phosphorus added to lake water. These studies along with studies by
Solski (1962) and Nichols and Keeney (1973) have shown that upon death and
decay the plants release a large portion of the phosphorus to the water.
Recent studies (McRoy and Barsdate 1970; Boyd 1971; Reimold 1972; McRoy et
al. 1972; Lie 1978) have suggested that rooted aquatic macrophytes can also
assimilate phosphorus from the bottom sediments. This phosphorus can also
be released to overlying waters. Lie (1978) calculated that aquatic plants
recycled about 5000 kg of bottom sediment phosphorus per year in Lake
Shagawa, Minnesota. This phosphorus release, according to Lie (1978), was
sufficient to maintain the high phosphorus concentrations observed in Lake
Shagawa after nutrient reduction programs were initiated.
Aquatic plants can also indirectly affect cycling of phosphorus. When
large quantities of organic matter are sedimented to bottom water, decompo-
sition often produces anoxia. Mortimer (1941, 1942, 1971) has shown that
when the oxygen supply is depleted, phosphorus is released from sediments.
Mortimer further noted that this release increased markedly as the redox
potential decreased.
There is little doubt that aquatic plants can affect phosphorus concen-
trations in lakes. Perhaps the most pressing research needs involve elucida-
tion of the role of aquatic macrophytes in the cycling of phosphorus. Lab-
oratory studies strongly indicate aquatic macrophytes assimilate phosphorus
in water from bottom sediments. These laboratory studies also indicate that
macrophytes release phosphorus. In order to manage algal problems, water
resource managers need to have quantitative information on the quantity of
inflowing phosphorus that aquatic macrophytes will assimilate and how these
quantities vary with changes in aquatic plant biomass. Quantitative informa-
tion must also be obtained on how much phosphorus aquatic macrophytes release
to water if nutrient reduction programs are to be effective. Lie's (1978)
work suggested that aquatic macrophytes release sufficient phosphorus to
maintain Lake Shagawa in a eutrophic state. An important question is how
much reduction in aquatic plant biomass is required to reduce phosphorus
release to tolerable rates. Additional research should also be conducted
to determine how different quantities of aquatic macrophytes affect the
phosphorus budgets of different aquatic systems.
31
-------
Silica, Iron, Manganese and Sulfur. Seasonal variations in silica
concentrations (Figure 23) are directly correlated with development of diatom
populations (Meloche et al. 1938; Lund 1964; Wetzel 1975). Kilham (1971) has
observed that diatoms can reduce surface water silica concentrations to less
than a few micrograms per liter. After death and sedimentation, mineraliz-
ation causes increases in hypolimnetic silica concentrations (Stangenburg
1961).
Aquatic plants seldom directly influence the cycling of iron, manganese,
and sulfur in the aquatic environment, although aquatic plants do have meta-
bolic demands for these elements. The major effect of aquatic plants on
cycling of these elements results from depletion of oxygen during decompos-
ition of aquatic plant remains. With loss of oxygen from water overlying
sediment, redox potential falls (Wetzel 1975), and when low enough, iron and
manganese can be released from the sediments. Sulfate can be reduced to
hydrogen sulfide and additional hydrogen sulfide can be produced by bacterial
decomposition of sulfur-containing organic matter.
There is a lack of quantitative data that can be used to predict the
effect of aquatic plant biomass on the silica, iron, manganese, and sulfur
cycles of aquatic systems. Because these elements are important water qual-
ity parameters in many domestic and industrial water uses, water resource
managers need data that can be used to predict their concentrations. It is
especially important to be able to relate anoxia to amount of dead organic
matter in order to predict release of iron, manganese, and sulfur from the
sediments.
Taste, Odor, and Toxic Substances. Taste and odor are major water qual-
ity problems. Excessive growth of certain aquatic plants can produce taste
and odor in water (Palmer 1959; Neel et al. 1961). The plants that cause
taste and odor problems most often are phytoplankton, especially certain
diatoms, blue-green algae, and pigmented flagellates (Palmer 1959). These
algae can impart a fish taste to the water as well as cause a fishy, musty,
septic, or pigpen odor (Table 6).
Studies by Prescott (1948), Ingram and Prescott (1954), and Rose (1953,
1954) suggest that blue-green algae and dinoflageHates can produce toxic
substances that can kill fish, birds, and mammals. These toxic substances
seem to be more lethal when there are dense growths of algae. Prescott
(1968) noted that domestic animals such as horses, cattle, sheep, hogs, and
birds have been killed or made severely ill by drinking algae-infested
water. He noted that death often resulted within 1 to 24 hours. Human
consumption of fish and shellfish which have injested large quantities of
dinoflagellates, especially Gonyaulax, can also result in death or serious
poisoning (Prescott 1968).
A major research need is definition of factors that enhance growth of
algae that produce taste, odor, and toxic substances. In addition, quanti-
tative information is needed on the amount of aquatic plant biomass that
produces problems. With this information it might then become possible to
manage taste and odor prolems as well as tpxicity problems.
32
-------
2
0 50 100 150 0
/jg N FIXED m J day '
100
200
00
00
Figure 22. Variations in nitrogen fixation with depth (A) in Lake Winder-
mere, and (B) in Esthwaite Water, England, 30 August 1966.
(Modified from Home and Fogg, 1970). From Wetzel, 1975.
' ' I9SG
I T I I I I I I I t II I I I I I I I I I I I I I I I
I I I I t I I I I
Figure 23. The periodicity of the diatom algae Asterionella formosa, Fragi-
laria crotonensis, and Tabellaria flocculosa in relation to fluc-
tuations in the concentration of dissolved silica, 0-5 m in Lake
Windermere, England, 1945 - 1960. From Lund, 1964.
-------
Table 6.
Odors, tastes, and tongue sensations associated with algae in water.
From The Practice of Water Pollution Biology, 1969.
Odor when algae are
Tongue
Algal genus Moderate Abundant Taste sensation
Actinastrum Grassy, musty
Anabaena Grassy, nasturtium, Septic
musty.
Anabaenopsis , Grassy
Anacystis Grassy Septic Sweet
Aphanizomenon Grassy, nasturtium, Septic Sweet Dry:
musty.
Asterionella Geranium, spicy Fishy
Ceratium Fishy Septic Bitter
Chara Skunk, garlic Spoiled, garlic
Chlamydomonas Musty, grassy Fishy, septic Sweet Slick.
Chlorella Musty
Chrysosphaerella Fishy
Cladophora Septic
(Clathrocystis)
Closterium Grassy
(Coelosphaerium)
Cosmarium Grassy
Cryptomonas Violet Violet Sweet
Cyclotella Geranium Fishy
Cylindrospermum Grassy Septic
Diatoma Aromatic
Dictyosphaerium Grassy, nasturtium Fishy
Dinobryon Violet Fishy Slick.
Eudorina Fishy
Euglena Fishy Sweet
Fragilaria Geranium Musty
Glenodinium Fishy Slick.
(Gloeocapsa)
Gloeocystis Septic
Gleotrichia Grassy
Gomphosphaeria Grassy Grassy Sweet
Gonium Fishy
Hydrodictyon Septic
Mallomonas Violet Fishy
Melosira Geranium Musty Slick.
Meridion Spicy
(Microcystis)
Nitella Grassy Grassy, septic Bitter
Nostoc Musty Septic
Oscillatoria Grassy Musty, spicy
Pandorina Fishy
Pediastrum Grassy
Peridinium Cucumber Fishy
Pleurosigma Fishy
Rivularia Grassy Musty
Scenedesmus Grassy
Spirogyra Grassy
Staurastrum Grassy
Stephanodiscus Geranium Fishy Slick.
Synedra Grassy Musty Slick.
Synura Cucumber, muskmelon, Fishy Bitter Dry, metallic,
spicy. slick.
Tabellaria Geranium Fishy
Tribonema Fishy
(Uroglena)
Uroglenopsis...' Cucumber Fishy ". Slick.
Ulothrix Grassy
Volvox Fishy Fishy
-------
Impact of Aquatic Plants on Biological Characteristics
Productivity. In aquatic systems, plants are the primary producers of
organic matter on which many organisms depend for food. Because most inves-
igators assumed phytoplankton were the primary producers of readily digest-
stable organic matter, there now exists extensive literature on algal primary
production (Table 7; Wetzel 1975). Studies have shown that annual algal
primary production rates can range from about 4 g C/m /yr (Kalff and Welch
1974) to over 600 g C/nT/yr (Tailing 1965; Lewis 1974) in natural lakes.
Wetzel (1975) summarized the literature and noted that primary production
rates in oligotrophic lakes ranged from 4 to 24 g C/nr/yr while mesotrophic
lakes ranged from 75 to 250 g C/rrr/yr and eutrophic lakes ranged from 350
to 700 g
Although it has long been recognized that productivity is greater in
nutrient-rich aquatic systems, there has been considerable debate on the
factors that control algal primary productivity. In single lake studies, it
has often been difficult to show that algal productivity increases with in-
creases in nutrients. Smith (1979), by using data from a large number of
lakes, has shown that mean growing season photosynthesis expressed per unit
volume of the euphotic zone is highly correlated to mean total phosphorus
concentrations (Figure 24) and mean chlorophyll concentrations (Figure 25).
These data are particularly useful to water resource managers as they permit,
for the first time, the prediction of the impact of changing nutrient inputs
on algal productivity and the impact of controlling algal populations on
algal productivity. Because the relationships in Figures 24 and 25 are
derived from a group of lakes representing a wide range of limnological
conditions, the results from Smith's (1979) study should be applicable to
a range of lakes.
Although most of the work on aquatic plant production in the aquatic
environment has centered on algal primary production, the early works of
Rickett (1922, 1924) and Wilson (1935, 1937) showed that the weight of the
total crop of aquatic macrophytes in lakes could be significant. Since this
work, a number of studies have measured aquatic macrophyte biomass and prod-
uctivity (Table 7). The importance of aquatic macrophytes to the total prod-
uctivity of the aquatic environment, however, was often overlooked because
many workers considered aquatic macrophytes to be an unimportant food source.
Odum and de la Cruz (1963), however, suggested that the organic matter pro-
duced by aquatic macrophytes was an important source of detritus which sup-
ported aquatic organisms. Studies by Nygaard (1958), West lake (1963), and
Wetzel (1964) further suggested that aquatic macrophytes were extremely
important in determining the total productivity of the aquatic environment.
Recent studies have provided data which support this hypothesis (Table 8) .
Rich et al. (1971) showed that in a southern Michigan marl lake aquatic
macrophytes contributed 48% of the total annual production while phyto-
plankton contributed 30%. Wetzel et al. (1972) showed that macrophytes
contributed 51% of the total annual production in Lawrence Lake, Michigan,
while phytoplankton contributed 25% and attached algae contributed 24%.
These data strongly suggest that while photoplankton may contribute as much
as 99% of the total production in some lakes (Schindler et al. 1973), the
role of macrophytes and attached algae in other lakes must be considered.
35
-------
Table 7.
Comparison of rates of primary production of phytoplankton in selected fresh waters of varying fertility3 and
representative estimations of annual above-ground biomass and productivity of aquatic macrophytes. From Wetzel,
1975.
Lake
OLIGOTROPHIC:
Char, N.W.T., Canadian
arctic (lat. 74°)
(Kalff and Welch, 1974)
Meretta, N.W.T., Canada
(Kalff and Welch, 1974)
Castle, Calif.
(Goldman, pers. comm.)
Lunzer Untersee,
Austria (Jonasson, 1972)'
Lawrence, Mich.
(Wetzel; cf. Table 14-7)
Ransaren, Sweden
(Rodhe, 1958)
Lake Superior, USA-Canada
(Putnam and Olson, 1961)
Lake Huron, USA-Canada
(Vollenweider, et al., 1974)
Borax, Calif.
(Wetzel, 1964)
Lake Michigan, USA
(Vollenweider, et al., 1974)
Lake Ontario, USA-Canada
(Vollenweider, et al., 1974)
MESOTROPHIC:
Erken, Sweden
(Rodhe, 1958)
Clear, Calif.
(Goldman and Wetzel, 1963)
Esrom, Denmark
(Jonasson and Mathiesen,
1959)
Fureso, Denmark
(Jonasson and Mathiesen,
1959)
Walter, Ind.
(Wetzel, 1973)
Basin A
Basin B
Basin C
Basin D
Oliver, Ind.
(Wetzel, 1973)
Olin, Ind.
(Wetzel, 1973)
Martin, Ind.
(Wetzel, 1973)
Pretty, Ind.
(Wetzel, 1966b)
1963
1964
Remarks
80% of total production
by benthic flora
Polluted by sewage
Deep, alpine
Small, alpine
Small, hardwater;
7-year averge
Small lapplandic
Most unproductive
of Laurentian Great Lakes
Offshore stations
Large, shallow saline lake
(25% of total productivity)
Offshore stations
Offshore stations
Large, deep, naturally
productive
Very large, shallow
Large, moderately deep
Large, deep, many
macrophytes
Series of 4 interconnected
marl lakes
Large, deep, marl lake
Large, deep, marl lake
Deep, stained marl lake
Moderate-sized, deep, marl
lake
Mean Daily
Productivity
for Entire Year
(mg Cm"2 day ~']
1.1
3.1
98
(123)
112.6
249
285
438
370
462
418
210
276
437
336
374
561
440
305
Range Observed
(mg Cm"2 day "')
0-35
0-170
6-317
5-497
23-66
50-260
150-700
10-524
70-1030
60-1400
40-2205
2-2440
23-422
0-1380
102-1395
12-535
30-1048
98-1458
32-775
89-996
27-1708
68-1850
6-895
Annual
Productivity
(g C m "2year "')
4.1
11
36
45
41.1
calOO
91
ca 130
calSO
104
160
260
168
153
77
101
160
123
137
205
161
111
36
-------
Table?.
Continued.
Lake
Crooked, Ind.
(Wetzel, 1966b)
1963
1964
Little Crooked, Ind.
(Wetzel, 1966b)
1963
1964
Goose, Ind.
(Wetzel, 1966a)
Lake Erie, USA-Canada
(Vollenweider, et al., 1974)
Western stations
Central stations
Eastern stations
EUTROPHIC:
Wintergreen, Mich.
(Wetzel, et al., unpublished)
Frederiksborg Slotsso
Denmark (Nygaard, 1955)
Minnetonka, Minn.
(Megard, 1972)
Sollerod So, Denmark
(Steemann Nielsen, 1955)
Sylvan, Ind.
(Wetzel, 1966a; cf.
Table 14-9)
Lanao, Philippines
(Lewis, 1974)'
Victoria, Africa
(Tailing, 1965)
DYSTROPHIC:
Kattehale Mose, Denmark
(Nygaard, 1955)
Smit Hole, Ind.
(Wetzel, 1973)
Store Gribso, Denmark
(Nygaard, 1955)
Grane Langso, Denmark
(Nygaard, 1955)
Remarks
Large, deep, hardwater lake
Small, deep, kettle lake
Small kettle lake
Shallow, extensive nutrient
loading
shallow, enriched
Extremely complex basin.
large, deep
Complex basin, large.
shallow
Large, deep tropical lake
Large, deep, equatorial lake
Very shallow, acidic, peat
bog
Shallow, humic stained
Deep, acidic, humic stained
Deep, acidic, humic stained
Mean Daily
Productivity
for Entire Year
(mg Cm"2 day "')
469
359
618
508
729
.
1012
1030
(820)
1430
1564
1700
1750
80
194
230
248
Range Observed
(mg Cm"2 day "')
142-1364
23-870
263-1903
9-2431
166-1753
30-4760
120-1690
140-1440
60-2240
12-4160
0-3800
9-4959
400-5000
1 700-3800
0-400
24-5960
4-680
20-880
Annual
Productivity
(g C m ~2 year "')
171
131
226
218
266
(310)
(210)
(160)
369
376
(300)
522
570
620
640
29
71
84
91
37
-------
Table 7.
Continued.
Type and Lake
SUBMERGENTS DOMINATING:
Trout L, Wise, (softwater)
Sweeney L., Wise, (softwater)
Weber L, Wise, (softwater)
Lowes L., Scotland (dystrophic)
Spiggie L., Scotland (dystophic)
L. Mendota, Wise, (hardwater)
Lawrence L., Mich, (hardwater)
Submersed Scirpus
subterminalis
Chara
Annuals
Croispol, Scotland
Borax L., Calif, (saline lake.
Ruppia)
River Ivel, England (Berula;
very fertile)
River Test, England (Ranunculus)
River Yare, England
(Potamogeton)
Saline channels, Puerto Rico
(Jhalassia)
FLOATING:
New Orleans, La. (Eichhorina)
EMERGENTS DOMINATING:
Ladoga L., USSR
Onega L, USSR
Blanket bog, England (Sphagnum)
on hummocks
in pools
on '.'lawns'
Polish lakes (emergent species)
Minnesota wetlands (Carex)
Surlingham Broad, England
(Glyceria, Typha, and Phragmites)
Opatovicky Pond, Czechoslovakia
(Phragmites)
Aerial
Below ground
Cedar Creek, Minn. (Typha)
Seasonal Maximum
Biomass or Above-
Ground Biomass
(g dry m"2)
0.07
1.73
16.8
32
100
202
338
110
130
400
60
500
100-400
380
700-7300
630-1472
0.4-10.7
0.1-33
440-830
850
800-1100
1100-2200
6000-8560
4640
Productivity
(g m"2 year"')
565
155
199
64
1 500-4400
180
290
340
738
2500
Source
Wilson, 1941
Wilson, 1937
Potzger and Engel,
1942
Spence, et al., 1971
Rickett, 1921
Rick, et al., 1971
Spence, et al., 1971
Wetzel, 1 964
Edwards and Owens,
1960
Owens and Edwards,
1961
Owens and Edwards,
1962
Burkholder, et al.,
1959
Penfound and Earle,
1948
Raspopov, 1971
Clymo and Reddaway,
1971, 1974
Szezepariska, 1973
Bernard, 1973
Buttery and Lambert,
1965
Dykyjova and
Hradecka, 1973
Bray, et al., 1959
38
-------
2000
'E
u
Ut
E
1000 -
50
100
TP, mg Total Phosphorus -m
150
-3
200
Figure 24. Correlation between growing season mean photosynthesis (per unit
volume euphotic zone) and mean total phosphorus concentration for
38 north temperate lakes. From Smith, 1979.
39
-------
2000
i>
20
30 40 50 60
mg Chlorophyll-m"3
70
80
Figure 25. Correlation between growing season mean photosynthesis (per unit
volume euphotic zone) and mean chlorophyll concentration for 49
North American lakes. Limits shown are 95% confidence limits for
individual points around regression line. -Lake Washington;
D - Tuttle Creek Reservoir; & -European lakes. From Smith, 1979.
The recent work of Smith (1979) is most promising. Additional research
on other lakes should be done to further expand the data base, but more im-
portant quantitative data should be gathered on aquatic macrophytes and
organisms. Studies by Nygaard (1958), Westlake (1963) and Wetzel (1964)
further suggested that aquatic macrophytes were extremely important in
determining total productivity of the aquatic environment. Recent studies
have provided data that support this hypothesis (Table 8). Rich et al.
(1971) showed that in a southern Michigan marl lake aquatic macrophytes
contributed 48% of the total annual production while phytoplankton
contributed 30%. Wetzel et al. (1972) showed that macrophytes contributed
51% of the annual production in Lawrence Lake, Michigan, while phytoplankton
contributed 25% and attached algae 24%. These data strongly suggest that,
while phytoplankton may contribute as much as 99% of the total production in
some lakes (Schindler et al. 1973), the role of macrophytes and attached
algae must be considered in other lakes.
40
-------
Table 8.
Examples of annual net productivity of phytoplankton, littoral algae, and macrophytes of several lakes in which
productivity estimates of attached algae were made on natural substrata. From Wetzel, 1975.
Mean
Area Depth
Lake (ha) (m)
Borax. Calif 39 8 XD 5
Phytoplankton
Littoral algae
Macrophytes
Marion, British
Columbia 133 22
Phytoplankton
Littoral algae
Macrophytes
Lake 239, Ontario 56 1 105
Phyioplankton
Littoral algae
Macrophytes
Lake 240. Ontario 44 1 61
Phyioplankton
Littoral algae
Macrophytes
Lawrence. Mich. 5.0 5 9
Phytoplankton
Littoral algae
Macrophyies
Wmgra. Wise 1396 ca 2
Phyioplankton
Methaphyton (Summer, 1971)
(Oedogonium) (Summer. 1972)
Macrophytes
Annual
Mean
(mg C m"2
day")
2493
731 5
765
21 9
1096
49.3
118 9
2003
2408
1200
30
55
3205
Annual
Mean (kg C
lake"' day"')
101 0
75.5
1 36
029
11 3
65
2153
1977
4360
1675
4 2
7 6
447
Kg C ha"
Surface
926
692
12
1630
8
310
180
498
823
8 1
N.D
ca 831
501
90
N 0
ca510
434
399
879
1712
4380
11 1
19 9
1170
5581
of Lake
year
(%)
(568)
(425)
(07)
(1 6)
(622)
(361)
(990)
(1 0)
(982)
(1 8)
(254)
(233)
(51 3)
(786)
(0 4)
(21 0)
Remarks
Saline lake, benthic algae, primarily epilithic,
some epiphytic and metaphyton. single macrophyte
species Ruppia mantima. '*C methods for all
components (Wetzel, 1964)
Sofrwater, oligotrophic lake, benthic algae.
primary epipelic, 02 techniques, from which
net production was estimated (Efford, 1967;
Margrave, 1969. Gruendling, 1971)
Softwaier, oligotrophic lake, probably under-
estimates since winter production is not included;
benthic algae, primarily epilithic, macrophytes
probably insignificant, COy utilization methods
(Schmdler. el al . 19731
(Same as above for Lake 239)
Hardwater. obligotrophic marl lake, benthic
algae, primarily epiphytic on sparse submersed
macrophytes. "C methods (Wetzel. et al . 1972)
Large, shallow hardwater eutrophic lake, large
littoral zone with dominant submersed mac-
rophyte Mynophytlum and metaphytic mats
of macroalga Oedogonium, "C methods for all
components, mostly only summer values
(McCracken. et al . 1 974. Adams and McCracken.
1974. J F Koonce. personal communication)
41
-------
Interactions Between Macrophytes and Phytoplankton. In recent years,
there has been considerable interest in the relationship between algal and
aquatic macrophyte growth. This interest has developed because of concerns
about what the impact of reducing algal or aquatic macrophyte growth might
be on subsequent algal or aquatic macrophyte growth. A number of investiga-
tions have suggested that an inverse relationship exists between development
of phytoplankton and aquatic macrophyte populations (Kofoid 1903; Scnreiter
1928; Hasler and Jones 1949; Hogetsu et al. 1960; Stangenburg 1968; Goulder
1969- Dokulil 1973) (Table 9). This relationship may exist because of
shading by the aquatic plants (Hasler and Jones 1949; Dendy 1963; Westlake
1968- Goulder 1969), nutrient competition (Embody 1928; Weibe 1934; Bennett
1942- Fitzgerald 1969) or because the plants secrete inhibitory substances
(Fitzgerald 1969). Other studies (Nichols 1973; Gasith et al. 1976; Lie
1978) suggest that aquatic macrophytes may enhance development of phyto-
plankton populations by mobilizing sedimentary phosphorus and releasing it
to the water where phytoplankton assimilate the phosphorus.
Management of phytoplankton and aquatic macrophyte populations requires
that water resource managers be able to predict how changes in phytoplankton
and aquatic macrophyte densities interact. There are no quantitative data
available that can be used to determine how phytoplankton density affects
macrophyte density and vice versa. There is a strong need for quantitative
data on phytoplankton and aquatic macrophyte biomass from a large number of
aquatic systems representing a range of physical, chemical, and biological
conditions in order to determine the relationship between phytoplankton and
aquatic macrophyte biomass. This information would be especially useful to
water resource managers because it would permit prediction of how these
plants would respond to control. For example, if there is an inverse rela-
tionship between phytoplankton and aquatic macrophytes it might be possible
to manage one group of plants to a level where phytoplankton and aquatic
Table 9.
Effect of dense aquatic plant growth on abundance of phytoplankton.
From Hasler and Jones, 1949.
Plant-free Plant-filled
105 organisms per liter 105 organisms per liter
Year . .
Pond . .
Month
July
Aug.
Sept.
1 945
A
D
1946
B C
1 P45 1 QAfi
B
C
A
D
Week
1
2
3
4
1
2
3
4
1
M
99
15
21
31
8.
7.
130.
134.
143.
65.
.00
.60
.00
60
.40
80
00
00
00
60
15
10
2
44
26
56
56
39.
102.
39
.40
.60
.70
.60
.00
.00
.00
01
00
13
9.13 2.07
47.82 8.59
28.06 15.86
50.00 24.40
90.28 13.42
272.06 42.70
394.06 9.42
No sample
No sample
127.35 16.63
71
5
2
2
3
6.
3
4.
16.
12.
.00
30
.90
.90
.20
30
80
10
00
83
4.80
5.60
5.40
3.80
2.00
4.40
1.90
4.30
9.30
4.61
4.39
4.39
2.70
3.05
1.95
5.24
8.27
10.97
11.91
2.56
31.72
5.61
7.17
1.70
No sample
No sample
4.27
10.23
42
-------
macrophyte problems are both minimized. However, if aquatic macrophytes
support phytoplankton growth by releasing sedimentary phosphorus, it might
not be possible to control algal biomass through reductions in nutrient
inputs until aquatic macrophyte populations are controlled.
Effects on Invertebrates. There has been considerable interest con-
cerning the effect of aquatic plants on invertebrate populations that are
often the main food source for fish. Early workers (Forbes 1887; Reighard
1915; Baker 1918; Shelford 1918; Wilson 1924; Needham 1929, 1938; Surber 1930
Pate 1932, 1934; and many others) noted that there was often a considerable
increase in the population of invertebrates when aquatic macrophytes were
present. Lundbeck (1927) showed that in central European ponds, plant-free
bottom areas supported 6.28 g of organisms/m2; the submersed aquatic
macrophytes 6.41 g/m2; the emergent plants 8.29 g/m . Surber (1930)
showed that snails were six times more abundant in weedy areas than non-weedy
areas. In streams, the studies of Needham (1929) and Pate (1932, 1934) have
shown that pools with weedy vegetation had between 17 and 38 times more
invertebrates than bare pools.
The greater numbers of invertebrates in macrophyte areas has been
attributed to the ability of plants to provide a substrate and shelter
(Shelford 1918). Shelford (1918) noted that Elodea was an excellent plant
for production of invertebrates. Andrews and Hasler (1943) noted that
aquatic macrophytes with the most dissected surface area harbor the largest
populations of invertebrates (Table 10). Martin and Shireman (1976) found
that many invertebrates (450 g/kg vegetation) live in hydrilla. Welch (1952)
noted that aquatic macrophytes provided a place of attachment to bryozoans,
mollusks, annelids, and insects. Aquatic macrophytes may also increase
invertebrate production indirectly by creating a favorable habitat. For
example, by reducing wave action, macrophyte beds create important nursery
areas for mosquitoes (Bishop and Hollis 1947; Beadle and Harmston 1958;
Myklebust and Harmston 1962).
Table 10.
Number of organisms per kilogram of different types of aquatic plants. From Andrews and Hasler, 1943.
Most productive Ceratophyllum demersum 52.000 animals per kg. of plant
(coontail)
Myriophyllum exalbescens 29,000 animals per kg. of plant
(water milfoil)
Moderately productive Potamogeton pectinatus 21,000 animals per kg. of plant
Chara sp. 17-20,000 animals per kg. of plant
Less productive Potamogeton americanus 18,000 animals per kg. of plant
Richardsonii 10,000 animals per kg. of plant
amplifolius 5,000 animals per kg. of plant
Poorly productive Vallisneria americana 3,000 animals per kg. of plant
43
-------
Aquatic plants are an important source of food for invertebrates.
Rawson (1955) showed that bottom fauna biomass (Figure 12) increased as
lakes become shallower and more productive. Noonan (1979) showed that total
number of zooplankton was directly related to algal biomass (Figure 26).
These studies strongly suggest that systems of high primary productivity
support more invertebrates than systems of low primary productivity.
Aquatic plants can enhance development of invertebrate populations, but
can adversely affect the animals directly and indirectly. Toxins released by
algae and macrophytes can kill invertebrates. Hutchinson (1975) reviewed the
literature demostrating that both aquatic macrophytes and macroalgae of the
family Characeae secrete inhibitory or insecticidal organic compounds that
affect mosquitoes. Pennak (1973) also reviewed literature that showed
organic secretions to inhibit or repel zooplankton. McLachlan (1970) report-
ed that obstruction of light by plants caused a reduced diversity and biomass
of benthic fauna. A major indirect effect occurs when plant decomposition
reduces oxygen concentration below the tolerance levels of invertebrates.
There is a great deal of information on the relationships of aquatic
plants to invertebrate populations, but most of the data are based on animal
numbers or are only qualitative. There is, as of yet, no means by which
water resource managers can predict the effect of aquatic plant growth on
invertebrate populations. Biomass data on phytoplankton and aquatic macro-
phytes should be related to invertebrate biomass data from a range of aquatic
systems. This should permit development of general equations to predict the
impact of phytoplankton or aquatic macrophytes on invertebrate populations.
Fish. There has been considerable interest on relationships of plant
and fish populations. Rawson (1955) showed that commerical fish production
was inversely related to mean lake depth. His data strongly suggested that
more productive lakes support more fish. However, as lakes become more
productive, composition of fish communities may change, resulting in devel-
opment of undesirable fish populations. Dense populations of aquatic plants
can also reduce fish populations. Studies by Prescott (1932), Carl (1937),
Prescott (1948), Mackenthun et al. (1948), Lefevre et al. (1952), and
Shelubsky (1951) suggest that toxins released by blue-green algae have killed
fish. When aquatic plants die and decompose, oxygen concentration is
reduced, resulting in summer or winter fish kills (Tomlinson 1935; Sears
1936; Hutchinson 1936; Moore 1942).
Effects of aquatic macrophytes on fish populations are still debated.
Early workers (Reighard 1915; Welch 1916, 1924; Baker 1918; Klugh 1926;
Frohne 1938) noted that aquatic macrophytes supported many organisms eaten
by fish. Klugh (1926), who reviewed much of the literature on relationships
of invertebrates to aquatic macrophytes, concluded that the plants could be
used as an index of fish production. Bailey (1978) reported that condition
factors for bluegill and redear sunfish in Arkansas generally improved with
removal of aquatic vegetation by grass carp. Colle and Shireman (1980)
44
-------
107-
LU
o
CO
106-
105-
8 1°4'
M
103-
LOG ZOOPLANKTON ABUNDANCE =
0.59 LOG CHLORPHYLL a + 4.3
Present Study
Walker 1975
Patalas 1971
Haertel 1976
Anderson et al. 1 955
Watson and Carpenter. 1974
Stockner and Northcote 1974
.1
10
CHLOROPHYLL a MG/M3
100
1000
Figure 26. Lake mean zooplankton abundance versus mean chlorophyll a for
data from the literature and this study. Values from Patalas
(1971) include mean depth data from Brumskill and Schindler
(1971) and chlorophyll £ data from Armstrong and Schindler
(1971). From Noonan, 1979.
45
-------
found that condition factors of bass, bluegills, and redears were influenced
by the amount of hydrilla in two Florida lakes. Harvestable largemouth bass
had low condition values once hydrilla coverage was greater than 30%; how-
ever, smaller bass were not as adversely affected until cover exceeded 50%.
Bluegill and redear condition and weight-length relationships were not af-
fected by hydrilla until it occupied the majority of the water column. Cope
et al. (1969) and Cope et al. (1970) found that bluegill and redear sunfish
in Oklahoma ponds grew faster when submersed vegetation was controlled, and
reasoned that a greater amount of food was available for growth in such situ-
ations. Ricker (1942) reported no association between bluegill growth rates
and abundance of aquatic vegetation in a series of Illinois ponds. Buck and
Thoit (1970) attributed delayed bass spawning to dense Potamogeton mats,
which caused elevated pH (10.2). The fish spawned after vegetation- was re-
moved and pH values reduced. Stocking of Lake Lichen with grass carp caused
alterations in fish species composition and number. Spawning grounds for
some fish were destroyed as grass carp ate vegetation (Opuszynski 1972).
Other studies (Hotchkiss 1941; Tilghman 1962; Wahlquist 1969; Gaevskaya 1969;
and others) have shown that aquatic macrophytes provide sites for attachment
of eggs and shelter for adult and larval animals, leading to increases in
population size of some fish species.
Other studies suggest that aquatic macrophytes are detrimental to fish
populations. Smith and Swingle (1941a, 194lb) and Rasmussen and Michaelson
(1974) suggested that dense growths of aquatic macrophytes cause overcrowding
by providing shelter to small fish. They also suggested that aquatic macro-
phytes sequester nutrients which could be used by phytoplankton, zooplankton,
and fish. Bennett (1948) showed an apparent inverse relationship between
fish yields and densities of aquatic macrophytes and suggested that macro-
phytes reduced fish production. Surber (1961) suggested dense stands of
macrophytes interfered with fish production. Boyd (1967), however, noted
that while many workers consider aquatic macrophytes detrimental to fish
production, there are very little data that show any correlation between
production and macrophytes biomass.
CONTROL METHODS IN THE SUNBELT
Floating Aquatic Weeds
Historically, floating aquatic weeds have caused problems in the Sunbelt
states. They are restricted to this area primarily because they cannot
survive prolonged freezing and ice cover. The most widely spread floating
aquatic plant requiring control is the water hyacinth, which is believed to
have originated in South America. It was first introduced into Louisiana
and soon spread throughout the South.
46
-------
Prior to 1950, chemical control of water hyacinth was accomplished by
spraying the foliage with sodium arsenite, copper sulfate, and other desic-
cating mineral salts. Since 1950, water hyacinths have been controlled
principally by 2,4-D (2,4-dichlorophenoxy acetic acid). Several salts,
esters, and formulations of 2,4-D have been tested, and over 90% of the water
hyacinths that are controlled chemically are with 2 to 4 Ib acid
equivalent/acre of the dimethlamine salt. The liquid 2,4-D formulation
(Weedar 64) is registered for use on water hyacinths and has been granted
fish and potable water tolerances. In 1977, it was estimated that
approximately 300,000 Ib of 2,4-D were used in Florida by state, federal, and
local agencies responsible for water hyacinth control. Control programs in
Florida cost an estimated $3.5 million and resulted in spraying of about
100,000 acres of water hayacinths (Haller 1976).
Because this herbicide formulation is a water-soluble liquid, applica-
tion of 2,4-D to water hyacinth is simple and accomplished by several means,
including hand-held spray guns, truck-mounted sprayers, and fixed wing or
helicopter aircraft. Most water hyacinth spraying, however, is accomplished
from airboats. Spray crews composed of an airboat and spray operator, can
spray 10 to 20 acres of water hyacinth daily. The majority of water hya-
cinth spray programs require a 16-foot airboat, a 10-gallon per minute piston
pump, and a saddle tank for holding the 2,4-D. The chemical is metered into
the suction side of the pump by inserting orifice plates of various sizes
into the the delivery line. The water portion of the spray mix is obtained
through water ports at the back of the airboat. Typically, about 70 to 100
gallons of spray mix are applied with a handgun to each acre of water hya-
cinths. Spray mix (100 gallons for example) contains 99 gallons of water and
one gallon (4 Ib/gal) of 2,4-D. Most water hyacinth control programs keep
the plant under maintenance control. Hyacinths are sprayed in isolated mats
and in fringes of swamps and backwaters to maintain a tolerable number of
plants because eradication is impossible in most watersheds, and a few plants
help support a diversity of fish and wildlife.
Acreage of water hyacinth sprayed in a particular watershed varies from
year to year. A 2- or 3-year succession of mild winters in the South permits
vigorous water hyacinth growth the following summer and frequent and exten-
sive control operations are often needed. Periods of heavy rainfall can
raise water levels and flush water hyacinths from backwaters into flowing
streams, lakes, and reservoirs where they must be controlled or they would
fill the waterway.
Other floating aquatic plants which periodically require chemical treat-
ment are water lettuce, Pistia sp., and duckweed. Water lettuce is found
only in tropical areas, whereas duckweed is a cosmopolitan species occurring
throughout the United States. In states that have functioning aquatic weed
control programs, both water lettuce and duckweed are treated similarly to
water hyacinths, but a different chemical (diquat) is used. Generally,
2,4-D does not kill water lettuce and duckweed. Diquat is substituted for
2,4-D in the spray mix and sprayed directly on the foliage. Other chemicals
are used in some states for duckweed control, legally and illegally, depend-
ing upon the state registrations. Substituted ureas, triazines, and other
herbicides applied at low rates (2 to 3 Ib/acre) control duckweed very
effectively particularly in small ponds.
47
-------
Recent studies of dense water hyacinth infestations conducted at the
University of Florida have delineated the environmental impact of this
species on the aquatic ecosystem. Three 2-acre ponds were selected as test
sites for detailed studies. Two ponds were chemically treated for hydrilla
control, leaving the third pond as a check.
Water beneath hyacinth mats is devoid of life. Neither oxygen nor light
penetrates beyond one-third meter, and anaerobic conditions prohibit fish
survival. Natural organic turnover of the hyacinth mat (standing crop of
12 tons of dry matter per acre) leads to organic siltation rates of approx-
imately 0.2 kg/n^/yr. Chemical control and eradication of water hyacinths
produced sediment accumulation of 0.6 kg/nr/yr of treatment. On a longterm
basis (10 years, for example), the natural untreated water hyacinth mat con-
tributes 2 kg/m2 dry matter to the sediment. Eradication of the mat pro-
duces 0.6 kg/m of water hyacinth sediment plus whatever may be contribute
by the phytoplankton in the remaining years after hyacinth control.
In conclusion, water hyacinths are the major floating weeds requiring
treatment in the United States. Most water hyacinths are under maintenance
control, and minimal spraying is required to keep them in check. The broad-
leaf herbicide 2,4-D is the most widely used to control water hyacinths and
may be applied by hand-held spray equipment or from the air in major plant
population areas. If the chemical control of water hyacinths was banned,
there are no substitute control means, and vast water bodies in the South
would become covered with water hyacinths.
Emergent and Ditchbank Weeds
Although often overlooked as a group, ditchbank and emergent weeds com-
prise a serious problem in irrigation and drainage districts, particularly in
southern Florida and in the western United States. The problem with this
group of weeds is their wide diversity. Saltcedar in the Southwest, grasses
in general in the West, and Melaleuca, Brazilian pepper, willows, and various
grasses in the Southeast create problems in irrigation ditches.
Economic data for the control of ditchbank weeds are difficult to assem-
ble and are at best rough estimates. In 1977, Florida expended $8.2 million
for ditchbank aquatic weed control. The U.S. Department of Agriculture
(USDA) reported in 1968 that 27 states chemically treated 1.7 million acres
of ditchbanks at an average cost of $20 per acre (total cost $34 million).
Few herbicides are labelled for ditchbank, use. The generalization can
be made, however, that dalapon is used for control of ditchbank and emergent
grasses, whereas various formulations of 2,4-D are used to control broadleaf
or dicotyledonous species. These chemicals are not effective against all
ditchbank and emergent plants and those that are resistant to them cannot be
controlled by currently approved chemicals. Because these species cause lo-
cal problems, chemical industries consider them a minor problem and are not
developing new, effective chemicals for their control. Consequently, weed
control agencies increase application rates of dalapon and 2,4-D to high
levels in order to control these plants and re-treat the areas more frequent-
ly.
48
-------
Submersed Weeds
Submersed aquatic weeds are very difficult to control. The basic
reasons are: 1) phytotoxic chemicals are diluted and degraded by water and
aquatic organisms; consequently, it is difficult to get a toxic dose of the
herbicide onto or into submersed aquatic weeds; 2) most waters in the United
States have multipurpose uses and flow from one area to another. The same
water, treated to control aquatic weeds, may be used by sportfisherman, for
agricultural irrigation, sports, and ultimately as potable water; 3) a
plethora
of exotic aquarium plants have been introduced into the United States. Lit-
tle is known about their life cycles, physiology, or biology. Without this
knowledge, it is difficult to formulate treatment schedules.
Because of the diversity of submerged aquatic weeds, there currently are
many chemical treatment methods varying from dragging a burlap bag of copper
sulfate behind a rowboat to sophisticated invert emulsion applications from
an airboat.
Many submersed aquatic weeds are treated by dissolving copper sulfate in
ponds and lakes. Total volume treatment methods evolved from this tradi-
tional method to use of chemicals such as an endothall, diquat, diuron,
acrolein, and various triazine herbicides. Two major problems became evident
when total volume treatments were used: fish were frequently killed and/or
the cost of chemical treatments became exorbitant. The use of copper sulfate
was decreased due to its toxicity to fish, corrosiveness, and ineffectiveness
in hard water where copper ions precipitate. Total volume treatment with
diquat (for example, in 1 acre 10 feet deep) costs well over $300 for
chemicals alone. Less expensive chemicals (diuron, triazines, etc.) often
control submersed vegetation on a total volume basis more cheaply; however,
these chemicals are persistent and water cannot be used for irrigation, etc.,
after treatment. Endothall and acrolein are toxic to fish at total volume
concentrations that kill submersed weeds.
Although expensive short-term residual (1-3 weeks) herbicides are clear-
ed for use in aquatic situations, endothall, diquat, organic-chelated liquid
cooper, and 2,4-D are predominately used. Total volume treatments are used in
irrigation ditches of the western states where acrolein or other contact
phytotoxic chemicals are injected into the water from bridges or banks. A
recent review of herbicides in flowing irrigation water has been published
(Bowmer et al. 1979). Anhydrous ammonia has also been used successfully in
western irrigation ditches for Najas and algae control. Basically, total
volume treatment methods have been replaced by newer application methods,
particularly in the southern states where aquatic weed problems have occurred
for many years.
Deep-water injection techniques have evolved in an attempt to place
phytotoxic chemicals in close proximity to or onto submersed weeds for more
effective uptake, control, and cost reduction. Deep-water injection is used
in some areas and has been used with success with invert or polymer carrier
systems. Deep-water injection is conducted regardless of water depth from an
49
-------
airboat or other spray platform equipped with a pump capable of pumping tank
mix or chemical directly into weighted hoses 12- to 15-feet long trailing
behind the spray boat. Chemicals generally used in this application are
endothall or diquat-copper complex combinations at roughly 4 to 8 gallons of
chemical per acre. The ends of the trailing hoses are fitted with lead pipes
to hold them down in the submersed weeds.
An additional modification, the invert system, allows placement of
aquatic herbicides onto the target plants in slowly flowing water. Origi-
nally researched and developed for drift control, the invert system involves
mixing an oil phase and an aqueous phase through a blending process into a
mayonnaise-like homogenate. The homogenate is injected below the water
surface and the droplets sink onto the submersed aquatic weeds. The droplets
dissolve slowly, releasing the herbicide close to the target plant. The
invert system has wide acceptance in the southern states, particularly for
hydrilla control in Florida.
The basic mix generally used is (per acre):
Oil Phase Water Phase
4 to 8 gal diesel fuel 30 to 40 gal water
2 to 3 gal inverting oil 2 gal diquat
4 gal (weighting agent)
liqud copper
This technique requires two tanks, an invert pump, and a below-surface
injection system (weighted trailing hoses).
The previous disscussions of chemical treatment covered application of
liquid formulations. Currently, two grandular formulations are widely used
for hydrilla control [Hydout^ pellets - mono (N,N-dimethylalkylamine salt
of endothall)] and for control of Eurasian watermilfoil [(AquacleanR
granules - 2,4-D butoy ethanol ester (2,4-D BEE)].
The use of Hydout pellets for hydrilla control has increased in Florida
and other states. Endothall pellets reduce the toxicity of the compound to
fish and provide for good hydrilla control in areas where liquid herbicides
have been less effective. The pellets are applied at rates of 100 to 300 Ib
per acre (22.4% active ingredient) depending upon water depth and degree of
weed infestation. Application is generally accomplished from a granule
fertilizer spreader on the bow of a boat. The pellets sink onto the weed
mass and slowly dissolve, releasing herbicide. Aerial application of Hydout
is used for very large scale treatments and is very effective.
Milfoil is effectively controlled with 2,4-D formulations, the most ef-
fective being 2,4-D BEE. TVA reservoirs have been experiencing milfoil pro-
blems for many years, and an increase in incidence of disease carrying mos-
quitoes was directly correlated to flood conditions and severity or extent
of submersed weed infestations. Flood-water mosquitoes could be effectively
50
-------
controlled by draining standing water and regulating water levels, but biting
insects (mosquitoes and tabanid flies) associated with aquatic weeds were
controlled only by massive application of insecticides.
A large scale milfoil control program was initiated in several of the
problem reservoirs in 1973. During the summer of 1974, over 400,000 pounds
of 2,4-D granules were applied to milfoil. The operation was successful and
currently milfoil is spot-treated and maintained by fluctuating water levels
(Bates, personal communication, and Goldsby et al. 1978).
Herbicide pellets or granules are used extensively for control of cer-
tain aquatic plants. The advantages of using solid granules are ease and
safety of application, simple application equipment, and release of the herb-
icide into the zone of the target plants. A particular disadvantage of using
granules is the greater expense of manufacture. For example, the liquid
endothall in the Hydout pellets costs approximately $10 per Ib of active
ingredient. After formulation, packaging and shipping, the pelletized form-
ulation costs approximately $12 per Ib active ingredient.
Currently, the most extensive submersed aquatic weed control programs
are in Florida, Texas, and Louisiana where hydrilla is a problem. Hydrilla
infests virtually every watershed in Florida, and only a small portion of the
total acreage of hydrilla is treated annually due to economic restrictions.
For example, Orange Lake in 1977 contained 10,000 acres of hydrilla, but only
200 acres were controlled by mechanical or chemical means. It is estimated
that expenditures for hydrilla treatments in Florida in 1977 were $9.1 mil-
lion. These treatments covered approximately 45,000 acres at an average cost
of $200 per acre. Approximately 400,000 Ibs. of phytotoxic chemicals are
placed in Florida's waters each year for submersed weed control.
Chemical weed control has been emphasized in this report because govern-
mental agencies spend more than $20 million each year on aquatic weed control
in Florida, at least 90% of the expenditure is for chemical aquatic weed
control.
WATER LEVEL FLUCTUATION
Habitat alterations, such as water drawdowns, pond liming, fertilization
disking, etc., have been used with various degrees of success for vegetation
control. Water level fluctuations on large reservoirs have generated great
interest in the past decade. This method can be used in many waterways,
especially southern reservoirs, but fluctuation schedules must be established
individually for each target species. The following are examples where water
fluctuation has been used to manage aquatic vegetation:
1. Ross Barnett Reservoir in Jackson, Mississippi, was drawndown
several feet during the winter of 1978 to control Najas species;
2. Tennessee Valley reservoirs are being fluctuated regularly for
Eruasian watermilfoil control,
51
-------
3. Louisiana has drawndown lakes during winter control of Egeria,
Cabomba, and Hydrilla species;
4. Florida has drawndown Lake Myacca and Lake Ocklawaha for Hydrilla
control;
5. A 5-foot increase in water level was largely responsible for 90%
control (for 2 years) of hydrilla in Orange Lake, Florida, in 1977.
Water-level fluctuation has controlled some aquatic weeds very effect-
ively. It appears that properly timed water manipulation is a key factor,
but research has been sporadic and often improperly monitored.
Hydrilla control by drawdown is similar to chemical control because only
the above-ground portions of hydrilla are killed, leaving viable tubers in
the hydrosoil to cause reinfestation. Drawdowns timed with respect to form-
ation and germination of hydrilla tubers offer partial management of tuber
populations.
Figure 27 outlines a drawdown schedule that provides hydrilla control in
northern Florida. Drawdown 1 is considered optional because its primary pur-
pose is to stimulate tuber germination after water levels are returned to
normal the following summer. Hydrilla tubers germinate only once, and dis-
appear after producing a new plant. The optional drawdown can be conducted
any time in late winter or early spring, and its duration depends on local
conditions. The drawdown must be long enough to dry the top few inches of
the hydrosoil.
Proposed Drawdown ScheduleNorth Florida Region
Drawdown 1
I 1
optional
Tuber formation
Drawdown 2
I -i
essential
Tuber formation
Tuber germination
Jan Feb Mar Apr May Jun Jul Aug Sep Oct Nov Dec Jan
Figure 27. A drawdown scheme which will provide hydrilla control in North
Florida.
52
-------
Drawdown 2 is essential because it kills hydrilla plants which have
sprouted from tubers and prevents tuber formation the following winter.
This drawdown should be started before October (before new tubers are
formed) and should be continued until water temperatures drop to around
55 F. This inhibits tuber germination and the establishment of new plants,
thus preventing new tuber formation. Limited research indicates that after
one year of both drawdowns, followed by two years of the essential drawdown,
it will be possible to proceed with the essential drawdown every other year.
Water-level fluctuation for aquatic weed control would have greater po-
tential use if more were known about the life cycles of target species and
their responses to water-level changes. Not only could detrimental plants be
controlled, but growth of beneficial plants could be encouraged. For exam-
ple, Panicum sp. is a beneficial plant for fish production. This plant
spreads and grows rapidly in deep water when water levels are reduced.
MECHANICAL CONTROL
Mechanical control is not stressed in this report because it is not
widely used and in general is very expensive, producing only short-term
effects. Extensive research efforts have been made to develop economical
harvesting and other mechanical control methods, but most aquatic vegetation
is 92 to 96% water, which makes it expensive to harvest and place on shore.
Nevertheless, mechanical equipment of various types are available and include
the following:
1. Cutters which cut or free submersed and floating vegetation,
permitting it to flow downstream;
2. Harvesting systems which cut the vegetation, lift it from the
water, and place it on shore;
3. Dredges and pumps designed to remove aquatic plants, organic
detritus, and plant roots from the hydrosoil;
4. Draglines and backhoes.
Advantages of each of the systems are apparent. Most people, however,
prefer to harvest aquatic weeds in order to remove nutrients in the plants.
Studies in Florida have shown that mechanical harvesting of submersed vege-
tation removes an insignificant amount of nutrients and that damage to fish-
eries could result. Small sport and forage fish species live in submersed
vegetation and become entrapped in vegetation as it is harvested. Recent
studies have shown that 32% of a fish population can be removed by a single
mechanical harvesting operation (Haller et al. 1980). Generally, very little
mechanical harvesting or control of aquatic weeds is conducted in the South
due to high cost.
53
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BIOLOGICAL CONTROL
It is widely held that exotic weeds become problems in their new habi-
tats because they lack natural predators or controls which co-evolved with
them in their native habitat. Submersed weeds are spreading most rapidly in
the south and are by far the most expensive to control. Prospects for dev-
velopment of new, inexpensive herbicides or more efficient mechanical har-
vesting equipment within the immediate future are not great. Consequently,
more effective weed control programs in the field of biological control are
needed.
Discovery and successful widespread use of the alligator-weed flea
beetle (Agasicles hygrophila Selman and Vogt) for control of alligator-weed
in the 1960s have provided impetus to study biological control of other
aquatic plants (Maddox et al. 1971; Spencer and Coulson 1976). Three insects
have been studied and released on water hyacinths in the United States, and a
host-specific fungal pathogen (Cercospora rodmanii Conway) is currently being
developed (Freeman 1977; Charudattan 1979).
There are hundreds of thousands of insect species in the world, but
relatively few live in the submersed aquatic habitat where hydrilla thrives.
The search for insect biocontrols for hydrilla has not yet yielded signifi-
cant insect candidates (Zeiger, personal communication). Recently, a fungal
pathogen (Fusarium culmorum) was isolated from plants from Holland and is a
promising organism for hydrilla control. Extensive field studies were plan-
ned to begin in 1979 (Charudattan 1979).
Biological control studies of hydrilla with snails (Marisa cornuaretis
L.) have been extensively studied in Florida. It was found that Marisa was
not a significant biological control because it was temperature sensitive,
and very high stocking densities were required (Blackburn et al. 1971).
Several species of fish have been considered as candidates for the bio-
logical control of submersed aquatic weeds (Blackburn et al. 1971; Legner et
al. 1975). The Chinese Grass Carp or White Amur (Ctenopharyngodon idella
Val.) recieved most attention for aquatic weed control in the South.
The grass carp was introduced into the United States in the 1960s and
was first evaluated for aquatic weed control in Arkansas and Alabama. Arkan-
sas has been the leading state in the use of carp for weed control: its major
waters have been stocked with excellent control of many aquatic weed species.
In other states, fishery biologists and environmental groups have proceeded
with caution, and the grass carp has been banned from many southern states.
The first field research in Florida was initiated in three small (0.08
ha) earthen ponds in central Florida in 1971. This nonreplicated study
(complete in January 1973) showed that stocking rates of 50 grass carp/ha
controlled hydrilla without catastrophic effect on the aquatic environment
(Haller and Sutton 1976).
54
-------
Further major research was undertaken by the Florida Department of Nat-
ural Resources and the Florida Game and Fresh Water Fish Commission in 1972.
Four natural ponds in widely scattered geographical locations were stocked
with grass carp for collection of baseline data after one year. The diverse
interpretation of the results of these nonreplicated studies has become
widely known among the world's fishery scientists (Beach et al. 1976; Gasaway
and Orda 1978).
In 1974, six lakes (each more than 50 ha) and one reservoir (2000 ha)
were stocked with grass carp to further determine their weed control capabil-
ities and potential environmental impact. This research is incomplete as
three of the lakes have become weed-free, and three other lakes still contain
hydrilla infestations. In one lake, apparent lack of biocontrol resulted
from low grass carp populations (Colle et al. 1978). Restocking programs
have begun on the remaining vegetated lakes and the reservoir.
Due to unpredictable results obtained with grass carp, and unanswered
questions concerning its possible impact on sport fish populations, there
remains considerable controversy among biologists with regard to widespread
use of the fish in hydrilla control programs.
Currently, the State of Florida allows private possession of grass carp
by individuals with weed problems in lakes (10 ha or less) that are not con-
nected to other water bodies. Stocking is permitted only in private waters
that meet specific criteria (size, weed problems, and lack of infall or out-
fall) and the grass carp is currently in use in golf course ponds, fishery
ponds, and waters of similar nature.
Widespread application of the grass carp to solve hydrilla problems in
large lakes has been deferred until further studies are conducted. The
problem remains: hydrilla continues to spread and current control measures
are expensive.
IMPACT OF CONTROL METHODS
Nutrient Reduction Techniques
1. Reduction of Plant Nutrient Inputs. Studies on natural lakes
(Edmondson 1961, Sakamoto 1966; Vollenweider 1968, 1969; Shannon and
Brezonik 1972; Schindler et al. 1973; Dillon and Rigler 1974a, 1974b; Jones
1974; Kirchner and Dillon 1975; Jones and Bachmann 1976; and others) have
shown that input of plant nutrients is an important determinant of lake
nutrient concentrations and plant biomass, particularly phytoplankton bio-
mass. For this reason, it has been assumed that reduction in nutrient inputs
would reduce excessive growth of phytoplankton and aquatic macrophytes.
Edmondson (1966, 1969, 1970, 1972a), Michalski and Conroy (1973), and
Schindler (1975) have shown that lakes respond to nutrient reductions: their
studies have shown significant reductions in nitrogen and phosphorus concen-
trations. The biomass of phytoplankton as measured by chlorophyll £i concen-
trations was also reduced. In general, there was an overall improvement in
water quality.
55
-------
Although some lakes have responded to nutrient reduction programs almost
immediately, lakes with a long history of high-nutrient loading rates have
often failed to respond or have recovered very slowly (Alhgren 1972; Bjork
1972; Larsen et al. 1975). Schindler (1976) suggested that this difference
in response time is related to the degree to which bottom sediments are sat-
urated with nutrients. He suggested that the sediments can release nutri-
ents into the water for long periods of time, thus delaying recovery of lakes
that have recieved long-term, high-nutrient loading. Studies by Lamarra
(1975a, 1975b), Gasith et al. (1976), and Lie (1978) have shown that bottom-
feeding fish and rooted aquatic macrophytes can also enhance recycling of
nutrients from bottom sediments.
2. Reduction of Internal Nutrient Recycling. Bottom sediments of an
aquatic environment represents a potentally significant plant nutrient
source, and sediment removal by dredging is often advocated. Bengtsson et
al. (1975) reported that removal of sediments from Lake Trummen, Sweden,
minimized nutrient recycling. They noted that phosphorus, nitrogen, silica,
and phytoplankton concentrations declined, while phytoplankton diversity and
water clarity increased.
Dredging to reduce nutrients does not always result in water quality
improvement. In fact^ water quality can deteriorate significantly during
dredging operations. Peterson (1979) reviewed the literature on the impact
of dredging on the aquatic environment and noted that dredging caused re-
lease of toxic substances and nutrients from sediments, increases in water
turbidity, depletions of oxygen, and changes in pH and temperature. Many of
these effects, however, last only a short time.
Another method of reducing nutrient recycling and nutrient concentra-
tions in the aquatic environment is the application of chemicals that
precipitate nutrients. Aluminum, iron, calcium, zirconium, and lanthanum
(Gahler 1969) have been used to remove nutrients from water. Funk and
Gibbons (1979) recently reviewed some of the literature concerning effects of
nutrient inactivation techniques on the aquatic environment. They noted that
reductions in total phosphorus, ammonia, and Kjedldahl nitrogen and iron
concentrations, as well as algal standing crops, have occurred following
chemical treatment. Improvements in water clarity and hypolimnetic oxygen
concentrations have also been reported. However, Funk and Gibbons (1979)
noted that these effects were not always consistent among lakes. They cited
a Cooke and Kennedy (1977) report that chemical treatments were only
partially successful in two of the lakes observed in their study. Cooke and
Kennedy (1977) suggested that macrophytes and bottom fauna mediated release
of nutrients from sediments, thus reducing the effectiveness of nutrient
precipitation. Funk and Gibbons (1979) also noted that there were very
little data concerning effects of chemical treatments on fauna of the lakes.
Artifical aeration has often been used to prevent regeneration of plant
nutrients from sediments during anaerobic conditions. It was thought that
aeration could prevent development of algal growth by preventing regeneration
of nutrients from the sediments. Fast (1979) recently reviewed the litera-
ture on aeration and found that nutrient concentrations, particularly that of
phosphorus in deep water, nearly always declined during destratification of
, 56
-------
lakes with aerators. Studies by Robinson et al. (1969) and Haynes (1971)
have shown that destratification of lakes results in reduction of algal
concentration. However, there is also evidence that destratification can
result in increased algal numbers, particularly green algae (Hooper et al.
1952; Robinson et al., 1969; Haynes 1971; Ridley 1971). Fast (1979) noted
that circulation of lake water could raise the temperature of deep-water
sediments. With increased temperature and oxygen, Fast suggested these
sediments could be colonized by benthic organisms and that their activities
(Lee 1970; Brinkhurst 1972; Davis 1974) could increase nutrient exchange
rates. Fast (1979) noted that there is no clear evidence that aeration
reduces nutrient concentration within the euphotic zone of lakes. In fact,
he showed that there is evidence that destratification upwells nutrients to
surface water.
Aeration of lakes does have important effects on other components of the
aquatic environment. Aeration of anaerobic waters reduces the concentrations
of iron, maganese, nitrogen, and sulfur (Irwin et al 1967; Wirth and Dunst
1967; Symons et al. 1970; Haynes 1971). In addition, aeration can increase
the oxidation rate of organic matter in bottom sediments and the water column
(Mercier 1955; Fast 1971). Fast (1971) found that aeration resulted in rapid
invasion of deep-water sediments by benthic organisms and an increase in the
vertical distribution of zooplankton. Aeration of waters can also increase
depth distribution of fish (Gebhart and Summerfelt 1975; Brynildson and Serns
1977). Despite all the beneficial effects noted for aeration, Fast (1979)
reported that not all have been thoroughly documented, and in some lakes,
beneficial effects noted in other lakes do not always occur.
Sediment exposure and desiccation has been suggested as a method for
reducing internal recycling of nutrients because oxidation of the sediment
surface should retard release of nutrients (Mortimer 1941; 1942) and increase
the binding capacity of sediments (Fitzgerald 1970). Sediment exposure can
also curb sediment nutrient release by physically stablizing the upper zone
of flocceluent sediments (Lee 1970). However, other studies (Sneisko 1941;
Neess 1946; Davis and Lucas 1959) suggest sediment desiccation will accel-
erate microbial mineralization of organic matter, thus making inorganic nu-
trients available for plant growth upon reflooding. Drawdown and reflooding
are used widely to increase plant growth in marshes and fish culture ponds.
Effects of Herbicides on the Aquatic Environment
For this report, we will concentrate on the five most widely used aquat-
ic herbicides and their various formulations. Federal Laws, particularly the
Federal Insecticide, Fungicide, and Rodenticide Act (FIFRA), have greatly re-
duced the number of herbicides used in aquatic ecosystems. All chemicals
currently used for aquatic weed control must have either federal registration
(full or interim) and/or must be used on a large enough basis to warrant
state special need registration.
57
-------
Because of these restrictions, the predominant herbicides are those sold
in quantities great enough to offset the expense (toxicology, residue, meta-
bolism, and ecological studies) required for registration. Consequently,
various formulations of copper, endothall, Dalapon, 2,4-D, and Diquat consti-
tute the vast majority of aquatic herbicides used in the United States
(Haller 1979).
It should also be stated that the products mentioned in this section
have full registration, or at least have interim tolerances established for
water and fishes. These compounds are, as far as is known, safe to use in
the aquatic environment. Nevertheless, effects of these compounds on water
quality, non-target organisms, and other selected parameters will be
reviewed in this section.
2,4-D (2,4-dichlorophenoxy) Acetic Acid
2,4-D is the parent acid used to make many formulations (salts, amines,
esters, etc.). These formulations have differential herbicidal activity on
different species of weeds and also have different residue, toxicology, and
metabolite characteristics.
The dimethyl amine salt of 2,4-D is a wide-spectrum broadleaf weed kill-
er. It is not effective on grass species (monocots). It is most widely used
in water hyacinth control (2 to 4 Ib/acre) and for control of ditchbank
dicots (6 to 10 Ib/acre) (brush, etc.). The butoxy-ethanol ester of 2,4-D,
(2,4-D BEE), often formulated on granules, is used for control of waterlily
species (Nuphar, Nymphaea) and Eurasian watermilfoil. Application rates vary
between 10 and 40 Ib active ingredient per acre, depending upon species.
Residues in Water and Soil. The half-life of 2,4-D in water varies with
environmental conditions and habitat sprayed. In general, high water temper-
atures, dissolved oxygen, dissolved organic matter, etc., cause rapid degrad-
ation for all herbicides.
2,4-D dimethyl amine sprayed on foliage of water hyacinths or on ditch-
bank weeds results in very little 2,4-D contamination of water. It has been
shown that of 4 Ib 2,4-D sprayed on an acre of water hyacinth, only 20 to
40% of the chemical is detected in the water, presumably the result of spray
runoff from plant leaves. Thus, at the rate of 4 Ib/acre, between 1 and 2 Ib
might be detected in the water, but the concentration would vary with depth.
In a slow-flowing situation (e.g. canal), it is possible that 2,4-D in water
would be undetectable due to the chemical being applied adjacent to water but
not over it.
Residues of 2,4-D dimethyl amine applied at normal rates of 2 to 4
Ib/acre dissipate from water to non-detectable levels in 10 to 20 days. Rate
of dis- sipation or degradation of 2,4-D from hydrosoil is similar to water
and varies depending upon environmental conditions (Tables 11 and 12).
Breakdown of 2,4-D in aquatic systems is apparently due primarily to
microbial metabolism (Frank and Comes 1967; Schultz 1973).
58
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Table 11.
Residues of the dimethylamine salt of 2.4-D in water (mg/1), hydrosoil(mg/kg)
and fish (mg/ kg) from ponds in Florida and Georgia treated with 2.24,4.48 and
8.96 kg 2,4-D per hectare. From Schultz and Gangstad, 1976.
Pond
Florida
Georgia
Florida
Georgia
Florida
Georgia
Florida
Georgia
Florida
Georgia
Rate
kg/ha
2.24
4.48
8.96
2.24
4.48
8.96
2.24
4.48
8.96
2.24
4.48
8.96
2.24
448
8.96
2.24
448
896
224
448
8.96
2.24
4.48
8.96
2.24
448
8.96
2.24
4.48
8.96
Depth
m
1.3
1.0
1.2
1.3
0.9
1.2
1 3
1 0
1 2
1.3
0.9
1.0
1.3
0.9
1 2
1 3
0.9
1.0
1.3
1.0
1.2
1.3
0.9
1.0
1.3
1.0
1.2
1.3
0.9
1.0
Temp.
C
34
31
31
27
29
30
30
30
31
29
32
30
31
31
31
26
27
29
32
32
31
28
31
30
30
32
30
27
31
30
Time
Days
01
01
01
01
01
01
03
03
03
03
03
03
07
07
07
07
07
07
14
14
14
14
14
14
28
28
28
28
28
28
Water
mg/l
0.025
0.155
0.312
0.025
0.233
0657
0.005
0.172
0345
0.087
0.390
0.692
0.005
0.048
0.025
0.059
0.400
0.395
0.005
0.005
0.005
0.027
0.008
0.050
0.005
0.005
0.005
0.005
0.005
0.005
Hydro-
soil
mg/kg
0.005
0.014
0.033
0.018
0.024
0.026
0.005
0014
0046
0.008
0018
0040
0.005
0010
0.008
0010
0.018
0042
0.005
0.010
0.013
0.005
0.005
0.005
0005
0.007
0.005
0.006
0.005
0.005
Fish
mg/kg
0.080
0.048
0.005
0.005
0.014
0.022
0.005
0.005
0.005
0005
0.005
0.005
0.005
0005
0.005
0.005
0.005
0005
0036
0.005
0043
0.005
0.005
0.005
0.005
0005
0.005
0005
0005
0.010
A literature review concerning herbicide residues or toxicology is dif-
ficult due to the tremendous variation in data from small-pond studies, lab-
oratory studies, and large-scale studies done under non-standardized condit-
ions. Residue data vary from 1 day to 6 months after treatment, and authors
often do not list pertinent environmental conditions in their reports.
Large scale field tests are probably more useful in determining the ef-
fect of herbicides on the aquatic environment. Joyce and Sikka (1977) pre-
sented data on the use of 2,4-D dimethyl amine in a major hyacinth control
program.
The St. Johns River is a slow moving Florida river approximately 300 miles
long with a long-standing water hyacinth problem. Joyce and Sikka (1977)
collected water samples at nine locations in the river and compared the amount
of 2,4-D applied to its residue in each location (Figues 28 and 29). Several
hundred pounds of 2,4-D were applied to shoreline marshes and backwaters of
the river over 6 months. The maxium concentration of 2,4-D detected in the
river was 1.4 parts per billion (ppb).
59
-------
W-9
3HWAY-192 (262)
W-8
3HWAY-520(232)
W-7
3HWAY-50 (209)
W-6
GHWAY-46 (190)
W-5
DELANO (145)
WC -4
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WC-3
PALATKA (80)
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12621
W-8
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(232)
W-7
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1209)
W-6
HIGHWAY- 46
(190)
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(145)
WC-4
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(100)
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(80)
WC 2
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SPRINGS 152)
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(17)
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-------
Table 12.
Physical, chemical and toxicological properties of some aquatic herbicides. From Newbold, 1975.
Herbicide
Dalapon
2.4-D
amine salt
Oiuron
Asulam
Copper
sulphate
Maleic
hydrazide
Maleic
hydrazide/
2.4-D
Maleic
hydrazide/
2.4-D/
chlorpropham
Chemical formula
2.2-dichloroproprionic
acid
2,4-dichlorophenoxy-
acetic acid (amine salt)
3-(3.4-dichlorophenyl)-
1,1-dimethylurea
methyi(4-aminobenzene-
sulphonyl)carbamate
CuSCv 5H20
6-hydroxy-3-(2H)-
pyridazinone
maleic hydrazide/2,4-
D/N-(3-chlorophenyl)
carbamate
Mode of
action
contact/
translocation
translocation
translocation
inhibits
photosynthesis
translocation,
inhibits cell
division
Translocation
suppresses
growth,
inhibits cell
division
mitotic poision
Persis-
tence
in muds
10to 60
days
1 month
4 months
Not
known
Not
known
Soils
1 to 8
weeks
Break-
down
in water
2 to 3
days
4 to 6
weeks
3 months
2 weeks
Not
known
Not
known
Treat-
ment
level
Form (mg/l)
Liquid 1.0
Liquid
Liquid 0 25-
0.5
Liquid
Crystal 0.2-
1.0
Liquid 1.0
Levels of toxicity
to fish
LCM
350 mg/l, 24 h Salmo
gairdnerii (Richardson)
(Alabaster, 1969)
250 mg/l. 24 h Salmo
gairdnerii (Richardson)
(Alabaster, 1969)
4.03 mg/l, 48 h Salmo
gairdnerii (Richardson)
(F.WP.C.A.. 1968)
Levels of toxicity
to Daphnia
LCM
6-0 mg/l, 48 hDaphnia
magna (Straus)
(F W.PC.A.. 1968)
> 100 mg/l, 48 h
Daphnia magna
(Straus) (Crosby &
Tucker, 1966)
1 -4 mg/l, 48 hDaphnia
pa/ex (De Geer)
(F.W.P.C.A., 1968)
5200 mg/l, 24 h RasboraNol known
heteromorpha (Duncker)
(Alabaster, 1969)
0.14 mg/l, 48 h Salmo
gairdnerii (Richardson)
(Shaw & Brown, 1974)
Information insufficient
Not known
Not known, likely to be
very toxic
(Greulach et a/.. 1961)
A large-scale submersed weed control operation was conducted in 1966 in
TVA lakes where dramatic increases in Eurasian watermilfoil hindered naviga-
tion and power production. Propogation of Anopheles mosquitoes necessitated
use of larvicides to control mosquitoes. A report (Anonymous 1961) stated
that "watermilfoil, like some other submersed aquatics, offers an extremely
favorable late season habitat for Anopheles quadrimaculatus and other perma-
nent pool mosquitoes."
Smith and Ison (1967) reported that 888 tons of 20% 2,4-D BEE granules
were applied to 8,000 acres of milfoil in seven TVA reservoirs. The treat-
ments spanned a 352-mile main channel distance and were applied between March
and December 1966.
Fish mortalities, differences in pre- and post-treatment populations of
Hexgenia, and toxic effects on benthic fauna did not occur. Smith and Isom
(1967) concluded that high application rates (up to 100 Ib/acre in flowing
water) of 2,4-D BEE did not produce adverse effects on aquatic fauna or water
quality (Tables 13, 14, and 15).
62
-------
Table 13.
Analyses of water samples. Watts Bar Reservoir. Gordon Branch Embayment. From Smith and Isom, 1967.
Alkalinity 2.4-D
Station
Date
Time
(FS)
Temp
Dissolved
Oxygen
(Mg/l)
(Surface samples collected prior
A
B
C
D
E
A
B
C
D
E
A
B
C
D
E
A
B
C
D
E
3-17-66
3-17-66
3-17-66
3-17-66
3-17-66
3-18-66
3-19-66
3-19-66
3-19-66
3-19-66
3-18-66
3-19-66
3-19-66
3-19-66
3-19-66
3-18-66
3-19-66
3-19-66
3-19-66
0505
0825
0835
0840
0855
(Surface
0920
0945
0955
1000
1020
(Surface
1220
1230
1240
1245
1305
(Surface
1550
1510
1515
1520
55.7
56.3
56.7
56.2
56.4
samples
55.4
56.1
37.2
53.0
56.1
samples
59.0
61.6
59.3
62.4
56.8
samples
64.4
64.7
62.2
67.2
10.0
10.2
11.4
10.9
11.4
collected 1 hour
11.7
10.4
11.4
8.8
1 1.4
collected 4 hours
12.8
10.8
11.5
12.6
10.4
collected 8 hours
11.6
10.4
11.4
10.4
pH
Phenol
(Mg/l)
Total
(Mg/l)
BEE Acid
U/g/l) (/;g/l as BEE)
to 2,4-D application)
8.1
8.6
7.8
8.1
8.2
after
8.8
7.8
8.3
6.7
8.4
after
9.1
7.9
8.5
8.6
8.5
after
8.7
8.2
8.5
8.6
0.0
1.4
0.0
0.0
0.0
70.1
35.9
33.4
34.9
34.5
2,4-D application)
6.0
0.0
0.0
0.0
1.0
94.0
35.0
34.0
19.5
32.5
<1 <1.45
<1 <1.45
<1 <1.45
37 <1.45
6 <1.45
2,4-D application)
18.0
0.0
2.0
3.0
1.5
92.0
35.0
36.0
36.0
35.0
<1 <1.45
6 <1.45
<1 <1.45
<1 <1.45
2 <1.45
2.4-D application)
5.0
0.0
2.0
3.0
91.0
36.0
38.0
35.0
<1 <1.45
<1 <1.45
<1 <1.45
<1 <1.45
(Sample missing)
Diquat [6,7-dihydrodipyrido (l,2-a:2',l'-c)pyra2inediium ion]
Diquat is probably the most widely used herbicide for submersed aquatic
weed control in the United States. It is effective at a rate of 2 ppm or
less on hydrilla, Elodea, milfoil, pondweeds, algae, and various other
species. There is only one formulation of diquat, the liquid diquat
dibromide.
The chemical is a contact weed killer, phytotoxic to plant tissue, and
kills by interfering with photosynthesis. In aquatic systems it is used
largely for control or submersed plants, but also is used in some instances
on duckweed, water hyacinth, water lettuce, or ditchbank grasses.
This discussion will emphasize use of diquat on submersed plants.
63
-------
Table 14.
Analyses of water samples, Guntersville Reservoir, Vicinity of Comer Bridge. From Smith and Isom, 1967.
Alkalinity
Station
Time
Date (CS)
Temp.
(°F)
Dissolved
Oxygen
(Mg/l) pH
(Surface samples collected prior to 2,
A
B
C
D
E
F
A
B
C
D
E
F
A
B
C
D
E
F
A
B
C
D
E
F
3-29-65 1120
1110
1101
1215
(Surface
4-5-66 0835
0845
0855
0925
0910
0905
(Surface
4-5-66
1220
1230
1240
(Surface
4-5-66 1455
1505
1515
1535
1530
1525
57.0
57.0
57.0
60.0
59.0
samples
56.5
56.0
56.3
56.0
56.0
56.0
samples
57.0
58.0
57.5
9.2
8.8
collected 1
8.8
8.7
8.6
9.0
8.6
8.6
collected 4
9.2
9.3
9.4
samples collected 8
56.5
57.0
57.0
58.0
58.0
58.0
8.6
8.5
8.5
9.7
10.1
9.9
7.7
7.5
hour after
7.2
7.6
7.4
7.6
7.6
7.7
hours after
7.6
7.6
7.8
hours after
7.4
7.5
7.5
7.8
8.0
7.9
Phenol
(Mg/l)
Total
(Mg/l)
2,
4-D
BEE Acid
(ug/\) (fjg/\ as BEE)
4-D application)
0.0
0.0
37.0
39.4
3
<0.5
7
7
6
7
<1.45
<1.45
<1.45
<1.45
<1.45
<1.45
2,4-D application)
0.0
0.0
0.0
0.0
0.0
0.0
36.0
41.0
41.0
36.0
38.0
39.0
91
157
36
5
5
3
<1.45
<1.45
<1.45
<1.45
<1.45
<1.45
2,4-D application)
0.0
0.0
0.0
36.0
37.0
36.0
8
130
19
<1.45
<1.45
<1.45
2,4-D application)
0.0
0.0
0.0
0.0
0.0
0.0
39.0
39.0
39.0
38.0
36.0
38.0
2
<1
4
18
64
21
<1.45
<1.45
<1.45
<1.45
<1.45
<1.45
64
-------
Table 15.
2.4-D analyses Watts Bar and Guntersville Reservoirs. From Smith and Isom. 1967
Sampler
No
19
25
18
16
21
26
27
22
23
15
8
32
34
33
6
10
48
49
50
47
43
44
45
46
Date
Collected
Prestudy
3-20-66
3-20-66
3-20-66
3-22-66
3-23-66
3-23-66
3-23-66
3-23-66
3-23-66
3-23-66
4-13-66
4-13-66
5-24-66
5-25-66
5-26-66
5-25-66
5-25-66
5-25-66
1-17-67
1-17-67
1-17-67
1-17-67
1-17-67
Hours 'Days
After
Treatment
Control
24 hours
24 hours
24 hours
72 hours
96 hours
96 hours
96 hours
96 hours
96 hours
96 hours
24 days
24 days
35 days
50 days
50 days
50 days
50 days
50 days
10 months
10 months
10 months
10 months
10 months
Mg/T
or Mg/Kg
Material BEE
Watts Bar
Fish
Watermilfoil
Watermilfoil
Watermilfoil
Fish
Mud
Mussel
Mud
Mud
Fish
Mussel
Mud
Mud
Mud
Fish
Fish
Fish
Fish
Fish
Fish
Mud
Mud
Mud
Mud
Reservoir
<0 14
<0.14
826
336
<0.14
560
038
2.8
095
<0 14
070
350
015
0 14
<0 14
<0 14-
<0 14
<0 14
015
<0 14
024
091
028
588
Station3
Control
Gordon Branch
do
do
do
do.
do
do.
do
do
do
do
do
do
do
do
do
do
do
do
do
do.
do.
do
Species
Lepomis macrochirus
Myriophyllum spicatum
Mynophyllum spicatum
Myriophyllum spicatum
Lepomis macrochirus
Assorted mussels
Lepomis macrochirus
Elhptio crass/dens
Lepomis macrochirus
Lepomis macrochirus
Ictalurus punctatus
Silzostedion canadense
Lepomis macrochirus
Pomolobus chrysochlons
Guntersville Reservoir
17
2
5
1 1
13
14
30
12
1
40
41
4
24
39
28
20
7
29
31
3
3
35
36
37
38
42
Prestudy
Prestudy
May 1966
4-06-66
4-06-66
4-06-66
4-06-66
4-06-66
4-08-66
4-08-66
4-08-66
4-1 1-66
4-1 1-66
4-21-66
4-11-66
4-11-66
5-17-66
5-17-66
5-17-66
5-17-66
5-17-66
1-20-67
1-20-67
1-20-67
1-20-67
1-20-67
Control
Control
Control
24 hours
24 hours
24 hours
24 hours
72 hours
72 hours
72 hours
72 hours
144 hours
144 hours
15 days
144 hours
144 hours
42 days
42 days
42 days
42 days
42 days
9 months
9 months
9 months
9 months
9 months
Mussel
Asiatic Clams
Mud
Mussel
Mussel
Mussel
Mussel
Mussel
Mussel
Mussel
Mussel
Mussel
Mussel
Fish
Mussel
Fish
Mud
Mud
Mud
Mussel
Mussel
Fish
Mud
Mud
Mud
Mud
<0 14
<0 14
0 14
025
024
<0 14
1 12
0 18
030
098
1 0
<0 14
<0 14
<0 14
<0 14
<0 14
<0 14
33 6
0.14
<0.14
020
<0.14
0.34
0.30
049
0.30
Control
la-1
"Control"
Out-1
ln-3
"Control"
ln-1
ln-3
ln-2
ln-1
ln-3
ln-2
Out-3
Out-1
Out-1
ln-3
Out-2
ln-1
ln-2
ln-2
ln-2
Out-2
ln-3
"Control"
Elliptic crassidens
Corbicula mamllensis
Elliptio crassidens
Elliptic crassidens
Elliptio crassidens
Elliptic crassidens
Elliptic crassidens
Elliptic crassidens
Elliptic crassidens
Elliptic crassidens
Elliptic crassidens
El/iptio crassidens
Ictalurus lurcatus
Elliptic crassidens
Lepomis macrochirus
Elliptic crassidens
Elliptio crassidens
Dorosoma cepedianum
' The C.W. England Laboratories converted 2.4-D and its esters to the methyl ester for reporting to TVA, however, for
comparison with published data on toxicity (2), all data were converted to the BEE equivalent.
2.4.D 2,4-dichlorophenoxyacetic acid
butoxyethanol ester
methyl ester
2 In = Inside embayment
Out = Outside in river channel external to embayment
Note The station labelled "Control" received an unplanned application of 2.4-D and. therefore, cannot be
a control
65
-------
Table 16.
Physical, chemical and toxicological properties of some aquatic herbicides. From Newbold, 1975.
Herbicide
Diquat
Paraquat
Dichlo-
benil
Chlor-
thiamid
Terbutryne
Chemical formula
l,1'-ethylene-2.2'.
dipyridylium dibromide
1.1'-dimethyl-4,4'-.
diphndyhum dichloride
2,6-dichlorobenzo-
nitrile
2.6-dichlorothio-
benzamide
2-rerr-butylamino-4-
ethylamino-6-methyl-
thio-1,3,5-tnazine
Mode of
action
contact
translocation
contact
translocation'
translocation
translocation
disrupts
photo-
synthesis
Persistence
in muds
6 months-
1 year
>2 years
6 months
6 months
Not
known.
likely to
be very
persistent
Break-
down Form
in water
8 to 11 Liquid
days
7 to 14 . Liquid
days
2 to 3 Granule
months
2 to 3 Granule
months
>3 Granule
months
Treat-
ment
level
(mg/l)
1.0
1.0
1.0-
3.0
1.0-
3.0
0025-
0.1
Levels of toxicity
to fish
LC,o
90 mg/l, 24 h Salmo
gairdnerii (Richardson)
(Alabaster, 1969)
840 mg/l, 24 h Rasbora
heteromorpha (Duncker)
(Alabaster, 1969)
1.6 mg/l, 10 days Rutilus
rutilus (L)
(Tooby, 1972)
41 mg I. 24 h Ras Bora
heteromorpha (Duncker)
(Alabaster, 1969)
3.5 mg/l, 96 h Salmo
gairdnerii (Richardson)
(Tyson, 1974)
Levels of toxicity
to Daphnia
LCso
7 Img/l, 24 h Daphnia
magna (Straus) (Crosby
& Tucker. 1966)
3.7 mg/l, 48 h Daphnia
pulex (De Geer)
(Sanders & Cope, 1966)
3.7 mg/l. 48 h Daphnia
pulex (Oe Geer)
(F.W.P.C.A.. 1968)
1 .4 mg/l, 48 h Daphnia
magna (Straus)
(Tyson, 1974)
Residues in Water and Soil. Diquat is inactivated by contact with soil
particles, so it cannot be used in turbid conditions. Newbold (1975) report-
ed complete degradation of 1.0 ppm in water in 8 to 11 days. Persistence in
bottom muds lasts from six months to one year (Table 16). Once the material
is tightly bound to hydrosoil it is, in effect, out of the system and remains
bound to soil particles. In this condition, it cannot be re-released into
the water column nor can it be absorbed and re-enter the ecosystem via plant
uptake. Diquat is degraded in hydrosoil microbially.
Mackenzie (1969) reported three residue trials conducted in southern
Florida in 1966. Diquat was applied to Florida elodea (hydrilla) at the rate
of 0.5 ppm. Water for residue analysis was collected at the surface, mid-
depth and bottom, and plant hydrosoil samples were collected 1 day, 8 days,
2 weeks, and 3 weeks after treatment.
Data in Tables 17, 18, and 19 indicate that except for Trial 1, diquat
residues were not found 14 days after treatment. Apparently, submersed
aquatic weeds rapidly absorb diquat scon after treatment. Rapid uptake by
plants is evidenced; water residue values averaged 0.22 ppm 1 day after
treatment with 0.5 ppm diquat. Laboratory studies have further demonstrated
rapid absorption of diquat by hydrilla (Sutton et al. 1972).
66
-------
Table 17.
Details of residue trials 1, 2. and 3, Dade County, Florida, 1966. From Mackenzie, 1969.
Residue
Trial Canal Dimensions
number Length
feet
1 2120
2 1500
3 1720
Width
feet
85
45
45
Depth
feet
13
6
12
Density of
Elodea'
30
75
100
Location
Andover "A"
Carol City
"A8"
Hefler
Homes "N"
Previous
Diquat History
No diquat
No diquat
Treated two times
in 1965 without
success
'Expressed as percentage of canal volume occupied by elodea.
Table 18.
Water samples diquat residues as related to time after treatment and depth in
residue trials 1, 2 and 3. From Mackenzie, 1969.
Sample level
in water
Residue
profile Trial No.
Top foot
of water
Mid section
of profile'
Bottom foot
of water
1
2
3
1
2
3
1
2
3
1 day
ppmw
0.28
0.20
0.24
0.09
026
0.45
0.27
0.22
0.01
Time After
8 days
ppmw
0.10
0.04
0.00
0.10
0.03
0.00
0.11
0.08
0.01
Treatment
2 weeks
ppmw
0.06
0.00
0.00
0.09
0.00
0.00
0.07
0.00
0.00
3 weeks
ppmw
000
0.00
0.00
0.00
0.00
0.00
000
000
0.00
'1 and 36 ft. below surface
23 ft. feet below surface.
Table 19.
Diquat residues in elodea based on dry weight as related to time after treatment
in residue trials 1, 2 and 3. From Mackenzie, 1969.
Time after Treatment
Sample site
1 day
8 days 2 weeks 3 weeks
ppmw
Residue
Residue
Residue
Residue
by flow
Trial
Trial
Trial
Trial
from
1
2
3
3
treated area
treated area
- treated area
- area affected
9.
6.
0
08
58
.17
ppmw
11
9
0
55
23
.06
ppmw
Elodea
eradicated
663
0.04
0.29
ppmw
Elodea
eradicated
4.90
0.04
treated area
67
-------
Mackenzie's (1969) field residue trials also determined diquat content
of submersed weeds (Table 20) on a mg diquat/kg dry weight basis. One day
after treatment, hydrilla contained 9.08 ppm diquat in Trial 1 and 6.58 ppm
in Trial 2, but in Trial 3, which was affected by water flow, the plants only
contained 0.17 ppm diquat 24 hours after treatment.
Table 20.
Bottom soil samples - diquat residues based on dry weight as related to time
after treatment in residue trials 1, 2 and 3. From Mackenzie, 1969.
Residue
trial
number
1
2
3
Time after treatment
1 day
ppmw
0.33
0.85
0.00
8 days
ppmw
3.52
3.68
0.00
2 weeks
ppmw
6.84
2.60
1.74
3 weeks
ppmw
3.98
3.14
0.00
Bottom mud samples indicated that diquat persisted in hydrosoil longer
than in any other portion of the ecosystem. The data are variable and addi-
tional samples should have been taken, but Mackenzie's hydrosoil residue data
suggest maximum diquat accumulation in hydrosoil 2 weeks after treatment
(mean of his trials was 3.71 ppm).
Interestingly, this was the same length of time generally required for
submersed weeds to disintegrate and portions of them to sink to the hydro-
soil. Visual evaluations of the residue plots (Table 21) 1 and 2 weeks after
treatment showed that the greatest increase in plant control was obtained
during this time. This suggests that submersed weeds, soon after treatment
and in the process of sinking, carry diquat to the hydrosoil.
Table 21.
Florida elodea control results in residue trials 1, 2 and 3 treated with 0.5 ppmw diquat, 1 966.
From Mackenzie, 1969.
Deposits on Elodea
Residue trial Infestion
number of Elodea
%
1 30 marginal
2 75 - center open
3 100
Calcium
carbonate
None
Medium
Medium
Algae
Medium
Heavy
Heavy
1 wk.
%
75
70
35
2 wks.
%
99
85
65
Control
3 wks.
%
100
95
90
6 wks.
%
100
95
80
8 wks.
%
100
85'
702
'Retreatment with 0.5 ppmw diquat at 4 months gave 100% control for the remainder of 1966.
2Two retreatments at 0.5 ppmw diquat did not give satisfactory control in 1966.
68
-------
In summary, diquat applied at 0.5 to 1.0 ppm is widely used to control
submersed aquatic plants. Data indicate that plants nave absorbed more than
50% of the diquat within hours after treatment. Suspended soils, hydrosoil,
and breakdown by ultraviolet radiation all contribute to removal of the chem-
ical from the water (Simsiman et al. 1976).
Endothall [7-oxabicyclo(~.2.1)heptane-2t3-dicarboxylic acid)]
Endothall, as an organic acid, has the potential of being manufactured
in many salt, amihe, or other formulations. The formulations most widely
used in aquatic weed control are the potassium salt and the dimethyl-
-alkylamine salt of endothall.
Endothall is used primarily to control submersed aquatic weeds and is
applied at rates from 1 to 2 ppm, depending on the plant species to be con-
trolled and on formulation. It is injected either as liquid into water or
as sinking pellets applied to the surface.
Residues in Water and Soil. Hiltibran (1962) conducted some of the ear-
liest published residue research concerning endothall. Endothall was approv-
ed for aquatic weed control in 1960, and is one of the older, most exten-
sively studied organic aquatic herbicides.
Hiltibran's (1962) studies compared several combinations of tap water,
lake water, fish, mud, and plant debris placed in aquariums on degradation of
endothall. There were considerable differences in degradation among treat-
ments (Table 22). His data indicate that microcosms that included variations
of lake water, mud, plant debris, and fish have the most rapid degradation of
the disodium salt of endothall.
Further field studies were conducted in plastic enclosures placed in
farm ponds (Table 23) where there were less variation and more rapid degrad-
ation after application of 1 ppm endothall; none was detectable an average
of 36 hours after treatment. None was detectable 66 hours after application
of 5 ppm, and none was detectable 72 hours after application of 10 ppm.
A more recent study and review of endothall residues has been compiled
by Sikka and Rice (1973). These authors discussed variation in the rate of
degradation of the various salts of endothall, and particularly the dis-
appearance of endothall from water, accompanied by smaller but concurrent
increases in endothall concentrations in hydrosoil.
Dipotassium endothall (Aquathol^ K) was applied to a 0.1 acre non-
flowing pond (4 ft deep) at a rate of 2 ppm. Within the first 3 days after
treatment, the endothall concentration was decreased by 55% in the water
(Figure 30). As the endothall concentration in the water decreased between
days 4 and 22 after treatment, there was a slight increase in hydrosoil
endothall content. After 36 days, endothall could not be detected in the
water but persisted in the hydrosoil, becoming non-detectable 44 days after
treatment.
69
-------
Table 22.
Disappearance of endothall from laboratory aquaria containing various combinations of tap water, lake water,
4 to 6 live fish, plant debris, and mud. From Hiltibran, 1962.
Endothal
applied
ppm
1
1
1
1
1
1
1
1
1
1
5
5
5
5
5
5
5
5
5
5
5
5
10
10
10
10
10
10
10
Tank contents
Lake water,8 mud, fish
Lake water, fish
Tap water, fish, mud
Tap water, fish
Tap water, 1/2, lake water, 1/2; fish, mud
Lake water"
Lake water
Lake water, fish
Lake water, fish, mud
Lake water, fish, mud
Lake water
Lake water
Lake water, mud
Lake water, mud
Lake water
Lake water, fish
Lake water, fish
Tap water, mud
Tap water, 1/2, lake water, 1/2; fish, mud
Lake water, fish, mud
Lake water, plant debris
Tap water, fish
Lake water, plant debris
Lake water, mud
Lake water, plant debris, mud
Tap water, mud
Tap water, plant debris
Tap water, plant debris, mud
Lake water, plant debris
Hours to
reach
0.5 ppm
endothal
98
98
98
265
98
403
32
32
32
32
957
478
236
236
146
166
166
166
166
172
502
364
364
364
337
337
337
172
Days to
reach
0.1 ppm
endothal
8
8
8
21
8
61
40
13
13
9
5C
6C
6C
6C
6C
12
21
22
22
22
20
20
20
12
aLake water from Arrowhead Pond unless otherwise specified.
bLake water for this tank was from Allerton Lake.
cConcentration of endothal was 0.3 ppm instead of 0.1 ppm.
Table 23.
Disappearance of endothall from plastic-enclosed test plots of aquatic
vegetation in farm ponds. From Hiltibran, 1962.
Pond
Arrowhead
Hay's
Sage's
Hay's
Arrowhead
Hays'
Sage's
Hays
Arrowhead
Hays'
Sage's
Hays'
Plot size,
feet
35 x 8
15 x 15
20 x 5
10x 10
30 x 8
15 x 15
10x 10
10x 10
15 x 15
10 x 10
10 x 10
10x 10
Rate
ppm
1
1
1
1
5
5
5
5
10
10
10
10
Time since
application,
hours
2.0
0.5
0.5
0.5
24.0
24.0
48.0
48.0
24.0
168.0
48. 0
72.0
Cone. Hours until
ppm endothal was
not detected
1.0
1.7
1.0
1.0
0.8
1.7
0.8
0.8
0.8
1.0
' 1.0
1.0
24
24
48
48
48
48
72
96
48
72
96
70
-------
I 7fi -'
<
c=
2
O
O
Q
2
< 075-
0 50 -
025
8
40
^2 16 20 24 28 32
DAYS AFTER TREATMENT
Figure 30. Endothall residues in water and the top 1-inch of hydrosoil
of a treated farm pond, with time. The bars represent the
range of duplicate values. From Sikka and Rice, 1973.
In most studies, degradation of herbicide occurred most rapidly in the
field. However, aquaria studies by Sikka and Rice (1973) indicated more
rapid degradation in aquaria than in the pond study. One day after treatment
with 2 and 4 ppm dipotassium endothall, 80 and 86% of the residue, respect-
ively, had disappeared from the water (Figure 31). Endothall was not detect-
ed seven days after treatment, compared to 36 days in field studies. Bio-
degradation of endothall on hydrosoil also occurred more rapidly in aquaria.
Treatment with 2 ppm resulted in non-detectable hydrosoil endothall at day
22, and rates of 4 ppm were below detection at day 35 after treatment.
WATER -4 ppm
O O WATER 2 ppm
HYDROSOIL-4 ppm
o o HYDROSOIL -2 ppm
12 16 20 24
DAYS AFTER TREATMENT
28
32
36
Figure 31.
Endothall residues in water and hydrosoil of aquaria treated
with 2 and 4 ppm of the herbicide. The bars represent the
range of duplicate values. From Sikka and Rice, 1973.
71
-------
As with the other submersed aquatic weed herbicides, endothall is absor-
bed rapidly after application by plants, organic matter, and hydrosoil. Re-
sidues in water drop at least 50% after 24 hours and gradually decrease to
non-detectable between two and three weeks after treatment. Endothall
persists slightly longer in hydrosoil, but hydrosoil concentrations rarely
exceed an order of magnitude below treatment levels.
Copper Salts and Organic Copper Complexes
Copper sulfate (copper sulfate pentahydrate;CSP;CuSO4.5HoO) has been
used for over 100 years as a fungicide and algicide. Before the devel-
opment of organic herbicides in the 1940s, CSP and other mineral salts were
applied to soil as soil sterilants and were sprayed on foliage of floating
aquatic plants. Copper compounds were, however, most effective in the
aquatic environment as algicides.
Copper is still widely used as an algicide, but this discussion will be
limited to discussion of copper for control of submersed vascular aquatic
plants. When metallic copper (usually CSP) was used to control vascular
submersed aquatic weeds, such as milfoil, Elodea, and hydrilla, it was soon
discovered that concentrations required to control aquatic plants (4-8 ppm)
were very close to the threshold of fish toxicity. Consequently, fish kills
often occurred when CSP was used for aquatic weed control.
In the mid 1960s, chelated organic copper complexes were developed and
greatly reduced the toxicity and corrosiveness of copper, principally by
formulating the sulfate (SO4~) radical out of the mixture (Table 24).
Concurrently, research was conducted with combinations of diqut and copper
compounds. During these studies (Mackenzie and Hall 1967; Sutton et al.
1972; and Haller and Sutton 1973), it was discovered that copper-diquat and
copper-endothall mixtures had a synergistic effect on aquatic weeds, part-
icularly hydrilla. The use of 4 Ib diquat and 2 to 4 Ib copper per acre
provided good (albeit expensive) submersed weed control at herbicide concen-
trations well below the level of fish toxicity.
Liquid formulations of copper organochelate formulations have almost
entirely replaced CSP. Copper complexes are generally injected into sub-
mersed weed mats in combination with diquat or endothall. There are several
copper chelate products on the market. The three principal products are:
copper triethanolamine, copper ethylene diamine, and copper diethylene
triamine.
Residues in Water and Soil. Major large-scale field tests of diquat and
copper combinations were conducted in 1971 by the U.S. Army Corps of Engi-
neers in cooperation with several state and federal agencies (Mobley e_t al.
1971). Herbicide residues, invertebrate populations, water quality investi-
gations, etc., were monitored concurrently in several stations in the
250-acre treatment site. The herbicides diquat and CutrineR (copper
triethanolamine) were applied in combination at rates of 0.75 ppm diquat and
0.14 ppm copper. The area treated was 85 acres, which comprised 694 acre
feet of water.
72
-------
Table 24.
Toxicity of commonly used aquatic herbicides to bluegill fingerlings.
Mortality is cumulative across the table. From Haller, unpublished data.
Percent mortality
n=20 2 hr 20 hr 72 hr 96 hr
Diquat:
0.5
1.0
2.5
5.0
10.0
20.0
Cutrine Plus:
5.0
10.0
20.0
40.0
80.0
Copper sulfate:
5.0
10.0
20.0
40.0
Weedar 64:
(2.4-D)
5.0
100
20.0
40.0
80.0
Control:
A
B
0
0
0
0
0
0
0
0
10
40
5
0
60
100
100
0
0
0
0
0
0
0
10
0
10
5
5
5
0
5
100
100
100
0
' 100
100
100
5
0
0
15
100
0
0
10
0
10
5
10
30
0
5
100
100
100
0
100
100
100
5
0
0
15
100
0
0
10
0
10
5
10
35
0
5
100
100
100
0
100
100
100
5
0
0
15
100
0
0
In an adjacent area (165 acres), diquat (1.0 ppm) and 0.8 ppm of copper
as copper sulfate were applied to nearly 1,400 acre feet of water.
The total area treated was 250 acres in a several thousand acre impound-
ment. The site was a hydrilla-infested embayment on the north side of a res-
ervoir. Weed control was achieved in the diquat-copper sulfate treatment
area, but minimal vegetation control occurred in the diquat-Cutrine site due
to the lower total herbicide concentrations used and internal water exchange
in the reservoir.
Copper residue data from the water are presented in Table 25. The back-
ground, or pretreatment water copper, varied from 0.001 to 0.011 ppm Cu.
Fourteen days after treatment, a range of 0.005 to 0.014 ppm copper was
found. During this period, it appeared that the aquatic plants had absorbed
most of the copper from the water column. The information presented in Table
26 indicates very high copper concentrations in the hydrilla plants almost
immediately after chemical treatment.
73
-------
Table 25.
Copper in solution after treatment of Inglis Reservoir. Samples for
stations 1 to 4 are from the area treated with diquat plus cutrine and stations
5 to 8 from the diquat plus CSP (T = top, B = bottom). From Mobley et a I., 1971
Station
Number
1.
2
3
4
5
6
7
8
Water
Depth
T
8
T
B
T
B
T
B
T
B
T
B
T
B
T
B
Copper in solution
0
.004
.009
.010
.006
.009
.002
.005
.002
.009
.002
.011
.005
.004
.005
.006
.001
Days
1
.021
.036
.010
.024
.125
.048
.005
.018
.060
.165
.038
.034
.350 '
.070
.215
.100
(ppmw)
after treatment
3
.014
.016
.012
.014
.024
.018
.011
.010
.027
.026
.024
.026
.080
.083
.125
.025
7
.012
.012
.012
.012
.007
.015
.006
.006
.012
.019
.006
.007
.012
.019
.012
.007
14
.007
.007
.005
.007
.005
.006
.006
.005
.008
.013
.007
.010
.008
.012
.014
.011
Table 26.
Copper content .of hydrilla after treatment of Inglis Reservoir. Plant samples
for stations 1 to 4 are from the area treated with diquat plus cutrine and
stations 5 to 8 from the diquat plus CSP (T - top, B = bottom).
From Mobley et al., 1971
Station
Number
1
2
3
4
5
6
7
8
Water
Depth
T
B
T
B
T
B
T
B
T
B
T
B
T
B
T
B
Copper in
dry plant
material (ppmw)
Days after treatment
0
4
7
6
10
4
7
4
3
7
7
7
12
7
6
6
4
1
252
31
258
17
260
15
196
16
1,870
38
126
12
1,040
43
2,460
1,050
3
75
22
104
14
550
15
103
35
1,470
90
214
16
1,750
112
1,230
294
7
109
24
172
18
43
20
52
11
2,000
79
75
11
1,750
1,480
990
238
14
32
40
34
18
92
42
38
17
218
51
50
10
615
3,240
550
1,580
28
23
12
11
9
15
22
12
9
a
36
11
14
340
410
49
32
aNo sample available.
74
-------
The large-scale test indicated that in a heavily vegetated area, diquat
was essentially nondetectable in the water seven days after treatment.
Several factors account for disappearance of copper ion from the aquatic
ecosystem. In tests such as this, copper is distributed over time into the
rest of the waterbody and is absorbed and diluted by algae and aquatic
plants. In certain waters, notably hard waters, copper ions are precipitated
by carbonate ions and are thus incorporated into sediments.
Since the 1970 field test, copper and diquat rates have been reduced to
those currently recommended: 4-6 Ib diquat and 4-8 Ib copper per acre, in-
dependent of depth. In water six feet deep, these rates provide a concentra-
tion of 0.4 ppm diquat and 0.5 ppm of copper.
In summary, the hydrosoil is the primary sump when copper treatments are
conducted in closed ponds and non-flowing situations. In large reservoirs,
lakes, canals and rivers, none of which have ever been treated in total, the
copper residue is diluted in water, plants and hydrosoil where it cycles in
the ecosystem at near background concentrations.
Dalapon (2,2-dichloropropionic acid)
Dalapon is a selective herbicide which is most effective on grasses and
other monocots. As an acid, it can be formulated into several salts and
other compounds. The vast majority of dalapon is formulated as mixtures of
sodium and magnesium salts.
Dalapon has a very low toxicity to animals (Paynter et al. 1960) and is
safe for fish and invertebrates above 100 ppm. In aquatic systems it is used
for emergent weed control only, principally cattail and other ditchbank
species. It is often used in combination with 2,4-D in ditchbank weed
control. Dalapon kills or suppresses grass species and the 2,4-D formulation
controls the broadleaf (dicot) weed species.
As a ditchbank herbicide, very little dalapon is expected to be found in
water. Thus, dalapon's occurrence in water would be incidental, and signifi-
cant concentrations are unlikely to occur in the water or soil of lakes,
reservoirs, or rivers.
Residues in Water and Soil. It is likely that dalapon residues occur in
small agricultural ditches and along canals where cattail and grasses have
been sprayed. As a simple organic acid, its breakdown in water and hydrosoil
is expected to be rapid.
Dalapon is applied as a foliar spray on ditchbank grasses where it is
apparently rapidly absorbed through the leaves and roots and translocated to
meristematic regions (Blanchard et al. 1960).
75
-------
Dalapon is most widely used in terrestrial weed control. Holstun and
Loomis (1956) studied leaching and decomposition of dalapon in soils. Solu-
tions of dalapon (120 ppm) were applied to soils with varying moisture, temp-
erature, and organic matter contents. In general, the degradation of dalapon
is enhanced by soils which have high moisture, high temperature and high or-
ganic matter contents.
The most extensive compilation of the use, toxicity, degradation, and
environmental effects of dalapon was published by Kenaga (1973). Kenaga com-
pared the chemical oxygen demand (OCX)) and biological oxygen demand (BCD) of
dalapon in water and found that dalapon is completely broken down through
bacterial degradation. The end products of dalapon degradation are 002,
H20, NaCl, and Cl^ (Figure 32).
fungi
CH3CHCICOOH»
a-Chloropropionic
acid
Natural constituent
in respiration and
energy cycles of
cells (Krebs, lactic
acid, etc.)
CH3CCI2COOH*
Dalapon
fnicrobial,
hydrolysis
[CH3CCIOHCOOH]
CH3COCOOHT + Cl
Pyruvic acid
CH3 CHO + C02
Acetaldehyde
I
CH3COOH'
Acetic acid
i
C02 + H20
photodegradation
[CH3CCICOOH]
Figure 32. Chart of the major metabolic and degradative routes of dalapon:
the starred acids will be in equilibrium with their anions with
the position of the equilibrium depending on pH and the specific
cations in a particular environment and the compounds in brackets
will be transient. From Kenaga, 1973.
76
-------
Frank et al. (1970) applied dalapon at rates of 6.7 to 20 Ib (active
ingredient)/acre to ditchbank weeds along irrigation canals. Water coliectea
immediately after spraying contained insignificant levels of dalapon (°'^
0.37 ppn). It was concluded that dalapon at these levels would be degraded
rapidly in the aquatic ecosystem, and even the highest (0.37 ppm) posed no
threat to use for irrigating agricultural crops.
The most recent investigation concerning dalapon and its environmental
effects on the aquatic ecosystem is that of Brooker (1976). Brooker treated
emergent aquatic weeds along canal banks with dalapon and 2,4-D. Dalapon
(25 kg a.i./ha) was applied in split application once in the fall of f97/*,
and again (12 kg a.i./ha) in the following spring with 2,4-D (1 kg a.i./ha).
These treatments effectively controlled emergent grasses and broadleaf weeds.
The chemically treated canal reach was compared to a similar upstream reach
which received hand weeding.
Brooker analyzed 140 water samples to determine the residues of the two
chemicals following application. After the fall treatment, dalapon was found
in the water at 38 ppb (one hour post treatment). After 8 days, dalapon res-
idues were no longer detectable in the treated area, and only trace amounts
(
-------
As in other studies, residue of dalapon in canal water of Brooker's
study was insignificant. The effects of chemical control of emergent weeds
in the canals of Essex, England, generally increased the growth of replace-
ment aquatic plants (Lemna and Callitriche) compared to similar areas of
hand weeding.
Effects of Herbicides on Water Quality
Previous sections of this report have discussed herbicide residues in
the aquatic environment (flora, water, soil, and fauna). Effects of aquatic
herbicides on water quality will be discussed separately in this section
because the herbicides per se have no effect on traditional aspects of water
quality (physio-chemical parameters). The changes that occur in water
quality are not directly a result of herbicide application, but rather are
a result of dying vegetation and changing of aquatic succession.
Research studies conducted with the objective of determining effects of
weed control on water quality are numerous, and results are as variable as
the containers used in the studies: erlenmeyer flasks to 10,000 acre lakes.
Water quality changes associated with chemical aquatic weed control result
from release of nutrients and organic matter from the decaying weeds. Even
then, not all nutrients contained in aquatic vegetation are released to the
water, considerable amounts of nutrients are incorporated into the soil and
benthic community as the vegetation falls to the hydrosoil (Strange 1976).
In small, enclosed systems such as aquaria, plastic pools, and similar
nonflowing, unnatural systems, herbicide application and subsequent decay of
aquatic vegetation invariably leads to at least temporary increases in dis-
solved plant nutrients and algal populations (Hestand and Carter 1978;
Walker 1963).
Nutrient released from treatment of peripheral emergent aquatic plants
in noneuthrophic lakes or river systems produce no changes or only temporary
minor changes in water quality and phytoplankton populations.
Factors that determine whether or not major changes in water chemistry
occur after weed control operations are dependent upon the physical and
chemical characteristics of a given water body (Table 28).
Table 28.
Characteristics of water which, if combined, will invariably produce major changes in water quality if treated for
aquatic weed control.
High water temperature.
High plant biomass to be controlled.
Shallow water, preferably eutrophic
High percentage of water surface area treated.
Closed ponds, non-flowing situations.
78
-------
According to Table 28, an algal bloom and deterioration of water quality
will invariably occur in the following situation: eutrophic shallow nonflow-
ing pond (1-2 m deep), 90% covered with water hyacinth (200 metric tons fresh
wt/acre), sprayed in July.
On the other end of the scale, the following treatment situation would
produce no measurable effect on water quality outside of the immediate treat-
ment area: oligotrophic, deep lake (100 acres), 100 feet long, 4 feet wide
fringe of shoreline Typha plants, sprayed in early spring at initation of new
growth.
These two extreme examples are used to emphasize the variability of
potential environmental effects of chemical control of aquatic weeds. A
trained and experienced aquatic biologist can accurately predict the outcome
of various weed control operations.
EPA and State pesticide laws require that permits be obtained by aquatic
weed control firms before control operations are conducted (particularly in
public waters). The permitting agency therefore must weigh the benefit-risk
ratio before permitting the operation.
Currently, Federal and State laws closely monitor and regulate aquatic
herbicide labelling directions and registration. The rigorous and extensive
requirements for herbicide registration for aquatic ecosystems insures that
only the safest possible chemicals are used in aquatic weed control. Any
detrimental environmental effects that occur as a result of chemical control
of aquatic weeds are not the fault of the chemical, but rather the reduction
and decay of the controlled vegetation.
Obviously, all alternative means of aquatic weed control must be eval-
uated before wise management decisions can be made. All means (mechanical,
biological, chemical, drawdowns, etc.) of aquatic weed control have environ-
mental impact, if for no other reason than eliminating undesirable vegeta-
tion.
Better aquatic weed control means control of aquatic plants more econom-
ically and without increasing the environmental impact of control operations.
Effects of Herbicides on Aquatic Organisms
The herbicide 2,4-D is used extensively for the control of aquatic
plants. Several formulations of this chemical are used. For example, the
diethylamine salt is used for water hyacinth control, whereas 2,4-D BEE is
used for milfoil. The chemical formulation often determines the toxicity of
the chemical to aquatic life. Hiltibran (1967) pointed out that the water
soluble derivatives of 2,4-D were less toxic than their ester derivatives,
and Walker (1964a,b) found that butyl, butoxyethanol and propylene glycol
butyl ether esters more toxic to fish than several other 2,4-D formulations.
The diethylamine salt of 2,4-D is the derivative used most often for hyacinth
control and appears to be the least toxic of the ^forms.
79
-------
2,4-D is usually applied at a rate of 2 mg/1 to the treatment area. It
is possible, however, to have areas of higher concentration, especially where
plants are thick and higher rates are applied. All formulations [2,4-D PGBE;
2,4-D BEE; 2,4-D IOE and 2,4-D (diethylamine salt)] are relatively nontoxic
to aquatic animals. Acute toxicity levels (LC50), as tabulated in Table 29
(Anon. 1973), indicate that LC50 values range from 100 g/1 (2,4-DPGBE) to
having no effect at 100,000 g/lN 48 hours. Other studies have reported
insignificant changes in the abundance or disti bution of freshwater plankton
or benthos (Pierce 1968, 1969). Alabaster (1969) reported an LC50 of
250 mg/1 (24 hours) for rainbow trout, Salmo gairdnerii. Hughes and Davis
(1962) found 48-hour TL^ values for bluegills (Lepomis microchirus) to
range from 0.8 to 840 mg/1 acid equivalent for several formulations of 2,4-D.
McCorkle et al (1977) considered 2,4-D acid and diethylamine salts of the
compound to be safe for use around catfish ponds as few mortalities occurred
when channel catfish (Ictalurus punctatus) were exposed to the chemical.
Application of 25 to 30 1 2,4-D/acre did not affect fish in hatchery ponds
(Mackenthun 1961). In most of the studies reported above, larger fish were
used. Hiltibran (1967) utilize small bluegill and fry of the fish in his
study (Table 30) and found 2,4-D to be relatively nontoxic to these organ-
isms. The diemethylamine salt had no effect' as all test fish held to term.
Mount and Stephen (1967) reported that fathead minnows (Pimephales promelas)
reproduced normally when exposed to 0.3 ppm 2,4-D.
The effects of 2,4-D on marine organisms has also been investigated.
Rawls (1965) tested eastern oysters, soft shell clams, and various fish
species in conjunction with herbicidal treatments. Chemicals were applied as
follows: 2,4-D acetamide 20 Ib/acre; polypropylene glycol butyl ether esters
of 2,4-D 20 Ib/acre; butoxy ethanol ester of 2,4-D 20 to 120 Ib/acre; isoctyl
ester of 2,4-D 20 to 60 Ib/acre. He concluded, after a one month trial, that
only 2,4-D acetamide appeared dangerously toxic. Butler (1965) found pink
shrimp (Penaeus duorarum) showed no effect at 1.0 ppm 2,4-D at 48 hours,
whereas brown shrimp (Penaeus aztecus) showed 10% mortality or paralysis at
2.0 ppm at 48 hours.
Residue studies indicate similar findings as do toxicity tests. Fish
residue studies indicate little 2,4-D uptake. Schultz and Whitney (1974)
determined residues in 60 fish samples after water hyacinths were sprayed
with the dodecyl-tetradecyl amine salts of 2,4-D (DTA-2,4-D) at a rate of
4.48 kg acid equivalent/ha. The initial treatment was followed with spot
treatments of DTA-2,4-D or the diemethylaine salt of 2,4-D (DMA-2,4-d), or
both, over a four-month period. Residues ranged from zero in 40 samples to
greater than 0.1 mg/kg in two samples (Table 31). Cope et al. (1970) found
no residues in whole bluegills with pond treatments of 0.5 ppm or less with
propylene glycol butyl ether esters of 2,4-D.
80
-------
Table 29.
Herbicides, fungicides, defoliants From Anon., 1973.
Pesticide
Organism ug liter //g liter Reference
/yg/ liter hours
ACRQLE1N AQUALIN
AMINOTRIAZOLE AMITROL
BALAN
BENSULFIDE
CHLOROXURON
CIPC
DACTHAL
DALAPON (SODIUM SALT)
DEXON
DICHLOBENIL CASARON*
FISH
Lepomis macrochirus
Satmo trutta
Lepomis macrochirus
CRUSTACEAN
Gammarus fasciatus
Daphma magna
Cypndopsis vidua
Asellus brevicaudus
Palaemonetes kadiakensis
Orconectes nais
FISH
Lepomis macrochirus
Oncorhyncus kisulch
CRUSTACEAN
Gammarus factatus
CRUSTACEAN
Gammarus facratus
FISH
Lepomis macrochirus
FISH
Lepomis macrochirus
FISH
Lepomis macrochirus
CRUSTACEAN
Simocephalus serrulatus
Daphma pule*
INSECT
Pleronarcvs caMomica
FISH
Pimephales promelas
Lepomis macrochirus
Oncorhynchus kisutch
CRUSTACEAN
Gammarus lacustns
INSECT
Pieronarcys caMomica
CRUSTACEAN
Gammarus lacustns
INSECT
Pteronarcys californica
CRUSTACEAN
Gammarus lacusins
Gammarus tasciaius
Oaphnia magna
Cypndopsis vidua
Asellus brevicaudus
Palaemonetes kadtakensis
Orconectes nais
FISH
Lepomis macrochirus
CRUSTACEAN
Gammarus lacustns
Gammarus fasciatus
Hyallella azteca
Simocephalus serrulatus
Daphma pulex
Daphma pulex
Oaphnia magna
Cvpndopsis vtdua
Asellus brevicaudus
Palaemonetes kadiakensis
Orconectes nais
INSECTS
Pteronarcys californica
Tendipedtdae
Cailibaetes sp
Umnephilus
Enallegma
FISH
Lepomis macrochirus
80 24
46 24
79 24
30000 46
32000 48
325000
1100
1400
25000
8000
700000
16000
11000
290000
290000
340000
100
2100
3700
24000
3900
48
96
96
48
48
48
48
48
96
96
48
96
96
96
96
96
20000
11000
10000
8500
5800
3700
3700
10000
7800
34000
9000
22000
7OOO
7800
10300
13000
20700
48
96
96
96
48
48
48
48
48
48
48
48
96
96
96
96
96
Bond el al 1960
Burdick el al 1964
100000 wg/l 48 hr Sanders 1970
100 000 t/g I 48 hr
100 000 f/g/1 48 hr
100.000
-------
Table 29.
Continued.
Acute toxicity LC50 Sub-acute effects
Pesticide
DICHLONE PHYGON XL
DIQUAT
DIURON
DIFOLITAN
DINITROBUTYL PHENOL
DIPHENAMID
DURSBAN
2.4-D(PGBE)
2.4-DIBEE)
2. 4-0 (BEE)
2,4-DIIOEI
Organism
CRUSTACEAN
Gammarus lacustns
Gammarus fasciatus
Daphnia magna
Cyprrdopsis vidua
Asselus brevicaudus
Palaemonetes kadiakensis
Orconectes nais
FISH
Lepomis macrochirus
Micropterus salmoides
CRUSTACEAN"
Hyallelta azteca
INSECTS
Callibaetes sp
Limnephilus
Tendipedidae
Enallagma
FISH
Lepomts macrochirus
Micropterus salmoides
Esox lucius
Stizostedion vitreum vitreum
Salmo gairdnen
Oncorhvnchus tshawytscha
CRUSTACEAN
Gammarus lacustns
Gammarus fasciatus
Simocephalus serrutatus
Daphnia pulex
INSECT
Pleronarcys californica
FISH
Oncorhynchus kisutch
CRUSTACEAN
Gammarus lacustns
INSECT
Pteronarcys californica
CRUSTACEAN
Gammarus fasciatus
CRUSTACEAN
Gammarus fascratus
Daphnia magna
Cypridopsis vidua
Asellus brevicaudus
Orconectes nais
CRUSTACEAN
Gammarus lacustris
INSECT
Pteronarcys califormca
Pteronarcella badta
Claassenia sabulosa
CRUSTACEAN
Gammarus lacustns
Gammarus fasciatus
Daphnia magna
Cypridopsis vidua
Asellus brevicaudus
Palaemonetes kadiakensis
Orconectes nais
CRUSTACEAN
Gammarus lacustris
Gammarus fasciatus
Daphnia magna
Cypridopsis vidua
Asellus brevicaudus
Palaemonetes kadiakensis
Orconectes nais
INSECT
Pteronarcys californrca
FISH
Pimephales promelas
CRUSTACEAN
Gammarus lacustns
yg/ liter
1100
100
25
120
200
450
3200
70
120
48
16400
33000
>!00000
> 100000
14OOO
35000
7300
16000
2100
11200
28500
160
700
2000
14OO
1200
16000
800
40
1800
56000
50000
58000
0 It
10
038
057
16OO
2500
100
320
2200
2700
440
5900
5600
1800
3200
1400
1600
5600
240O
//g/liter
hours
96
96
48
48
48
48
48
48
48
96
96
96
96
96
96
96
96 ,
48
96
48
48
96
96
48
48
96
48
96
96
96
48
48
48
96
96
96
96
96
96
48
48
48
48
96
96
48
48
48
48
96
96 1 500 MJ/I lethal to eggs
in 48 hour exposure
*
96
No effect
vg/ liter Reference
Sanders 1969
Sanders 1970
"
"
"
Bond et al 1960
Hughes and Davis 1962
Wilson and Bond 1969
Wilson and Bond 1969
"
PI
Gilderhus 1967
Surber and Pickering 1962
Gilderhus 1967
"
Bond et al 1960
Sanders 1969
Sanders 1970
Sanders and Cope 1966
Sanders and Cope 1968
Bond et al 196C
Sanders 1969
Sanders and Cope 1968
Sanders 1970
1OO.OOO /jg/t 48 hr Sanders 1970
10O.OOO(/g/l 48 hr
100,000 «g/l 48 hr
Sanders 1969
Sanders and Cope 1968
"
Sanders 1969
Sanders 1970
1 00.000 ug/l 48 hr
Sanders 1969
Sanders 1970
100.000 ug/l 48 hr
Sanders and Cope 1968
300 j/g/l 10 mo Mount and Stephen 1967
Sanders 196S)
82
-------
Table 29.
Continued.
Acute toxicity LC50 Sub-acute effects No effect
Pesticide
2.J D [DIETHYLAMINE SALT)
ENDOTHALL DISODIUM SALT
ENDOTHALL CHPOTASSIUM SALT
EPTAM
FENAC (SODIUM SALT)
HYAMINE 1622
HYAMINE 2389
HYDROTHAL 47
HYOROTHAL 191
HYDROTHAL PLUS
IPC
KURON
MCPA
MOLINATE
MONUHON
PARAQUAT
PE8ULATE
PICLORAM
Organism
CRUSTACEAN
Gammarus lacustns
Gammarus fasciatus
Daphnia magna
Crypidopsis vidua
Asellus brevicaudus
Palaemonetes kadiakensis
Orconectes nais
FISH
Pimephales notatus
Lepomis macrochirus
Micropterus salmoides
Notropis umbratilus
Micropterus salmoides
Oncorhynchus tschawylsch
CRUSTACEAN
Gammarus lacustns
FISH
Pimephales promelas
Lepomis macrochirus
CRUSTACEAN
Gammarus fasciatus
CRUSTACEAN
Gammarus lacustns
Gammarus fasciatus
Daphnia pulex
Daphnia magna
Cypndopsis vidua
Asellus brevtcaudus
Palaemonetes kadiakensis
Orconectes nais
INSECT
Pteronarcys californica
FISH
Lepomis
FISH
Pimephales promelas
Lepomis macrochirus
Oncorhynchus kisutch
FISH
Lepomis macrochirus
CRUSTACEAN
Gammarus fasciatus
CRUSTACEAN
Gammarus lacustns
Gammarus fasciatus
FISH
Lepomis macrochirus
CRUSTACEAN
Gammarus lacusiris
Gammarus fasciatus
Simocephalus serrulatus
Daphnia pulex
CRUSTACEAN
Daphnia pulex
FISH
Lepomis macrochirus
CRUSTACEAN
Gammarus lacustns
Gammarus fasciatus
Daphnia magna
Asellus brevicaudus
Palaemonetes kadiakensis
FISH
Oncorhynchus kisutch
CRUSTACEAN
Gammarus lacustns
Simocephalus serrulatus
Daphnia pulex
INSECT
PteronarcYS californica
CRUSTACEAN
Gammarus fasciatus
CRUSTACEAN
Gammarus lacustns
INSECT
Pteronarcys californica
40OO
8000
110000
125000
120000
95000
200000
136000
320000
160000
23000
12000
45OO
6600
55000
15000
16OO
1400
53000
2400
1200
510
500
480
3500
10000
19000
10000
10000
2400
2000
1500
4500
300
600
400
1OOO
5600
110000
11000
4000
3700
100OO
270OO
48OOO
48
48
96
96
96
96
96
96
96
96
96
96
48
96
48
96
96
96
96
96
96
36
96
48
96
96
48
48
48
48
48
96
96
48
48
48
48
48
96
48
48
96
96
96
83
ng/ liter yg liter
100.000 (/g/1 48 hr
100000;jg'l 48 hr
1 00.000 us/I 48 hr
100.000(^/1 48 hr
lOOOOOug-l 96 hr
1 00 COO ug 48 hr
1OOOOO fig 1 48 hr
tOO.OOO(/g 1 48 hr
100.000 ug'l 48 hr
100.000 wg/l 48 hr
lOO.COOug/l 96 hr
Reference
Sanders 1969
Sanders 1970
Walker 19643
Bond et al 1960
"
Sanders 1969
Surber and Pickering 1962
Sanders 1970
Sanders 1969
Sanders 1970
Sanders and Cope 1966
Sanders 1970
Sanders and Cope 1968
Hughes and Davis 1962
Surber and Pickering 1962
Bond et al 1960
Surber and Pickering 1962
Sanders 197C
Sanders 196S
Sanders 1970
Hughes and Davis 1964
Sanders 19>"9
Sanders 1970
Sanders and Cope 1966
Sanders and Cope 1966
Hughes and Davis 1964
Sanders 1969
Sanders 1970
"
Bond el al 1960
Sanders 196S
Sanders and Cope 1966
Sanders and Cope 1968
Sanders 1970
Sanders 1969
Sanders and Cope 1968
-------
Table 29.
Continued.
Acute toxicity LC50 Sub-acute effects No effect
Pesticide
PROPANIL
SILVEX (BEE)
SILVEX IPGBE\
SILVEX HOE!
SILVEX (POTASSIUM SALTI
SIMAZINE
TRIFLURALIN
VERNOLATE
Organism
CRUSTACEAN - -
Gammarus fasctatus
CRUSTACEAN
Gammarus fasctatus
Daphnia magna
Cvpndopsis vidua
Asellus brevicaudus
Orconectes mas
FISH
Lepomis macrochirus
CRUSTACEAN
Gammarus fasciatus
Daphnia magna
Cvpndopsis vKJua
Asellus brevicaudus
Palaemonetes kadiakensis
Orconectes nais
FISH
Lepomis macrochirus
FISH
Lepomis macrochirus
FISH
Lepomis macrochirus
CRUSTACEAN
Gammarus lacustns
Daohnia magna
Cvpridopsis vidua
Asellus brevicaudus
Orconectes nais
FISH
Oncorhynchus kisutch
CRUSTACEAN
Gammarus lacustns
Gammarus fasciatus
Oaphnta magna
Daphnia pule*
Simocephalus serrulatus
Cypridopsis vidua
Asellus brevicaudus
Palaemonetes kadiakensis
Orconectes nais
INSECT
Pteronarcvs califormca
CRUSTACEAN
Gammarus lacustns
Gammarus fasciatus
Daphnia magna
Cypndopsts vidua
Asellus bravicaudus
Palaemonetes kadiakensis
Orconectes nais
ng/tner
16000
250
2100
4900
40000
800
60000
1100
840
180
200
500
3200
16600
16000
83000
13000
1000
3200
6600
2200
1000
560
240
450
250
200
1200
50000
3000
1800
13000
1100
240
5600
1900
24000
pg/liter /jg/ltter Reference
hours
96
96
48
48
48
48
48
48
96
48
48
48
48
48
48
48
96
48
48
48
96
96
48
48
48
48
48
48
48
96
96
96
48
48
48
48
48
Sanders 1970
Sanders 1970
,.
Hughes and Davis
Sanders 1970
-
-
1OO.OOO yg/l 48 hr
Hughes and Davis
Hughes and Davis
Hughes and Davis
Sanders 1969
100,00 j/g/l 48 hr Sanders 1970
10000Oug 1 48 hr Sanders 1970
100.000 /yg 1 48 hr
100.000 t/g 1 48 hr
Bond et al I960
Sanders 1969
Sanders 1970
Sanders and Cope
Sanders 1970
Sanders and Cope
Sanders 1969
Sanders 1970
..
1963
1963
1963
1963
1966
1968
84
-------
Table 30.
Effects of various concentrations of herbicides on small bluegills and fry from
four species of fish From Hiltibran. 1967
Common name
Amitrole
Arsenite
Atrazine (G)2
(WP)2
Dalapon
Dichlobenil (G)
Diquat cation
Endothall (L):
(G)
Fenac
Fenuron TCA (G)
Monuron TCA (G)
2.4-D
PGBE ester (G)
Mixed isopropyl-
butyl esters (L)
2-ethylhexyl
ester (G)
Dimethylamme
salt (L)
Sodium salt
2.4-DP
Isooctyl ester (G)
2,4.5-T
Isooctyl ester (L)
Isooctyl ester (G)
Sodium salt
Silvex
PGBE ester (L)
(G)
Potassium salt (L)
(G)
Sodium salt
Simazine (G)
(WP)
Effect
of
herbi-
cides
on
blue-
gill"
25'
10
5
50'
20
10
100
50
50
20
20
2
3
50
40
100
20
1
10
50
2
20
50
30
50
10
30
E
Q.
a
50'
15
8
10
10
50'
25
10
25
1 3
25'
10
20
10
10
5
1
4
1
10
5
25
10
1.5
4
1
10
2.4
1.0
10
20
10
10
10
Days
CD
3
CO
Term3
7
Term
Term
Term
Term
3
Term
Term
5
Term
Term
1
2
2
5
Term
Term
Term
Term
Term
Term
2
Term
Term
Term
Term
survival
Green
sunfish
Term
Term
Term
Term
Term
Term
Term
Term
Term
5
5
Term
1
4
Term
1
Term
Term
4
Term
Term
7
5
of fry
.0
r.
o w
-*
-------
Diquat
Diquat is usually applied at a rate of 0.5 to 1.0 ppm over the area
treated. It is applied as a spray to foliage of floating weeds or injected
below the surface for control of submersed weeds. For most species this
treatment should not have detrimental effects (Tables 29 and 32). Several
authors, however, reported toxicity concentrations for certain organisms that
approached these levels. Lawrence (1962) reported 0.5 + 0.5 mg/1 as toxic to
Pimephales promelas , Lepomis gibbosus , and Micropterus salmoides . His
studies were accomplished under field conditions. The values he obtained for
M. salmoides were not in agreement with the report of other authors (Tables
29 and 32). Hughes (1973) reported 1.0 mg/1 was toxic to Morone saxatilis.
Tatum and Blackburn (1965) reported values of 0.5 mg/1 killed chironomids and
oligochaetes. Wilson and Bond (1969) reported 0.048 mg/1 was toxic to
Hyalella azteca and Gilderhaus (1967) reported 1.0 mg/1 was toxic to Daphnia
pulex (Table 32). Hiltibran (1967) reported the effects of diquat cation on
small bluegills and fry of other species (Table 30) and concluded that at
recommended rates the chemical would not greatly effect survival of fish
reproductive products.
Endothall
Sodium and potassium salts, aluminum oxides and acid of endothall are
used for control of aquatic weeds. A derivative (dime thy lamine salt) is
especially effective for hydrilla control (Dumas 1976). This derivative is
toxic to fish if it exceeds 0.2 mg/1 in water. Dipotassium or disodium
endothall is recommended for use at 5 mg/1 and should be safe at this level
(Dumas 1976). Hilitbran (1967) postulated that endothall would be one of the
safest herbicides to use during the spawning season due to its low toxicity
to fish (100 mg/1). Other authors have reported similar results (Tables 29
and 33).
Copper
Copper sulfate is used for algae control and can be toxic. Cairns
(1974) lists toxicities of heavy metals including copper (CuSO4) (Table 34)
to 13 species of protozoa. Toxicities were quite low; however, the author
does not stipulate the concentration in his tables. This author also pres-
ented dose-response curves for various protozoan species which indicated that
species responded differently in a time sequence. He stated that differences
in sensitivity might be explained if more were known about sensitivity of
copper at the cellular level. Copper ion toxicities are also presented in
Table 34. This ion is toxic to marine and freshwater organisms; however,
toxicities often varied according to experimental conditions. For example,
Palmer and Maloney (1955) reported that 0.18 ppm CuSCs in acute static
bioassays was toxic (96 hr LC50) to Pimephales promelas. Tarzwell and
Henderson (1960) found CuS04 toxicity varied according to water conditions:
the compound was toxic at 0.05 ppm in soft water and 1.4 ppm in hard water.
Doudoroff et al. (1966) used a copper cyanide complex in their tests and
found this compound was toxic at 1.5 ppm. Mount (1968) reported copper to be
toxic at 470 mg/1 in continuous flow bioassays with _P. promelas. Similar
examples can been seen for other species, indicating the importance of
standardized tests and data presentation.
86
-------
Table 31.
Residues of 2,4-D in fish from Loxahatchee National Wildlife Refuge. From Schultz and Whitney,
Station
1
1
1
1
3
3
3
1
1
2
2
2
3
3
3
1
1
1
2
Date
4-27-71
4-28-71
4-28-71
4-28-71
4-28-71
4-28-71
4-28-71
6-17-71
6-17-71
6-18-71
6-18-71
6-18-71
6-18-71
6-18-71
6-18-71
6-19-71
6-19-71
6-19-71
6-20-71
Species
Redear sunfish
ILepomis
microlophus)
Largemouth bass
(Micropterus
salmoidesl.
Brown bullhead
(Ictalurus
nebulosusl
Gar
(Lepisosteus
sppj
Chubsuchers
lErimyzon spp.l
Largemouth bass
Redear sunfish
Redear sunfish
Brown bullhead
Chubsuckers
Gar
Redear sunfish
Chubsuckers
Redear sunfish
Largemouth bass
Gar
Redear sunfish
Brown bullhead
Redear sunfish
Length,
CM
178
152
280
280
305
25.4
356
330
30 5
38 1
483
330
279
279
279
254
203
203
178
17 8
178
17 8
17 8
279
254
305
229
229
305
254
254
279
229
457
38 1
457
330
356
203
17 8
17 8
229
178
330
330
305
279
279
229
229
203
178
356
279
279
406
483
559
330
229
203
254
254
279
229
15 2
203
Weight.
G
91
45
227
228
454
227
182
136
91
182
454
368
272
272
227
182
227
227
136
91
91
91
91
343
227
454
182
182
454
227
227
272
91
454
136
368
91
182
182
136
91
272
91
590
454
409
227
227
182
182
136
91
454
318
318
227
454
454
136
227
136
227
227
318
272
46
136
Residues
of 2.4-D.
MG/K.G'
0000
0000
0000
0000
0000
0000
0.000
0 101
0012
0000
0000
0000
<0.010
0000
0000
0000
<0010
0000
0000
0000
0000
Station Date
3 6-20-71
1 6-20-71
3 6-24-71
3 6-24-71
1 7-1-71
2 7-2-71
2 7-2-71
2 7-2-71
3 7-2-71
3 7-2-71
1 7-15-71
1 7-15-71
2 7-15-71
2 7-15-71
3 7-15-71
1 8-12-71
1 8-12-71
1 8-12-71
1 8-12-71
2 8-12-71
3 8-12-71
3 8-12-71
Species
Gar
Redear sunfish
Gar
Chubsuckers
Redear sunfish
Chubsuckers
Redear sunfish
Bluegill
ILepomis
macrochirust
Redear sunfish
Chubsuckers
Bluegill
Redear sunfish
Chubsuckers
Bluegill
Bluegill
Gar
Brown bullhead
Redear sunfish
Chubsuckers
Redear sunfish
Chubsuckers
Largemouth bass
Length.
CM
38 1
330
63.5
457
43 2
203
38 1
356
279
305
254
254
279
21 5
178
330
305
203
254
22 9
203
22 9
20 3
203
20 3
330
203
203
189
18 9
22 9
305
330
178
17 8
203
203
203
17 8
203
22 9
17 8
22 9
38 1
483
432
38 1
305
279
178
178
178
17 8
17 8
25 4
27 9
27 9
254
279
22 9
22 9
178
229
17 8
30 5
330
254
27 9
254
27 9
Weight,
G
227
136
1090
368
368
590
545
368
409
318
318
368
227
113
499
454
182
272
182
182
227
182
182
182
454
136
227
113
113
294
454
567
227
204
227
227
227
113
227
340
182
227
136
454
272
681
318
272
91
91
91
91
91
227
272
318
227
318
227
227
91
227
91
454
567
227
272
227
227
Residues
of 2.4-0.
MG-KG'
0022
0000
0.000
0000
0000
0000
0000
0000
0022
0000
0000
0000
0010
<0010
0000
0000
0000
<0010
0000
0 106
<0010
<0010
<0010
0 162
0000
0012
87
-------
Table 31.
Residues of 2,4-D in fish from Loxahatchee National Wildlife Refuge. From Schultz and Whitney, 1974.
Length,
Station Date Species CM
2 6-20-71 Gar 559
40.6
48.3
38.1
2 6-20-71 Brown bullhead 27 9
25.4
279
3 6-20-71 Largemouth bass 27 9
3 6-20-71 Chubsuckers 25.4
27.9
25.4
229
254
25.4
279
27.9
279
305
3 8-12-71 Redear sunfish 20.3
20.3
22 9
178
17.8
1 10-7-71 Brown bullhead 305
343
35 6
38.1
Residues
Weight. of 2.4-D.
G MG/KG1 Station Date Species
727 0.028 1 10-7-71 Gar
227
368 1 10-7-71 Redear sunfish
227
272 0.000
227
227
227 0000 2 10-7-71 Redear sunfish
272 0 000
272
272
182
227 2 10-7-71 Chubsuckers
227
272
227
368
368 i 3 10-7-71 Largemouth bass
227 0.024 ,
227 !
368 !
136
136 J 3 10-7-71 Redear sunfish
368 0.000
545
681 i
681 3 10-7-71 Chubsuckers
i
Length.
CM
635
66 1
205
20.3
20.3
19.1
242
22 3
21 6
21 6
229
220
369
356
267
343
356
356
305
299
305
330
22 9
22 9
21 6
21 6
330
35 6
330
330
Weight.
G
1362
1644
227
182
227
182
368
227
227
227
318
182
726
681
272
636
681
636
409
318
409
590
227
227
227
227
545
772
590
636
Residues
of 2.4-D.
MG/KG'
0000
0000
<0010
0000
0000
0070
0000
T
'When only one set of data is given fora series of fish, the fish were composited for analysis
88
-------
Table 32.
Diquat: toxicity to aquatic organisms From Folmar, 1977.
Organism
Tendipedidae
Unnamed chironomids
Callibaetis spp.
Lemnephilus spp.
Daphnia pu/ex
D. magna
Hyalella azeteca
Unnamed oligochaetes
Cardium edule
Crassostrea virginica
Enallagma spp.
Libellula spp.
Penaeus setiferus
Crangon crangon
Aquatic insects.
amphipods, copepods,
ostracods
Pimephales promelas
P. promelas
P promelas
Fundulus similis
Carassius auratus
Rasbora heteromorpha
Icta/urus punctatus
(fry)
/. punctatus
Lepomis cyanellus
L gibbosus
L macrochirus
(fry)
L macrochirus
(fingerlings)
L macrochirus
L macrochirus
(fry)
L. macrochirus
L macrochirus
Micropterus salmoides
(fry)
M. salmoides
M. salmoides
M. salmoides
M salmoides
M. salmoides
Type of test
L. ST, A
FP
L, ST, A
L, ST, A
L, ST, A
L, ST, A
L, ST, A
FP
L, ST, A
L, CFT, A
L, ST, A
L, ST, A
L, CFT, A
L, ST, A
FP
L, ST, A
L, ST, A
FP
L, CFT, A
FP
L, ST, A
L, ST, A
L, ST, A
FP
FP
L, ST, A
L, ST, A
FP
L, ST, A
L, ST, A
L, ST, A
L, ST, A
L. ST, A
FP
L, ST, A
L, ST, A
FP
Experimental
conditions
a
a
a
a,b,c,d,e
a,b,c,d,e,f
a
a
a
a
a,b,c,d,e
a,b,c,d,e
a,b,c,d.e
a,b,c,d,e
a
a,b,c,d,e
a,b,c,d,e
a,b,c,d,e
a,b,c,d,e
a
a,b,c_d,e
a.b.c.d.e
a,b,c,d.e
a,b,c,d,e
a
a
Toxicity
>100T(96 h)
0.5 K (incomplete)
16. 4 T (96 h)
33.0 T (96 h)
1.0T(8d)
3.0 K (8 d)
7.1 IC50(26 h)
0.048 T (96 h)
0.5 K (incomplete)
>10.0T(24 h)
1.0 NTE(96 h)
>100T(96 h)
>100T(96 h)
1.0 NTE (48 h)
>10.0T(24 h)
1.0 NTE (7 d)
10.0 NTE (96 h)
Soft water
14.0 T (96 h)
Hard water
14.0 T (96 h)
0.5 +0.5
Paraquat NTE
1.0 NTE (48 h)
35.0 T (96 h)
70.0 T (48 h)
10.0 NTE (72 h)
10.0 NTE (96 h)
1 ,000 NTE
0.5 + 0.5
Paraquat NTE
10.0 NTE (12 d)
525 T (24 h)
150T(48 h)
25.0 T (96 h)
4.0 NTE (72 h)
10.0 NTE (96 h)
Soft water
MOT (96 h)
Hard water
140T(96 h)
1.0 NTE (72 h)
Soft water
7.8T(96 h)
1 ,000 NTE
11.0T (48 h)
10.0 NTE (96 h)
0.5 + 0.5
Paraquat NTE
Reference
Wilson and Bond (1969)
Tatum and Blackburn (1965)
Wilson and Bond (1969)
Wilson and Bond (1969)
Gilderhaus(1967)
Crosby and Tucker (1966)
Wilson and Bond (1969)
Tatum and Blackburn (1965)
Portmann and Wilson (1 971 )
Butler (1965)
Wilson and Bond (1969)
Wilson and Bond (1969)
Butler (1965)
Portman and Wilson (1971)
Hilsenhoff (1966)
Lawrence et al. (1965)
Surber and Pickering (1962)
Lawrence (1962)
Butler (1965)
Gilderhaus (1967)
Alabaster (1969)
Jones (1965)
Lawrence et al. (1965)
Yeo(1967)
Lawrence (1962)
Hiltibran (1967)
Hughes and Davis (1962)
Gilderhaus (1967)
Jones (1965)
Lawrence et al. (1965)
Surber and Pickering (1962)
Jones (1965)
Surber and Pickering (1962)
Yeo(1967)
Muirhead-Thompson (1971)
Lawrence et al (1965)
Lawrence (1962)
89
-------
Table 32.
Continued.
Organism
Morone saxatilis
(larvae)
M. saxatilis
(fingerlings)
M. saxatilis
(fry)
M. saxatilis
(fingerlings)
Stizostedion vitreum
Esox lucius
Trout
Onchorynchus
tshawytscha
Salmo gairdneri
S. gairdneri
S. gairdneri
Type of test
L, ST, A
L, ST, A
L, ST. A
L, ST, A
' FP
FP
L, ST, A
L, ST, A
FP
L, CFT, A
Experimental
conditions
a,b,c,e
a,b,c,e
a,b,c,d,e
a
a,b,c,d,e
a,b,c,d,e
a
a,b,c,d,e
a,b,c,d,e
Toxicity
LOT (24,
48, 72, 96 h)
35.0 T (24 h)
25.0 T (48 h)
15.0T(72 h)
10.0T(96 h)
35.0 T (24 h)
25.0T(48 h)
15.0T(72 h)
10.0T(96 h)
31 5 T (24 h)
155T(48 h)
80.0 T (96 h)
2.1 T(96 h)
16.0T(96 h)
20.0 T (24 h)
29.0 T (48 h)
5.0 NTE (96 h)
11.2T(96 h)
10.0 NTE; no
avoidance
Reference
Hughes (1973)
Hughes (1973)
Hughes (1969)
Wellborn (1969)
Gilderhaus(1967)
Gilderhaus (1967)
Holden(1964)
Muirhead-Thompson (1971)
Lawrence et al. (1965)
Gilderhaus (1967)
Folmar (1976)
1. Type of test, letters represent:
T-Toxicity test, used in conjunction with
ST= static
CFT = continuous flow
A = acute
C = chronic
L= Laboratory toxicity test
F= Field study, used in conjunction with
R = river, stream, or creek
M = marine
E = estuarine, and
0 = other
2. Toxicity, active ingredient. All values in mg/1 (ppm) unless otherwise noted. Letters included with numerical
values represent:
T= LC50 (also accompanied by a time factor, eg., 96 h)
K=Kill
SB = sublethal effects
NTE = No toxic effect
3. Experimental conditions. Factors reported in cited articles:
a - water temperature
b = pH
c= alkalinity
d = dissolved oxygen
e = dissolved solids
f = photoperiod
90
-------
Table 33.
Endothall: toxicity to aquatic organisms. From Folmar, 1977.
Organism
Semotilus atromaculatus
Lepomis macrochirus
Salmo gairdneri
Gammarus fasciatus
Salmo gairdneri
Lepomis macrochirus
(fingerlings)
L. macrochirus
L macrochirus
(fry)
L macrochirus
(fingerlings)
L. macrochirus
M/cropterus salmoides
(fry)
Icta/urus punctatus
Daphnia magna
Gammarus lacustris
G. lacustris
Mercenaria mercenaria
(eggs)
M. mercenaria
(larvae)
Crassostrea virginica
Notropis lutrensis
N. umbratilis
Pimephales notatus
P promelas
Erimyzon sucetta
Type of test
L, ST, A
L, ST, A
L, CFT, A
L, CFT, A
L, ST, A
L. ST, A
RP
L, ST, A
L, ST, A
L, A
L, ST, A
L, ST, A
L, ST, A
L, CFT, A
L, ST, A
L, ST, A
L, ST, A
L, ST, A
L. ST, A
L, ST, A
L, ST, A
L, ST, A
L, ST, A
Experimental
conditions
a,b,c,e
a,b,c,e
a,b,c,d,e
a,b,c,d,e
a,b,c,d,e
Pathological
examination
a.b.c.d.e
a,b,c,d,e
a,b,c,d,e
a,b,c,d,e
a,b,c,d,e,f
a,b,c,d,e
a
a
a
a
a,b,c,e
a,b,c,e
a,b,c,e
a,b,c,d.e
Toxicity
1,600 NTE (24 h)
428T(24h)a
268 T(48 h)
10 NTE, no avoidance8
3.1 T(24 h)b
2.1 T(48 h)
0.48 T (96 h)
1.5T (48 h)b
0.8 T (24 h)b
0.8 T (48 h)
S.B. 0.3 0.03 (28 d)b
0.75 NTE (72 h)c
0.3 T (24 h)c
0.3 T (48 h)
376 (3-4 weeks)
i.p. injection0
0.075 NTE (72 h)c
0.2 NTE (72 h)c
46.0 IC50 (26 h)a
100T(6 h)d
>320 T (96 h)°
50.0 T (48 h)d
>10.0(10d)d
>25.0(12 d)d
95.0 T (96 h)d
105T(96 h)d
110-120T(96 h)d
Soft Water
320 T (96 h)d
Hard water
6.0 T (96 h)
25.0 NTE (12 d)e
Reference
Gillette et al. (1952)
Davis and Hughes (1963)
Folmar (1976)
Sanders (1970)
Cope (1965)
Hughes and Davis (1962)
Eller (1969)
Jones (1965)
Hughes and Davis (1962)
Walker (1964b)
Jones (1965)
Jones (1965)
Crosby and Tucker (1966)
Sanders (1969)
Nebecker and Gaufin (1964)
Davis and Hidu (1969)
Davis and Hidu (1969)
Davis and Hidu (1969)
Walker (1963, 1964a)
Walker (1963, 1964a)
Walker (1963, 1964a)
Surber and Pickering (1962)
Hiltibran (1967)
(fingerlings)
91
-------
Table 33.
Continued.
Organism
Carassius auratus and
Cyprinus carpio
Ictalurus punctatus
(fry)
/. nebulosus
/. me/as
Lepomis cyanellus
L. macrochirus
(fry)
L. macrochirus
(fry)
L. macrochirus
(fingerlings)
L. macrochirus
(fingerlings)
L. macrochirus
L macrochirus
L. microlophus
Micropterus dolomieui
(fingerlings)
M. dolomieui
(fingerlings)
M. dolomieui
M. salmoides
(fry)
M. salmoides
(fingerlings)
M. salmoides
Morone saxatilis
(fingerlings)
Notropis umbratilis
N lutrensis
Pimephale notatus
Enmyzon sucetta
(fingerlings)
Ictalurus punctatus
Lepomis macrochirus
L. macrochirus
(fry)
L. macrochirus
(fry)
L. macrochirus
(fingerlings)
L. macrochirus
L. cyanellus
(fingerlings)
Micropterus salmoides
M. salmoides
(fry)
Salmon
Salmo gair drier/
Type cf test
L, ST. A
L, ST, A
L. ST, A
L, ST, A
FP
L, ST, A
L, ST, A
L, ST, A
L, ST, A
L, ST, A
L, ST, A
L, ST, A
L, ST, A
L, ST
FP
L, ST, A
L, ST, A
L, ST; A
L. ST, A
L, ST, A
L, ST, A
L, ST, A
L, ST, A
L, ST, A
L, ST, A
L, ST, A
U ST, A
L, ST, A
L, ST, A
L, ST, A
L, ST, A
L, ST, A
L, ST
L, ST, A
Experimental
conditions
a,b,c,e
a,b,c,d,e
a,b,c,e
a,b,c,e
a,b,c,d,e
a,b,c,d,e
a,b.c,d,e
a,b,c,e
a,b,c,d,e
a,b,c,e
a,b,c,d,e
a,b,c,d,e
a,b,c,d,e
a.b.c.e
a,b,c,d,e
a.b.c.d.e
a,b,c,d,e
a,b,c,d,e
a,b,c,d,e
Toxicity
145-210 T (96 h)e
100NTE(72h)e
170-175T(96 h)a
180-185T(96 h)d
3.0 NTEa
50.0 NTE (72 h)°
100 NTE (12d)e
Soft water
1 SOT (96 h)d
Hard water
160T(6 h)
25.0 NTE (12 d)8
125-1 SOT (96 h)d
450 T (48 h)e
280 T (96 h)
125T(96 h)d
25.0 NTE (12 d)e
10.0NTE(12d)a
3.0 NTEd
10.0 NTE (72 h)d
Soft water"
200 T (96 h)
120-125 T(96 h)°
2,000 T (24 h)d
1,700T(48 h)
710 T (96 h)
400 NTE (21 d)'
40.0.NTE(21 d)'
40.0 NTE (21 d) ''
10.0 NTE (12d)'
50.0 NTE (72 h)'
100 NTE (21 d)r
2.0 NTE (72 h)1
50.0 NTE (12 d)'
10.0 NTE (12 d)'
650 T (48 h)1
280 T (96 h)
10.0 NTE (12 d)'
10.0 NTE (21 d)1
2.0 NTE (72 h)'
NTE (21 d)'
10.0 NTE (21 d)(
Reference
Walker (1963, 1 964a)
Jones (1965)
Walker (1963, 1964a)
Walker (1963, 1964a)
Yeo(1970)
Jones (1965)
Hiltibran (1967)
Surber and Pickering (1962)
Hiltibran (1967)
Walker (1963, 1964a)
Hughes and Davis (1965)
Walker (1963, 1964a)
Hiltibran (1967)
Hiltibran (1967)
Yeo(1970)
Jones (1965)
Surber and Pickering (1962)
Walker (1963, 1964a)
Wellborn (1971)
Lindaberry (1961)
Lindaberry (1961)
Lindaberry (1961)
Hiltibran (1967)
Jones (1965)
Lindaberry (1961)
Jones (1965)
Hiltibran (1967)
Hiltibran (1967)
Hughes and Davis (1965)
Hiltibran (1967)
Lindabrry (1961)
Jones (1965)
Lindaberry (1961)
Lindaberry (1961)
1 di K salt formulation
1 TD-191 (mono-N. N dimethylmine salt) formulations.
TD-47 (di-N, N dimethylmine salt) formulation.
di Na salt formulation.
di Na salt formulation liquid.
di Na salt formulation granular.
92
-------
Table 34.
Lethal and tolerant concentrations of 12 toxic compounds tested on 13 species of Protozoa. From Cairns, 1974.
Test organism
Chilomonas paramecium
Peranema trichophorum
Tetrahymena pyriformis
Paramecium caudatum
Paramecium
mu/timicronuc/eatum
Stentor coeruleus
Euglena gracilis
Chlamydomonas sp.
Chlamydomonas reinhardi
Blepharisma
Amoeba proteus
Eglena acus
Chaos carolinensis
Test organism
Chilomonas paramecium
Peranema trichophorum
Tetrahymena pyriformis
Paramecium caudatum
Paramecium
multimicronucleatum
Stentor coeruleus
Euglena gracilis
Chlamydomonas sp.
Chlamydomonas reinhardi
Blepharisma
Amoeba proteus
Euglena acus
Chaos caro/inensis
Cr8'
(as K2Cr2O7)
1000
(>18)
160
(100)
1000; >1000
(180); (750)
5000
(1000)
>1000
(320)
>1000
(180)
1000
(32)
Zn2* (as
ZnSO«
7H20)
>10; 3.2
(18); (5. 6)
(1000)
5.6
(10)
32
(15.5)
10
(0.56)
42
1000
(5000)
1.8
(1.0)
(100)
100; 32; 56
(10); (5.6); (5.6)
(1000)
>1000
(320)
Phenol
1500
(560)
2500
(1000)
3200
(1000)
10
(1.35)«1.
(1000)
>1000
(750)
(1000)
(1000)
Co2* (as
CoCI2
6H2O)
>2500
(1000)
>5000
(2500)
Toxicant
Cu2*(as
CuSo4 5H20)
0.056
(0.024)
>100
(1.8)
10
(0.32)
35)
0.1; >1.0
(0.032); (0.24)
1.0
>100; 500
(0.1); (5. 6)
56
(0.1)
56
(18)
3.2. 3.2; 1 8
(0.1); (0 18); (032)
10
(1.0)
1.0
(0.18)
2.4
(0.1)
Toxicant
Nitric Acetic
acid acid
10 100; 32
(7.5) (180); (100)
56 1 000
(56) (75)
18 320
(10) (180)
13.5
(10)
(1000) 1000
(560)
32
(54)
Pb2 + [as Mn7' (as
Pb(N03)2] KMnCM
320 3.2; 5.6; 7.5
(5.6) (1.0); (1.0); (1.0)
>32
(1000) (3.2)
>100 3.2
(24) (0.75)
56 10
(24) (0.65)
10; 100
(1000) (3.2); (3.2)
100 18; 5. 6
(42) (10); (0.32)
18
«1.0)
Al3* (as Sn2' (as
AICI3 SnCI2
6 H20) 2H2O) HCI
24 10
(1 0) (7.5)
>1000
(560)
32 32 3.2
(1.0) (7.5) (2.4)
(100)
(1000)
93
-------
Although considerable data are available in the literature pertaining
to toxicity and responses of organisms to toxic chemicals, conditions under
which each test was run should be considered. Many factors can influence
the responses of organisms to these materials. It is much easier to control
environmental variables in the laboratory; it is impossible to control these
variables in the field. When animals are subjected to bioassays, physical,
and chemical factors of the environment as well as physiological mechanisms
of the living organisms should be considered. Often the physiological re-
sponses, especially the cellular response of lower organisms, are not known.
Physical and environmental variables including size, acclimation, water
source, temperature, dissolved oxygen, and number of organisms per test con-
tainer can be controlled in laboratory tests. Weiss and Botts (1957) studied
effects of acclimation time, individual size, and temperature differences of
three fish species, Pimephales promelas, Lepomis cyanellus, and Carassius
auratus, when exposed to sarin (isopropyl methylposphonofluoridale). Gen-
erally, smaller fish consumed more oxygen than larger individuals. The size
factor, however, was not strictly proportional among the groups in their
response. Lepomis cyanellus was tested over a wider range (Figure 33). The
T50 values increased by a factor of 1.6, whereas the oxygen consumption rate
decreased by a factor of 1.57.
OXYGEN CONSUMED - MG- GRAM 2 HOURS T-50 MINUTES
03 04 150 200 250 300 350
0
I
O
LU
g
I
o
<
cc.
6
Figure 33. The effect of size of sunfish on time of response to sarin (10
ppb) and oxygen consumption. From Weiss and Botts, 1957.
The effect of time in laboratory stock tanks in response to sarin was
determined (Figure 34). Over a period of three weeks, the T50 of all sizes
increased, and oxygen consumption decreased (Weiss and Botts 1957). Accord-
ing to the authors, differences found in response to fishes at different
times may have been due to some unknown seasonal factor. These authors also
studied the influence of water temperature on oxygen consumption and con-
cluded that temperature acclimatization prior to test exposure is necessary
to allow for adjustment of the oxygen consumption rate.
94
-------
111111 11111111111 111111
0 5 10 15 20
DAYS' IN LABORATORY
Figure 34.
The effect of time in laboratory stock tanks on the time of
response of sunfish to sarin (10 ppb) and oxygen consumption.
From Weiss and Botts, 1957.
Dalapon
Dalapon (CgHoCL^Op) is used extensively for controlling ditch-
bank weeds and spot treatment of cattails. Toxicity of this compound to
aquatic organisms is low, probable less toxic in most cases than 2,4-D.
Toxicity values for dalapon for various aquatic animals are presented in
Tables 29 and 35. These tables are literature values compiled by Anon.
(1973) and Folmar (1977). Other authors report similar results. Chancellor
and Ripper (1960) obtained an LC50 for brown trout (Salmo trutta) of 400 ppm.
Kenaga (1973) tested 12 species of freshwater fish in static bioassays and
found that LC50 was always greater than 100 ppm. Fish fry were resistant to
dalapon (Table 35) and showed no mortality after exposure to 50 ppm. Johnson
(1978) reported toxicity at 19 g/1 for mosquitofish (Gambusia affinis).
95
-------
Table 35.
Dalapon: toxicity to aquatic organisms. From Folmar, 1977
Organism
Simocephalus serrulatus
Daphnia pulex
Cardium edule
Crangon crangon
Pimephales promelas
Erimyzon sucetta
Rasbora heteromorpha
Flounder
Lepomis macrochirus
(fry)
L. macrochirus
L macrochirus
L. cyanellus (fry)
Micropterus dolomieui
(fry)
Trout
Pteronarcys californica
Crassostrea virginica
Penaeus aztecus
Fundulus similis
Cyprinus carpio
Rasbora heteromorpha
Lepomis macrochirus
(fingerlings)
Salmo gairdneri
Lepomis macrochirus
Type of test
L, ST, A
L, ST, A
L, ST, A
L, ST, A
L, ST, A
L, ST, A
L, ST, A
L, ST, A
U ST. A
L, ST, A
L, ST, A
L, ST, A
L, ST, A
L, CFT, A
L, CFT, A
L, CFT, A
L, CFT, A
F, P
L, ST, A
L, ST, A
L, CFT, A
L, ST, A
Experimental
conditions
a,b,c,e
a,b,c,e
a,b,c,d,e
a,b,c,d,e
a,b,c,d,e
a,b,c,d,e
Pathological
examination
a,b,c,d,e
a,b,c,d,e
a,b,c,d,e
a,b,c,d,e
Toxicity
16.0IC50(48 h)
11.0IC50(48 h)
>100T(24 h)
>100T(24 h)
Soft water
390 T (96 h)
Hard water
290 T (96 h)
50.0 NTE
44.0 T (48 h)
>100T(24 h)
50.0 NTE (12 d)
Soft water
440 T (96 h)
115T(48 h)
50.0 NTE
50.0 NTE
340 T (24 h)
100 NTE (96 h)a
1.0 NTE (96 h)a
LOT (48 h)a
1.0 NTE (48 h)a
SB 250, 25,
2.5 (28 d)a
240 T (48 h)a
>1,OOOT(24 h)b
>1,OOOT(48 h)
1.0 SB
Avoidance3
>1 ,0001(24 h)c
>1,OOOT(48 h)
Reference
Sanders and Cope (1966)
Sanders and Cope (1966)
Portmann and Wilson (1 971 )
Portmann and Wilson (1 971 )
Surber and Pickering (1962)
Hiltibran (1967)
Alabaster (1969)
Portmann and Wilson (1 971 )
Hiltibran (1967)
Surber and Pickering (1962)
Cope (1965)
Hiltibran (1967)
Hiltibran (1967)
Holden(1964)
Sanders and Cope (1968)
Butler (1965)
Butler (1965)
Butler (1965)
Schultz(1971)
Alabaster (1969)
Hughes and Davis (1962)
Folmar (1976)
Hughes and Davis (1962)
Na salt formulation.
Granular formulation.
Wettable powder formulations.
A field experiment conducted by Brooker (1976) was designed to compare
directly the ecological effects of the use of herbicides (dalapon and 2,4-D)
and hand clearance for drainage channel maintenance of emergent weeds. He
concluded that invertebrate densities generally remained similar during the
two study years. No species were lost due to herbicide treatment. Some
animals such as Herpetocypris, Calospectra, and Planorbia vorticulus
increased in the herbicide treatment area. Replacement plant species (algae,
submersed and floating-leafed species) apparently provided living space for
these organisms. For additional information of 'the effects of dalapon on the
environment, the review article by Kenaga (1973) should be consulted.
96
-------
Chronic Effects of Herbicides
The long-term or chronic effects of most herbicides have not been stud-
ied to any extent and pathophysiology of the chemicals has not been deter-
mined. Most toxicological studies have been conducted with laboratory
mammals or live stock. Chronic or toxicological effects of herbicides on
fish and wildlife are few.
Effects of 2,4-D on the eggs of the pheasant (Phasianus colchicus), red
partridge (Perdix rufa), and gray partridge (Perdix perdix) were studied by
Lutz-Ostertag and Lutz (1970). Eggs from these species were sprayed with
2,4-D at a concentration of 1 to 2 1/ha. They found the chemical to be toxic
to embryos prior to the 19th day. In the majority of cases, surviving
embryos suffered partial or total paralysis and exhibited lordosis, atrophied
feet, and shriveled toes which remained clenched. Histological and cytol-
gical examination of the surviving embryos indicated sterility of half of the
males and females.
Cope et al. (1970) studied chronic effects of 2,4-D on bluegill sunfish.
Hematocrit values were highest at highest treatment concentrations but re-
turned to normal after 3 days. Histopathological changes were noted
primarily in the liver as parenchymal cells underwent shrinkage and loss of
vacuolation. A decrease in glycogen was also noted. They reported that
changes were most accentuated at the 14th day and were common at day 28 in
the 5-10 ppm treatment lots. By day 56, cyological changes were still
evident in the 10 ppm lot, but glycogen stores were nearly normal.
Concomitant with the liver changes were also vascular abnormalities.
Eosiniphilic deposits appeared throughout the vascular system. At least 30%
of fish from the 0.1 ppm, 50% from the 0.5 ppm, 60% from the 1.0 ppm, and all
fish from the 5 and 10 ppm treatments showed these symptoms at 72 hours.
These deposits were present in 30% of the fish in the higher dosage at day
28, but were not found after 28 days. Circulatory stasis also occurred.
After 84 days only occasional fish showed slightly increased capillary
networks within the brain. Although a single exposure to 2,4-D in this study
produced pathologic changes, all fish eventually returned to normal. As the
authors point out, responses to repeated exposures and responses of more
sensitive species are not known.
Eller (1968) studied effects of Hydrothol 191 on bluegills in Oklahoma
ponds. He found morpheme trie and systematic changes in gills, liver, and
testes. Gill filaments were fused on contact with the chemical at 0.3 ppm.
After 14 days, gill structure returned to normal. Liver damage was reduced
after 56 days and returned to normal by day 112. Hypertrophic cells appeared
in the testes at the 3rd day and disappeared from fish from the 14th to 28th
day. Blood dyscrasia occurred in fish from ponds treated with 0.5 ppm and
existed for 36 days.
These studies indicate that fish suffer pathological changes when ex-
posed to herbicides. These changes may be severe enough to cause death or
physiological changes that last for a period of time and then recover. A
question can be posed as to whether fish exposed to subacute levels contin-
uously would eventually die. That possibility exists.
97
-------
Chemicals added to the water might also influence fish behavior. Folmar
(1976) exposed rainbow trout (Salmo gairdneri) to concentrations of various
herbicides. Trout detected sub-lethal concentrations and avoided the treat-
ment area. The lowest concentrations of each chemical avoided were 0.0001
mg/1 CuSCL, 0.1 mg/1 xylene and acrolein, 1.0 mg/1 dalapon and 2,4-D, and
10.1 mg/1 glyphosate, dipotassium salt of endothall, and diquat. In most
cases, fish detected these herbicides at below toxic concentrations. The
author postulated that although the chemicals were not toxic to fish, they
could have influenced habitat selection.
Another factor that should be considered as a chronic or long-term
effect is the influence chemical control could have on food chains. From the
previous discussion pertaining to acute toxicity and LD50, certain zoo-
plankton have tolerance levels which could make them susceptible to herb
icide treatments. Reduction or eradication of these organisms at a time when
larva fishes are dependent upon them for food could seriously affect a year
class and cause a shift in fish species. It seems logical,. therefore, to
determine food webs and to determine the reliance of fishes on species that
could be intolerant to herbicidal chemicals.
Plant Destruction
Data presented in this paper indicate that aquatic herbicides are rela-
tively non-toxic to aquatic animals and their residues in water and soils are
short-lived. Water quality parameters are usually not affected and if dif-
ferences are seen, they are not long-lasting. The most detrimental effect
could be establishment of algal "blooms. These blooms, the result of nutrient
release by decaying weeds, are an indirect effect of the treatment. The
degree of severity depends upon the nutrients within the system and the
amount of weeds killed. Literature data indicate that algal blooms do not
create problems in large systems or systems through which water flows.
Problems occur in small systems such as aquaria, swimming pools, and small
ponds where there is very little water circulation and redeposition of
nutrients in the soil. Carter and Hestand (1977) investigated effects of
various herbicide formulations in plastic pools on phytoplankton succession.
They concluded that liquid herbicides caused a fast kill of plants and
release of nutrients, resulting in an explosive growth of phytoplankton and
bacteria (Table 36). System E (a pelleted form of endothall) was the least
detrimental of the herbicides. Due to its slow release, phytoplankton growth
was gradual. In all cases, except System E, phytoplankton and bacteria
populations returned to normal at approximately 84 days after treatment.
Further evidence for the nonimpact of herbicides on aquatic animals is
seen in a pond study by Simpson and Pimental (1972). The authors used 12
0.1 ha ponds treated and untreated with Fenac. Short-term effects were de-
pression of dissolved oxygen concentration, an increase in community respir-
ation, and reduction of carbonate ion one month after treatment. These
effects were short-lived and did not seem to affect phytoplankton and
zooplankton populations. The authors stated that this particular chemical
did not deter algal growth, particularly Chara, which may have buffered the
system from drastic change.
98
-------
Table 36. ,
Toxicity of copper to marine and aquatic life. From Anon., 1973. /
-V"'-
, ,n* * I ,.'
]'''M!'"'<- !"'''
l/v'/,.-
Acute dose
96 hr LC50
0.57 mg/l (2 hrl
3 85 mg/l (2 hr)
051 mg/l 12 hr)
2.9 mg/l (2 hrl
22.5 mg/l (2 hrl
0.4-0.5 ppm
12 day)
1 25 ppm
1 .04 ppm
26 0 ppm
5.2 ppm
52 ppm
430 mg/l
470 mg/l
1.0 ppm (6.5 day)
0.23 mg 1 16 hrl
0 46 mg/l (6 hr)
3.3 mg/l (24 hr)
0.74 ppm
7.0 mg/l (48 hr)
0.18 ppm
34 ppm (2 day)
75 mg/l
Species
Wanersipora
Bugula
Spirorbis
Galeolaria
Mytilus
Salmo gairdneri
Lepomis
macrpchirus
Ldpomis
macrochirus
"
-
adult minnows
Pimephales
Gasterosteus
aculeatus
Balanus balanoides
" crenatus
Orizias
Lepomis
macrochirus
"
Pimephales
promelas
Gambusia a"ffinis
Gambusia affinis
56 000 ppm (2 day) Gambusia affinis
38 ppm (1 day)
1.25 mg/l (lime
not given)
48 hr.
1.9 ppm
1.9 mg/l
1.4 ppm
0.05 ppm
10 ppm
0.2 ppm
1.9 mg/l
0.40 ppm
Salmo gairdneri
(fry)
Lepomis
macrochirus
Daphnia magna
Japanese oyster
oysters
Pimephales
promelas
"
Lepomis
marcrochirus
"
oysters
Limnodrilus hoff-
Conditions
copper sodium citrate
pH 7.0-8.2
copper sodium citrate
pH 7 0-8.2
copper sodium citrate
pH 7 0-8.2
copper sodium citrate
pH 7 0-8.2
copper sodium citrate
pH 7 2-8.2
static acute bioasssy:
a.c.d.e.f
static acute bioassy.
a.c.d.e; Cu**; fish
acclimatized 2 wks. in
syn. dil. water.
static .acute bioassay;
a.c.d.e: fish acclima-
tized 2 wks in syn.
dil. water copper-
copper acelic acid: all
fish acclimatized 2
wks. in syn. dil water.
a.c.d.e: static acute bio-
assay same as above
except that" copper-
acetaldehyde was
used.
same as above except
that acetone: copper
mixture was used
static test
Literature
Citation*
Wisely and Blick
1967'"
Wisely and Blick
1967'"
"
Brown 1968"
'
Cairns Jr. and
Scheier 1968"
-
Cairns Jr. and
Scheier 1968" .
"
Mount 1968"
Constit- Acute dose
uent 96 hr LC50
Copper /
(Cul
0 425 ppm
0.27 ppm
1.5 mg/l (2-3 d)
0.27 ppm
0.050 ppm
0.56 ppm (1 day)
90 ppm (1 day)
ISppmd day)
10-ppm (2 days)
5 ppm (3 days)
20 ppm (3 day)
40 ppm (1 day)
2 ppm (1 day)
0.1 ppm
2 ppm (2 hr)
0.2 ppm (48 hr)
1.5 ppm
19 ppm (12 days)
Species (
meisteri
Gyraulus
circumstriatus
Physa
heterostropha
Nereis
Conditions
hard water CuSo.
a.c.d.i
static acute bioassy;
a.c.d.i: hard water:
CuSO.
same as above
Physa heterostropha 21 C hard water as
"
*
Carassius auratus
Poecilia reticulata
toad and frog
tadpoles
"
"
Dragon fly larvae
daphnia longispina
Nereis virens
Salmo gairdneri
"
CuSO.
same as above: young
static acute bioassy:
a.c.f: CuSO. hard and
soft water.
cone, as copper sulfite
Cone, as copper sulfite
cone, as copper sulfite
M
time not specified
CuSO. 5HjO
"
Gammarus lacustris static acute bioassay:
Nereis virens
a.e. CuSo.
time not specified
continuous flow bioassay Mount 1968"
static acute bioassay;
a.c: using Cu(NOi)i
hypertonic seawater
hypertonic seawater
CuCI,2HjO
static acute bioassay;
a.c.d.e. distilled
aerated water
20 C; pH 8.3
static acute bioassay:
a.c.d.e.f: CuSO.
static acute bioassay:
turbid water:
a.c.d.a.g; CuSO.
24-27 C: using copper
sulfate in highly
turbid water
cupric oxide: static acute
bioassay a.c.d.e.g:
turbid water 19-20 C
CuSO-: a.c.e.f.i.p:
static acute bioassay
in soft water: 18-20 C:
Cuds
Copper sulfate
pH8.2; 12 C
static acute bioassay:
a.c.d.f. hard water:
CuSO.
same as above using
soft water
same as above using
hard water
same as above using
soft water
CuCI, 2HjO_
static acute bioassayl
Jones 1938"
Pyefinch and Mott
1948"" *
-
Ooudoroff and
Katz 1953"
Trama 1954aIM
Turnbull at al.
1954'30
Palmar and Ma-
loney 195590
Wallen at al.
1957'"
Wallen at al.
1957'"
Wallen et al.
1957'"
Turnbull-Kemp
1958'3'
Academy of Nat-
ural Sciences
I9601
Cataejszek and
Stasiak 1960"
Fujiya I96013
Fujiya 1960"
1961"
Tarzwell and Hen-
derson 1960'"
"
"
"
Fujiya 1961"
Wurtz and Bridges
0.980 ppm
2.8 ppm
0.8 ppm (2 day)
0.034 ppm (1 day)
32 0g/l (time not
given)
0.1 50 ppm (2 day)
2.800 ppm (2 day)
1.5 ppm
1.2 ppm
1.14 mg/l
10.2_mg/l
0.048 ppm
3.0 ppm
10 ppm (1 day)
1.0 ppm (6 day)
1.0 ppm (6 day)
0.25 mg/l
7
Lepomis
macrochirus
Lepomis
macrochirus
Salmo gairdneri
Salmo salar
juvenile salmon
Salmo gairdneri
Lepomis
macrochirus
Pimephales
promelas
"
Pimelometopon
pulchrum
Lepomis
macrochirus
Salmo salar
CuCli
static acute bioassay;
CuSO.a
a.c.e.f.l.m: field study
in a river
continuous flow, acute
bioassay g.c.f: with
3 w 1 Zn and 2 pg/l
Cu
in very soft water
(14 mg 1 hardness)
static acute bioassav. a.
CuSO.
same as above
as CN using copper
cyanide complex;
static acute bioassay:
a.c.: soft water
same as above except
cone, as Cu
in hard water: CuSO.
5H,O
in hard water
BSA: a: incipient lethal
level with 0 600 Zn
Orconectes rusticus continuous flow acute
"
Oroconectes rusti-
cus embryo
bioassay. a.c.e.f:
20 C; intermolting
stage
same as above: adult
crayfish used
same as above; juvenile
crayfish used
same as above: re-
cently hatched young
which remained cling-
ing to pleopods of
female during 1st
molt were used.
time not given
Literature
citation*
1961 J
"
"
Raymount and
Shields 1962
Wurtz 1 962
"
Wurtz 1962
Rochet al. 1963
Floch et al. 1963
"
Floch et al. 1963
*
w
Raymount and
Shields 1963
Herbert and Van
Oyke 1964
Nebeker and Gaufin
1964
Raymount and
Shields 1963
Cope 1965
"
Herbe.' "' al.
1965
Sprague 1965
Sprague and Ram-
sey 1965
Cope 1 966
Cope 1966
OoudoroH et al.
1966
"
Pickering and Hen-
derson 1966
Sigleretat. 1966
Hubschman 1967
"
,
"
Hubschman 1967
Of-l-' f J
y.t-i
99
-------
Table 36.
Continued.
Acute dose
Constitutent 96 hr LC50
84.0^g/l
75/yg/l
0.795-0.815 ppm
(5 day)
1 .25 ppm
0.2 mg/l (48 hr>
30 mg/l (48 hr)
100 mg/l
1 mg/l
430 ug/\
470 i*a/\
1.7 mg/l
0.039 mg/l
0.20 mg/l
48 hr
3.2 mg/l
Species
Pimelometopon
pulchrum
Nitzschia linearis
Lepomis
macrochirus
Penaeus duorarum
Penaeus aztecus
shore crab '
cockle
Pimelometopon
pulchrum
"
Capeloma decisum
Physa Integra
Gammarus pseudo-
limnaeus
Salmo gairdneri
Fundulus
heteroclitus
Conditions
soft water: static bio-
assay
" continuous flow
bioassay
static acute bioassay.
a.c.e; CuCI2
same as above
in the dark: 15 C;
CuSO.
"
"
"
static bioassay: hard
water
continuous flow bio-
assay: hard water
soft water
soft water
soft water
20-22 C. no feeding
during the 96 hrs.
aerated water
i '
Literature citation"
Mount and Stephen
1969
"
Patrick et al.
1958
Portmann 1968
"
"
"
Mount and Stephen
1969
Mount and Stephen
1969
Arthur and Leon-
ard 1970
Arthur and Leon-
ard 1970
Brown and Gallon
197C
Jackim et al.
1970
L= Laboratory bioassay
BS = bioassay static
BCF= bioassay continuous flow
BA= bioassay acute
BCH= bioassay chronic
a = water temperature
b= ambient air temperature
c=pH
d= alkalinity (total, phenolphthalein or caustic)
e= dissolved oxygen
f = hardness (total, carbonate, Mg or CaO)
g = turbidity
h= oxidation reduction potential
i= chloride as Cl
j= BOD, 5 day; (J) = BOD, short-term
k=COD
l= Nitrogen (as N02 or NOs)
m= ammonia nitrogen as NHs
n= phosphate (total, ortho-, or poly)
o= solids (total, fixed, volatile, or suspended)
p=C02
BOD = biochemical oxygen demand
100
-------
HERBICIDE
APPLICATION
PLANT DEATH
Breakdown of
plant material
Increased
respiration
Loss of substrate
for attachment.
production etc
\
Changn in oxygen carbon
dioxide balance
Increased light
penetration
Increased
turbulence
Decreased light penetration
I
REPLACEMENT MACRO-
OR MICRO-FLORA
Release of inorganic
nutrients
Detritus
Restoration of
oxvgen-carbon
dioxide balance
CHANGE IN STATUS OF
ANIMAL POPULATIONS
Food attachment
sues etc
Figure 35. Effects of herbicide application and the destruction of
submerged plants likely to be of consequence in determin-
ing faunal changes. From Brooker and Edwards, ]975.
101
-------
Destruction of aquatic vegetation can lead to continued phytoplankton
blooms as evidenced by Lake Apopka, Florida. This large lake (12,500 ha,
average depth approximately 2m) was formerly a clear-water lake with luxur-
iant vegetation and was noted for excellent bass fishing. At the present
time, the lake experiences continued algal blooms (EPA 1979). Several causes
are given for this change: 1) nutrient overloading; 2) vegetation removal in
1947 by a hurricane; 3) hyacinth spraying and fish poisonings which added
nutrients; and 4) water level stabilization. The first algal bloom was noted
after vegetation was removed in 1947. The buffering capacity (nutrient
removal and tie-up) was removed when the plants no longer existed. It is
interesting to note that one of the proposed beneficial effects after lake
drawdown will be the establishment of macrophyte growth. These plants will
compete with phytoplankton for nutrients and thus reduce algal blooms (EPA
1979).
Brooker and Edwards (1975), in a review paper, discussed changes that
can occur after aquatic weeds are killed (Figure 35). Most changes are due
to decay and disappearance of the weeds. These changes can be "good" or
"bad," depending upon the user group. The user group, however, often changes
their attitude once a change is made and becomes permanent.
Fish populations can be affected by removal of macrophytes. Many fish
depend upon organisms that are associated with vegetation, while others are
dependent upon vegetation for spawning sites. Data, however, pertaining to
changes in fish productivity are conflicting when a water body is treated
with herbicides. Bennett (1971) reported data that indicated productivity
may increase or decrease, depending on the situation.
Fishery biologists, especially in Florida, agree that vegetation is
needed to maintain good fish populations, but the optimal amount of vege-
tation needed is not known. Data from Orange Lake, Florida, indicate that
when 80% of the lake was covered with hydrilla, many small fish abounded in
the vegetation. When the vegetation decreased the next year, fewer small
fish were present, but growth and population structures improved (Shireman
and Haller 1979). Total vegetation infestation of the water column probably
exerts a physical constraint on foraging efficiency by eliminating the gra-
dient or ecotone between open water and submersed macrophytes. Colle and
Shireman (1980) found that bluegill and redear sunfish condition factors and
weight-length relationships were not adversely affected until hydrilla occu-
pied the majority of the water column. Elimination of the forage gradient
was given as a possible cause. Harvestable largemouth bass exhibited low
condition values once hydrilla coverage was above 30%; however, smaller bass
were not adversely affected until coverage exceeded 50%. Shireman et al.
(1979) compared condition factor values for bluegill and redear sunfish from
Lake Wales during two years of heavy hydrilla infestation with a central 50%
range of K(TL)'s from Carlander (1977). Condition for both bluegills and
redears only approached or exceeded Carlander's mean value during 1977 when
hydrilla coverage was reduced.
102
-------
Mechanical Harvesting
Removal of excessive aquatic plant growth from the aquatic environment
often enhances the usefulness of the water. Mechanical harvesting is often
considered the most environmentally safe method of removing aquatic plants,
particularly aquatic macrophytes, and has been considered as an important
method for removing nutrients (McNabb and Tiemey 372). Burton et al.
(1979) upon reviewing data from a range of lakes suggest that aquatic plant
harvesting has very little effect on reducing the nutrient content of many
lakes. In fact, they noted that recent research has shown that many rooted
aquatic macrophytes extract nutrients from the sediments (McRoy and Barsdate
1970; Bristow and Whitcomb 1971; DeMarte and Hartman 1974; Bole and Allan
1978) and that cutting or damaging the stems allows the macrophytes to pump
nutrients to the overlying water (DeMarte and Hartman 1974; Carpenter and
Adams 1977; Carpenter and Gasith 1978; Bole and Allan 1978), thus increasing
plant nutrient concentrations. This nutrient release often leads to develop-
ment of phytoplankton blooms or to growth of filamentous algae (Nichols 1973;
Burton et al. 1979). Decay of plant material not removed from the water
during harvesting can also release nutrients to the water column (Jewell
1971; Nichols 1973) and can further support algal growth.
Harvesting macrophytes can also affect the aquatic environment in other
ways. Removal of rooted macrophytes destablizes bottom sediments, thus al-
lowing resuspension of bottom sediments. This not only increases turbidity,
but increases the rate of erosion in the littoral zone. Resuspension of
organic materials can also decrease oxygen concentration in the water column
(Jewell 1971). Harvesting may also result in diminished growth of macro-
phytes in subsequent years (Neel et al. 1973; Nichols 1974) by decreasing
rate of input of detrital organic matter which is an important food source
for many fishes and benthic organisms. Reduced plant growth will also result
in a loss of substrate for attached algae and cover for invertebrates and
fish (Burton et al. 1979).
Biological Control
Management of aquatic plant growth, through use of living organisms,
offers a useful approach to managing the consequences of nutrient enrichment.
Shapiro (1979) suggested that phytoplankton populations could be reduced by
manipulations of fish populations. Shapiro cited studies that indicated that
loss of fish populations through either winter kill (Schindler and Comita
1972) or chemical removal (Bandow 1978) often resulted in significant reduc-
tion in algal biomass with a corresponding increase in water transparency.
Decline in algal biomass nearly always corresponds to large increases in
Daphnia populations. Large Daphnia seem to remove algae through grazing
(Burns 1969). However, Bandow (1978) noted that decline in algal abundance
could be attributed to decrease in bottom-feeding fish that pump nutrients to
the overlying water during feeding (Lamarra 1975a, 1975b).
103
-------
Shapiro (1979) suggested that complete removal of fish populations was
probably not necessary to obtain reduction in algal populations. Brooks and
Dodson (1965) showed that zooplanktivorous fish select larger zooplankton
during feeding. Shapiro (1979) suggested selective reduction in population
size of zooplanktivorous fish through use of chemicals or the stocking of
large predators that could reduce predation on the larger forms of zooplank-
ton. This would permit these zooplankton to exert greater grazing pressure
on algae.
Fish can also exert a degree of control on aquatic macrophytes. Rose
and Moen (1953) reported that large populations of carp (Cyprinus carpio),
buffalo (Ictiobus spp.), and sheepshead (Aplodinotus grunniens) suppressed
growth of macrophytes in Lake East Okoboji, Iowa. They noted that after
removal of 2,726 Ib rough fish/acre, biomass of bottom vegetation increased.
The feeding activities of the fish apparently uprooted vegetation. Recent
studies by Lamarra (1975a, 1975b) suggest that carp pump nutrients which can
support algal growth into the waters. Since Lake East Okoboji had large
numbers of blue-green algae, shading by algae might also have contributed to
reduction in plant growth. In fact, Smith and Swingle (1941a, 1941b) rec-
ommended fertilization of waters to reduce aquatic macrophyte growth.
Perhaps one of the most promising fish for management of aquatic macro-
phytes is the grass carp (Ctenopharngyodon idella). Possible introduction of
this fish, however, has caused considerable concern among many biologists. A
main concern is that the introduction of non-native fish will reduce popula-
tion size of native fish. Although Rose and Moen (1953) noted that common
carp caused reduction in game fish, grass carp normally do not spawn in
lakes, and thus population size can be controlled through stocking. These
fish, however, eat large amounts of vegetation and 100% removal of vegetation
could have important indirect effects on the aquatic environment. Grass carp
can release plant nutrients through feeding (Hickling 1966; Michewicz et al.
1972) and it is possible that very rapid removal of large amounts of macro-
phytes by these fish could result in algal blooms. If fish populations are
controlled, however, algal blooms probably will not occur. If algal blooms
occur, they might reduce populations of sight-feeding fish. Most studies,
however, indicate there is no direct competition between grass carp and other
fish (Kilgen and Smitherman 1971; Rottmann 1976).
The effect of grass carp -on water quality has not been fully investi-
gated. Studies by Lembi et al. (1978) in ponds indicate certain water
quality changes. Dissolved potassium concentration increased in all ponds
stocked with the fish and appears to be the water quality parameter most
sensitive to vegetation alteration. Nitrogen and phosphorus concentrations
were not statistically different after stocking. Shireman et al. (1979)
found similar results in a 300-acre Florida lake stocked with grass carp.
Potassium concentration increased three to four times when vegetation biomass
was reduced. Studies (Michewicz et al. 1972) on small enclosures stocked
with grass carp have shown an increase in water hardness and nitrogen but not
in phosphorus. Studies by Haller and Sutton (1976), however, reported an
increase in phosphorus during rapid hydrilla control by grass carp. They did
not, however, detect changes in water turbidity. Other authors (Terell 1974;
104
-------
Forrester and Lawrence 1978) reported no changes in water quality after grass
carp were stocked. Terrell (1974) found significant increase in sediment
concentrations of iron, magnesium, and orthophosphate; calcium and magnese
did not change.
Research Needs for Development of Effective, Environmentally Safe Aquatic
Plant Management Programs.
Private citizens, water resource managers, politicians, some aquatic re-
search scientists, and others, after examining the literature review in this
paper, might argue that there is no need for additional research as an over-
abundant accumulation of facts already exists. They will agree that all
monies should be expended on aquatic plant control. We agree that there is
an embarassingly large accumulation of data and that, if sufficient money is
available, we have the knowledge and techniques to kill aquatic plants. Most
aquatic weeds, however, are controlled by chemicals that environmental groups
and the public at large find unacceptable. . It is necessary to provide alter-
native control methodology and to document any potential hazard in using
certain chemicals in water. An outright ban on chemical control of weeds
would create health, recreational, economic, and flood problems. Development
of environmentally safe, effective aquatic plant management programs will
require a better understanding of how aquatic systems function. The major
research need, therefore, is the advancement of our knowledge concerning
functioning of aquatic systems influenced by aquatic management techniques.
If, with the accumulated data now available, we still need additional
information before we can develop effective, yet safe aquatic plant manage-
ment programs, the problems will involve how to advance our knowledge. Cer-
tainly, a poll of research scientists would show that there are numerous re-
search questions, and, hence, reseach projects that should be funded.
We feel that the following questions should be considered and research-
ed:
Plants and Water Quality
1. Are current river and lake classification systems appropriate? Should
separate classifications be established for specific user groups?
Should we attempt to provide multiple-use waters? Should separate clas-
sifications be established for aquatic plant management? Water bodies,
in most cases, have a primary use: for example, canals for water trans-
port, reservoirs for potable water, fish management lakes, and suburban
lakes for general recreational uses. These water bodies, because of
their various functions, probably need different management programs
that may be mutually exclusive.
105
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2. What are the standards against which water quality and aquatic weed
nuisances should be judged? User groups often establish their own
standards without regard for biological or ecological principles.
3. Are water quality measurements important as impact measurements? How
does water quality change with changes in the natural physical, chemi-
cal, and biological characteristics of aquatic systems? Methods must
be developed for evaluation of natural water quality changes versus
changes caused by management programs.
4. What are the relationships between plant nutrient concentrations in
lakes and subsequent aquatic macrophyte biomass?
5. Will the degree of an algal bloom and its persistence depend on the lake
trophic state? Can we expect all lakes to react in a similar manner
regardless of trophic state?
6. Are there critical algal and aquatic macrophyte concentrations that can
be achieved to minimize problems in a given lake? Can a balance be
reached where neither of these plants causes problems?
7. What physical, chemical, and biological factors control total aquatic
plant growth and growth of specific species of aquatic plants? Can
natural and/or anthropogenic factors, such as nutrient inputs, be manipu-
lated to enhance growth of desirable species and discourage growth of
non-beneficial aquatic plants?
Aquatic Organisms
8. What is the effect of aquatic plant biomass on water quality,
zooplankton, benthos, and fish?
9. At what biomass do aquatic plants, such as hydrilla or blue-green algae,
become detrimental to fish populations?
10. Can we manage problem aquatic macrophytes, such as hydrilla, or benefi-
cial plants, such as panicum, to satisfy maximum fish and wildlife needs?
11. Are there certain species of aquatic plants, such as panicum or green
algae, that can be managed to enhance fish populations.?
12. Can research develop genetically superior and hardier aquatic plants?
Emergent or floating plants in certain situations might be more bene-
ficial than submersed plants. Genetically superior plants might compete
more effectively for nutrients, causing a reduction in less desirable
plants. Plants could be selected to enhance fish production.
106
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Biological Control
13. Should there be a moratorium on introduction of potential biological
control organisms until better methods are developed and the potential
behavior of these organisms in U.S. ecosystems can be predicted? More
exhaustive searches should be made to determine if biological control
organisms are present in U.S. waters. Evaluation should be made to
determine effectiveness of existing biological controls for differ-
ent management programs. It is our feeling that many of the current
management strategies have not been developed for their use.
14. Can grass carp be effectively managed alone to control vegetation?
Data indicate that considerable time and effort must be expended to
assure that desired control is achieved. Grass carp stocked alone
often do not achieve desired results.
15. Can we begin ichthyofauna reconstruction in our nation's waters to
enhance aquatic plant management programs and thus control aquatic
problems with minimum energy inputs and environmental degradation?
Herbivores and other fishes that feed on algae or other organisms
could be utilized to reconstruct entire systems.
Chemical Control
16. Considering there is a spectrum of toxicity and persistence, is it
best to use a persistent chemical once every few years or a non-
persistent herbicide 3 or 4 times yearly? The persistent herbicide
would reduce the volume of weeds killed each year and would probably
decrease environmental problems such as those associated with weed
decay.
17. What are the short-term and long-term impacts of chemical usage in
aquatic systems? Are there long-term chronic effects on non-target
organisms that are not detected by current bioassay techniques? Is
the continued use of non-persistent, non-toxic herbicides detrimental
to fish and other aquatic organisms?
18. What is the best time to apply herbicides to control aquatic weeds
and reduce impacts on non-target organisms? Can treatments be timed
to allow organisms to reproduce and grow?
19. What is the selectivity of various herbicides to specific aquatic
plants? Selective herbicides would allow other plants to grow, thus
eliminating the problem of effects on all plant species. As previous
discussions indicate, complete eradication of aquatic weeds is detri-
mental to the aquatic system.
20. Are laboratory bioassay procedures relevant to impact observed on aqua-
tic systems? It appears to us that laboratory bioassay tests cannot
predict what a chemical will do in the lake or river system.
107
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21. Is the review process for labelling herbicides consistent with main-
taining effective herbicides on the market? The cost of review is
very expensive and might discourage chemical producers from developing
new and more Affective herbicides.
22. Are there aquatic plant growth regulators that could be applicable to
the management of aquatic plants? Growth regulators are currently in
use in terrestrial systems to retard plant growth. It should be
determined if such regulators can be used in aquatic systems.
Mechanical Control
23. Mechanical control devices have been developed that can be used in
certain situations and sites. Evaluations must be made to determine
their economic and energy costs under different conditions and whether
these methods are effective.
Integrated Control
24. What is the best integrative control method approach utilizing current
technology? For example, how should various chemical, biological,
physical, mechanical, and environmental manipulations be mixed to
achieve optimal aquatic plant management?
25. Can we develop integrated mechanical, chemical, and biological control
techniques that have acceptable cost-benefit ratios?
Miscellaneous
26. What are acceptable cost-benefit ratios for aquatic plant management?
How should cost-benefit ratios be determined? Should public monies
be spent on private water or minimize potential infestation of public
waters with problem plants?
27. What are the social and economic impacts of alternative aquatic control
techniques?
28. Should public monies be spent for development of control techniques?
If a chemical or other technique shows promise as a beneficial method,
should the research and development of this technique be funded with
public monies?
108
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Aquatic weeds can currently be controlled in most aquatic habitats with
one or a combination of chemical, mechanical, biological, or habitat manipu-
lation techniques. For the most part, the technique and extent of weed con-
trol are decided on the basis of water usage and cost. Environmental effects
are often evanescent and result from removal of weeds or changing the habitat
from littoral to limnetic.
The above list provides a basis for establishment of research priorities
and programs. It is not intended to be complete and cover all aspects of
aquatic weed research, but we feel it includes the major questions. This
list, however, should form a basis for discussion that will lead to estab-
lishment of research needs and priorities and sound research programs.
109
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