TECHNICAL SUPPORT MANUAL:
WATERBODY SURVEYS AND ASSESSMENTS FOR
CONDUCTING USE ATTAINABILITY ANALYSES
VOLUME III: LAKE SYSTEMS
U.S. ENVIRONMENTAL PROTECTION AGENCY
OFFICE OF WATER REGULATIONS AND STANDARDS
CRITERIA AND STANDARDS DIVISION
WASHINGTON, D.C. 20460
NOVEMBER 1984
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FOREWORD
The Technical Support Manual: Water Body Surveys and Assessments for
Conducting Use Attainability Analyses. Volume III; Lake Systems contains
guidance prepared by EPA to assist States in implementing the revised Water
Quality Standards Regulation (48 FR 51400, November 8, 1983). This docu-
ment addresses the unique characteristics of lake systems and supplements
the two previous Manuals for conducting use attainability analyses (U.S.
EPA, 1983j>, 1984). The purpose of these documents is to provide guidance
to assist States in answering three central questions:
(1) What are the aquatic protection uses currently being achieved in
the water body?
(2) What are the potential uses that can be attained based on the
physical, chemical and biological characteristics of the water
body?
(3) What are the causes of any impairment of the uses?
Consideration of the suitability of a water body for attaining a given use
is an integral part of the water quality standards review and revision
process. EPA will continue to provide guidance and technical assistance to
the States in order to improve the scientific and technical bases of water
quality decisions. States are encouraged to consult with EPA at the
beginning of any standards revision project to agree on appropriate methods
before the analyses are initiated, and to consult frequently as they are
conducted.
Any questions on this guidance may be directed to the water quality
standards coordinators located in each of the EPA Regional offices or to:
Elliot Lomnitz
Criteria and Standards Division (WH-585)
401 M Street, S.W.
Washington, D.C. 20460
Edwin L. Johnson, Director
Water Regulations and Standards
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CONTENTS
FOREWORD
CHAPTER I INTRODUCTION 1-1
CHAPTER II PHYSICAL AND CHEMICAL CHARACTERISTICS .11-1
INTRODUCTION II-l
PHYSICAL CHARACTERISTICS II-l
Physical Parameters II-l
Physical Processes II-6
CHEMICAL CHARACTERISTICS 11-23
Overview of Physico-chemical Phenomena in Lakes 11-23
Phosphorus Removal by Precipitation 11-27
Dissolved Oxygen 11-28
•Eutrophication and Nutrient Cycling 11-29
Significance of Chemical Phenomena to Use
Attainability 11-31
TECHNIQUES FOR USE ATTAINABILITY EVALUATIONS 11-32
Introduction 11-32
Empirical Models 11-33
Computar Models 11-48
CHAPTER III BIOLOGICAL CHARACTERISTICS III-l
INTRODUCTION III-l
PLANKTON III-l
Phytoplankton III-l
Zooplankton I11-10
AQUATIC MACROPHYTES I11-11
Response to Macrophytes to Environmental Change 111-11
Preferred Conditions II1-12
BENTHOS 111-13
Composition of Benthic Communities 111-13
General Response to Environmental Change 111-14
Qualitative Response to Environmental Change 111-14
Quantitative Response to Environmental Change III-22
FISH 111-31
Trophic State Effects II1-31
Temperature Effects 111-32
Specific Habitat Requirements II1-32
Stocking I11-34
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CHAPTER IV SYNTHESIS AND INTERPRETATION
IV-1
CHAPTER V
APPENDIX A
APPENDIX B
APPENDIX C
APPENDIX D
INTRODUCTION IV-1
USE CLASSIFICATIONS IY-1
REFERENCE SITES IV-4
Selection IV-4
Comparison IV-7
CURRENT AQUATIC LIFE PROTECTION USES IY-8
CAUSES OF IMPAIRMENT OF AQUATIC LIFE PROTECTION USES IY-8
ATTAINABLE AQUATIC LIFE PROTECTION USES IY-8
PREVENTIVE AND REMEDIAL TECHNIQUES IV-10
Dredging iy-11
Nutrient Precipitation and Inactivation IV-16
Aeration/Circulation IY-22
Lake Drawdown IV-30
Additional In-Lake Treatment Techniques IY-34
Watershed Management IY-39
REFERENCES Y-l
PALMER'S LISTS OF POLLUTION TOLERANT ALGAE A-l
U.S. ENVIRONMENTAL PROTECTION AGENCY'S PHYTOPLANKTON
TROPHIC INDICES B-l
CLASSIFICATION, BY VARIOUS AUTHORS, OF THE TOLERANCE
OF VARIOUS MACROINVERTEBRATE TAXA TO DECOMPOSABLE
WASTES C-l
KEY TO CHIRONOMID ASSOCIATIONS OF THE PROFUNDAL ZONES
OF PALEARCTIC AND NEARCTIC LAKES 0-1
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CHAPTER I
INTRODUCTION
EPA's Office of Water Regulations and Standards has prepared guidance to
accompany changes to the Water Quality Standards Regulation (48 FR 51400).
This guidance has been compiled and published in the Water Quality Stand-
ards Handbook (U.S. EPA, December 1983^). Sections in the Handbook present
discussion of the water quality review and revision process; general
guidance on mixing zones, and economic considerations pertinent to a change
in the use designation of a water body; the development of site specific
criteria; and the elements of a use attainability analysis.
One of the major pieces of guidance in the Handbook is "Water Body Surveys
and Assessments for Conducting Use Attainability Analyses." This guidance
presents a general framework for designing and conducting a water body sur-
vey whose objective is to answer the following questions:
1. .What are the aquatic life uses currently being achieved in the
water body?
2. What are the potential uses that can be obtained, based on the
physical, chemical and biological characteristics of the water
body?
3. What are the causes of impairment of the uses?
In response to requests from several states for additional information,
technical guidance on conducting water body surveys and assessments has
been provided in two documents:
1. Technical Support Manual; Water Body Surveys and Assessments for
Conducting Use Attainability Analyses (U.S. EPA. November 1983.b);
2. Technical Support Manual; Water Body Surveys and Assessments for
Conducting Use Attainability Analyses, Volume II: Estuarine'
Ifil
Us
Systems (U.STEPA, June 1984},
The first volume is oriented towards rivers and streams and presents
methods for freshwater evaluations. The second volume stresses those con-
siderations which are unique to the estuary. The current Manual, Volume
III; focuses on the physical, chemical and biological phenomena of lakes
and is presented so as not to repeat information that is common to other
freshwater systems that already appears in one of the earlier volumes.
Apart from the rare impoundment that is fed only by surface runoff or
underground springs, rivers and lakes are linked physically and exhibit a
transition from riverine habitat and conditions to lacustrine habitat and
conditions. Because of this physical link, the biota of the lake will be
essentially the same as the biota of the stream, although there are few
species that are primarily lake species. Given the ties that exist between
lake and stream under natural conditions, it is important that those who
will be conducting lake use attainability studies refer to Volume I on
rivers and streams for additional perspective.
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Each of the Technical Support Manuals provides extensive information on the
plants and animals characteristic of a given type of water body, and
provides a number of assessment techniques that will be helpful in per-
forming a water body survey. The methods offered in the guidance documents
are optional, however, and states may apply them selectively, or may use
their own techniques for designing and conducting use attainability
studies.
Consideration of the suitability of a water body for attaining a given use
is an integral part of the water quality standards review and revision
process. The data and other information assembled during the water body
survey provide a basis for evaluating whether or not the water body is
suitable for a particular use. Since the complexity of an aquatic eco-
system does not lend itself to simple evaluations, there is no single
formula or model that will serve to define attainable uses. Rather, many
evaluations must be performed, and the professional judgment of the
evaluator is crucial to the interpretation of data that is reviewed.
This Technical Support Manual on lakes will not tell the biologist or
engineer how to conduct a use attainability study, per se, rather, it will
lay out those chemical, physical and biological phenomena that are char-
acteristic of lakes, and point out factors that the investigator might take
into consideration while designing a use study, and while preparing an
assessment of uses from the information that has been assembled. The
chapters in this Manual focus on the following aspects of lakes:
Chapter II. Physical and Chemical Characteristics
o Circulation, stratification, seasonal turnover
o Nutrient cycling
o Eutrophication processes
o Computer and desktop procedures for lake evaluations
Chapter III. Biological Characteristics
o Benthos
o Zooplankton
o Phytopiankton
o Macrophytes
o Fi sh
Chapter IV. Synthesis and Interpretation
o Aquatic life use classifications
o Impairment of uses
o Reference site comparisons
o Preventive and remedial techniques
Chapter V. References
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CHAPTER II
PHYSICAL AND CHEMICAL CHARACTERISTICS
INTRODUCTION
The aquatic life uses of a lake are defined in reference to the plant and
animal life in the lake. The types and abundance of the biota are largely
determined by the physical and chemical characteristics of the lake. Other
contributing factors include location, climatological conditions, and
historical events affecting the lake.
Each lake characteristic such as depth, length, inflow rate and temperature
contributes to the physical processes of the water body. For example,
circulation may be the dominant physical process in a lake that is large
and shallow while for a deep medium size lake the dominant process may be
the annual cycle of thermal stratification.
The chemical characteristics of a lake are affected by inflow water quality
and by various physical, chemical and biological processes which provide
the biota with its sustaining nutrients and required dissolved oxygen.
Overenrichment with nutrients may accelerate the natural processes of the
lake, however, and lead to major upsets in plant growth patterns, dissolved
oxygen profiles, and plant and animal communities. The physical and
chemical attributes of lakes as well as the influence of physical processes
on chemical characteristics are discussed in this chapter.
In addition to a discussion of physical parameters and processes, and the
chemical characteristics of lakes, several techniques for use attainability
evaluations are presented in this chapter. These include empirical
input/output models, computer simulation models, and data evaluation
techniques. For each of these general categories specific methods and
models are presented with references. Illustrations of some techniques are
also presented.
The objective in discussing the physical and chemical properties of lakes
is to assist the states to characterize a lake and select assessment
methodologies that will enable the definition of attainable uses.
PHYSICAL CHARACTERISTICS
Physical Parameters
The physical parameters which describe the size, shape and flow regime of a
lake represent the basic characteristics which affect physical, chemical
and biological processes. As part of a use attainability analysis, the
physical parameters must be examined in order to understand non-water
quality factors which affect the lake's aquatic life.
Lakes can be grouped according to formation process. Ten major formation
processes presented by Wetzel (1975) include:
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o Tectonic (depression due to earth movement)
o Yolcanos
o Landslides
o Glaciers
o Solution (depressions from soluble rock)
o River activity
o Wind-formed basins
o Shoreline activity
o Dams (man-made or natural).
The origins of a lake determine its morphologic characteristics and
strongly influence the physical, chemical and biological conditions that
will prevail.
Physical (morphological) characteristics whose measurement may be of
importance to a water body survey include the following:
o Surface area, A (measured in units of length squared, L^)
o Volume, V (measured in units of length cubed, L )
o Inflow and outflow, Q. and Qn.It (measured in units of length
cubed per time, L3/T) 1n out
o Mean depth, 3
o Maximum depth
o Length
o Length of shoreline
o Depth-area relationships
o Depth-volume relationships
»
o Bathymetry (submerged contours).
Some of these parameters may be used to calculate other characteristics of
the lake. For example:
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o The mass flow rate of a chemical, say phosphorus, may be calcu-
lated as the product of concentration [P. ] and inflow, Q. , pro-
vided the units are compatible.
mass flow rate = [P.n, M/L3] x (Qin, L3/T) = M/T
where M denotes units of mass
o The surface loading rate is calculated as the quotient of inflow
and surface area, or the quotient of mass flow rate and area,
e.g.,
liquid surface loading rate = (Q.n, L3/T)/(A, L2) = L3/L2-T
mass surface loading rate = [Cin, M/L3] x (Qin, L3/T)/(A, L2) = M/L2-T
o The detention time is given by the quotient of volume and flow
rate, e.g.,
detention time = (V, L3)/(Q. , |_3/T) = T
The reciprocal of the detention time is the flushing rate, T"
o Mean depth is the quotient of volume and surface area, e.g.,
9 = (V, L3)/(A, L2) « L
The first seven parameters of the above list describe the general size and
shape of the lake. Mean depth has been used as an indicator of produc-
tivity (Wetzel, 1975; Cole, 1979) since shallower lakes tend to be more
productive. In contrast, deep and steep sided lakes tend to be less
productive.
Total lake volume and inflow and outflow rates are physical characteristics
which indirectly affect the lake aquatic community. Large inflows and
outflows for lakes with small volumes produce low detention times or high
flow through rates. Aquatic life under these conditions may be different
than when relatively small Inflows and outflows occur for a large lake
volume. In the latter case the detention time is much greater.
Hand (1975) has recommended a shape factor—the lake length divided by the
lake width—for lake studies. This shape factor was applied by Hand and
McClelland (1979) as a variable in a regression equation used to predict
chlorophyll-a in Florida lakes. Other parameters in that regression
equation are phosphorus, nitrogen, and the mean depth.
For the requirements of a more detailed lake analysis, information describ-
ing the depth-area and depth-volume relationships and information
describing the bathymetry may be required. An example of a bathymetric map
is shown in Figure II-1 for Lake Harney, Florida (Brezonik and Fox, 1976).
The roundness of this particular lake is typical of many lakes in Florida
whose morphometry has been affected by limestone solution processes (Baker,
et al., 1981). A typical representation of the depth-area and depth-volume
relationships for a lake is shown in the graph of Figure II-2 for the Fort
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EXPLANATION
Shaded area represents marsh area
Contour lines showing depth in feet at
mean low va«er
Figure II-l. Bathymetric Map of Lake Harney, Florida (from Brezonik, 1976)
II-4
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820
81 0
800
S 790
a
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Loudoun Reservoir, Tennessee (Hall, et al., 1976). Depth-area relation-
ships can be important to the biological activity in a lake. If the
relationship is such that with a slight increase in depth the surface area
is greatly increased, this then produces greater bottom and sediment con-
tact with the water volume which in turn could support increased biological
activity.
In addition to the physical parameters listed above, it is also important
to obtain and analyze information concerning the lake's contributing water-
shed. Two major parameters of concern are the drainage area of the con-
tributing watershed, and the land use(s) of that watershed. Drainage area
will aid in the analysis of inflow volumes to the lake due to surface run-
off. The land use classification of the area around the lake can be used
to predict flows and also nonpoint source pollutant loadings to the lake.
The physical parameters presented above may be used to understand and
analyze the various physical processes that occur in lakes. They can also
be used directly in simplistic relationships which predict productivity to
aid in aquatic use attainability analyses.
Physical Processes
There are many complex and interrelated physical processes which occur in
lakes. These processes are highly dependent on the lake's physical param-
eters, geographical location and characteristics of the contributing water-
shed. Individual physical processes are usually highly interdependent.
Five major processes—lake currents, heat budget, light penetration,
stratification and sedimentation—are discussed below. Each process can
affect the ecological system of a lake, especially the biota and the dis-
tribution of chemical species.
Lake Currents
Water movement 1n a lake affects productivity and the biota because it
influences the distribution of nutrientsv, microorganisms and plankton
(Wetzel, 1975). Lake currents are propagated by wind, inflow/out flow and
Coriolls force (a deflecting force which is a function of the earth's
rotation). The types of currents developed in lakes are dependent upon the
lake size and Its density structure.
For small, shallow lakes (especially those that are long and narrow),
inflow/outflow characteristics are most important and the predominant cur-
rent is a steady-state flow through the lake. For very large lakes, wind
is the primary generator of currents and, except for local effects, inflow
and outflow have a relatively minor affect on lake circulation. The
Coriolis force is another important determinant of circulation in larger
lakes such as the Great Lakes (Lick, 1976£)_.
Wind. Wind induced turbulence on the lake surface results in a variety of
current patterns that are characteristic of the lake's physical properties.
For shallow lakes, the wind induces vertical mixing throughout the water
column. Steady-state currents formed in deep lakes that have a constant
density are characterized by top and bottom boundary layers where vertical
II-6
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mixing is important, and by horizontal boundary layers near the shore where
horizontal mixing is important (Lick, 1976aj.
Under severe or prolonged wind conditions, the stress on the water surface
can cause circulation in the upper epilimnion region of a stratified lake
because of the inclination of the water surface. This then can cause a
counter flow in the lower hypolimnion region of the reservoir. This
condition is demonstrated by Fischer (1979) in Figure II-3. The flow
patterns are turbulent enough to disrupt the thermocline by tilting it
toward the leeward side of the lake. After the wind stops, internal water
movement causes the tilted upper and lower water regions, which are
separated by the thermocline, to oscillate back and forth until the pre-
wind stress steady-state condition returns (Wetzel, 1975). This type of
water movement caused by wind stress and subsequent oscillations is known
as a seiche.
Simply stated, an external seiche is a free oscillation of water, in the
form of long standing surface wave, reestablishing equilibrium after having
been displaced. The external seiche attains its maximum amplitude at the
surface while the internal seiche, which is associated with the density
gradient in stratified lakes, attains it maximum amplitude at or near the
thermocline (Figure II-4). In stratified waterbodies, the layers of
differing density oscillate relative to each other, and the amplitude of
the internal standing wave or internal seiche of the metalimnion is much
greater than that of the external or surface seiche. Because of the
extensive water movement associated with internal seiches, the resulting
currents lead to vertical and horizontal transport of heat and dissolved
substances (including nutrients) and significantly affect the distribution
and productivity of plankton (Wetzel, 1975).
Inflow and Outflow. Lake currents and the resultant mixing and horizontal
transport of the water mass may also be a function of inflow and outflow
patterns and volumes. Influent velocity generally decreases as the flow
enters the lake. Inflowing water of a given temperature and density tends
to seek a level of similar density in the lake. Three types of currents
may be generated by river influents, as shown in Figure II-5. Overflow
occurs when inflow water density is less than lake water density.
Underflow occurs when inflow density is greater than lake water density.
Interflow occurs when there is a density gradient in the lake, as during
periods of stratification, where inflow is greater in density than the
epilimnion but is less dense than the hypolimnion.
For a completely mixed lake where no density gradient exists, the outflow
draws on the totally mixed volume with little consequence to the net flow
within the lake. In stratified impoundments, where outflows could be from
different levels (e.g., reservoir release or withdrawal operations), the
discharge comes from only a limited zone (or layer) within the lake or
reservoir. The thickness of the withdrawal layer is a function of the
density gradient in the region'of the outlet.
Coriolis Effect. For very large lakes, like the Great Lakes, the Coriolis
effect can influence the currents within the lake. This effect is caused
by the inertial force created by the earth's rotation. It deflects a
moving body (water in this case) to the right (of the line of motion of the
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Id
Figure II-3.
Formation of baroclinic motions in a lake exposed to wind
stresses at the surface: (a) initiation of motion,
(b) position of maximum shear across the thermocline
(c) steady-state baroclinic circulation (from Fischer, 1979)
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>»WO IIC1M
Figure II-4.
Movement caused by (i) wind stress and (ii) a subsequent
internal seiche in a hypothetical two-layered lake,
neglecting friction. Direction and velocity of flow are
approximately indicated by arrows, o = nodal section.
(from Mortimer, 1952)
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(0) OVERFLOW Ptn < P
(0) uNCERPuOw P,n , p
FIGURE II-5. Types of inflow into lakes and reservoirs
(from Wunderlich, 1971)
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earth's rotation) in the Northern Hemisphere and to the left in the
Southern Hemisphere. The Coriolis effect causes the surface water to move
to the right of the prevailing direction of the wind. Under these con-
ditions in a stratified lake, less dense water tends to form on the right
side of the predominant current while denser water collects on the left
side of the current (Wetzel, 1975).
Heat Budget
The temperature and temperature distribution within lakes and reservoirs
affect not only the water quality within the lake but also the thermal
regime and quality of a river system downstream of the lake. The thermal
regime of a lake is a function of the heat balance around the body of
water. Heat transfer modes into and out of the lake include: heat trans-
fer through the air-water interface, conduction through the mud-water
interface, and inflow and outflow heat advection.
Heat transfer across the mud-water interface is generally insignificant
while the heat transfer through the air-water interface is primarily
responsible for typical annual temperature cycles in lakes.
Heat is transferred across the air-water interface by three different
processes: radiation exchange, evaporation, and conduction. The individ-
ual heat terms associated with these processes are shown in Figure II-6 and
are defined in Table II-l along with typical ranges of their magnitudes in
northern latitudes.
The expression that results from the summation of these various energy
fluxes is:
- (Hb + He ±
where
HM = net energy flux through the air-water interface,
Btu/ftz-day
H n = net short-wave solar radiation flux passing through the
interface after losses due to absorption and scattering
in the7atmosphere and by reflection at the interface,
Btu/ft -day
H = net long-wave atmospheric radiation flux passing through
the interface after reflection, Btu/ft -day
2
Hb = outgoing long-wave back radiation flux, Btu/ft -day
Hc s convective energy flux passing back and forth between
the interface and the atmosphere, Btu/ft -day
Hp = energy loss by evaporation, Btu/ft -day
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H8 Ha Hb
I ";
^ H
AIR-WATER
'INTERFACE
Hsn Han
Figure II-6. Heat Transfer Terms Associated with Interfacial Heat Transfer
(from Roesner, 1981)
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TABLE II-l
DEFINITION OF HEAT TRANSFER TERMS
ILLUSTRATED IN FIGURE II-6
Magnitude
Heat Term Units (BTU ft"2 day"1)
Hs •
Hsr =
Ha •
Har =
Hb '
He '
Hc '
where
H =
L =
T -
total incoming solar or 2 .
short-wave radiation HL T
-2 -1
reflected short-wave radiation HL T
total incoming atmospheric - ,
radiation HL'^T"1
-2 -1
reflected atmospheric radiation HL T
back radiation from the water ? .
surface HL" T"1
heat loss by evaporation HL" T"1
heat loss by conduction to - .
atmosphere HL" T"1
units of heat energy (e.g., BTU)
units of length
units of time
400-2800
40-200
2400-3200
70-120
2400-3600
150-3000
-320 to +400
SOURCE: Roesner, et al., 1981.
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These mechanisms by which heat is exchanged between the water surface and
the atmosphere are fairly well understood and are documented in the litera-
ture (Edinger and Geyer, 1965). The functional representation of these
terms has been defined by Water Resources Engineers, Inc. (1967).
The heat flux of the air-water interface is a function of location (lati-
tude, longitude and elevation), season of the year, time of day and
meteorological conditions in the vicinity of the lake. Meteorological
conditions which affect the heat exchange are cloud cover, dew-point
temperature, barometric pressure and wind.
Light Penetration
The heat budget discussed above is also descriptive of the light flux at
the air-water interface. The transmission of light through the water
column influences primary productivity, growth of aquatic plants,
distribution of organisms and behavior of fish.
The reduction of light through the water column of a lake is a function of
scattering and absorption where absorption is defined as light energy
transformed to heat. Light transmission is affected by the water surface
film, floatable and suspended particulates, turbidity, dense populations of
algae and bacteria, and color.
The intensity at a given depth is a function of light intensity at the
surface and the parameters mentioned above which attenuate the light.
Attenuation is usually represented by the use of a light extinction co-
efficient. ' ,
An important physical parameter based on the transmission of light is the
depth to which photosynthetic activity is possible. The minimum light
intensity required for photosynthesis has been established to be about 1.0
percent of the incident surface light (Cole, 1979). From the depth at
which this intensity occurs to the surface is called the euphotic zone.
Percent light levels can be measured by a subsurface photometer which can
be used to establish the depth of 1.0 percent illumination. A simple
measurement of light penetration depth is made with the Secchi disc which
is lowered into the water to record the depth at which it disappears to the
observer. The depth of the 1.0 percent surface light intensity may be
estimated as 2.7 to 3.0 times the Secchi disk transparency (Cole, 1979).
The percent of the surface incident light which reaches different depths is
highly variable for individual lakes. Cole (1979) presents examples of the
percent incident light by depth for various bodies of water, as shown in
Figure II-7.
Lake Stratification
Lakes in temperate and northern latitudes typically exhibit vertical
density stratification during certain times of the year. Stratification in
lakes is primarily due to temperature differences (i.e., thermal strati-
fication), although salinity and suspended solids concentration may also
affect density.
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a.
4)
a
1 -
2 -
3 -
5 -
Littl
0.1
I i
0.5 1.0
I I
5 10
50 100
Percent incident light
FIGURE II-7.
Vertical penetration of light in various bodies of
water showing percentage of incident light remaining
at different depths (from Cole, 1978)
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Lake stratification is best explained by a discussion of a generalized
annual temperature cycle. For a period in spring, lakes commonly circulate
from surface to bottom, resulting in a uniform temperature profile. This
vernal mixing has been called the spring overturn. As surface temperatures
warm further, the surface water layer becomes less dense than the colder
underlying water, and the lake begins to stratify. This stratified
condition, called direct stratification, exists throughout the summer, and
the increasing temperature differential between the upper and lower layers
increases the stability (resistance to mixing) of the lake.
The upper mixed layer of warm, low-density water is termed the epilimnion,
while the lower, stagnant layer of cold, high-density water is termed the
hypolimnion. The transition zone between the epilimnion and hypolimnion
has been called, among other names, the metalimnion. This narrow
transition zone is characterized by rapidly declining temperature with
depth, and it contains the thermocline which is the plane of maximum rate
of decrease in temperature. The region in which the temperature gradient
exceeds 1°C per meter may be used as a working definition of the thermo-
cline. A diagram of the three zones and the thermocline is presented in
Figure II-8, and Figure II-9 is a diagram of an annual temperature cycle in
which direct stratification occurs.
As surface water temperatures cool in the fall, the density difference
between isothermal strata decreases and lake stability is weakened.
Eventually, wind-generated currents are sufficiently strong to break down
stratification and the lake circulates from surface to bottom (fall
overturn). In warmer temperate regions, a lake may retain this completely
mixed condition throughout the winter, but in colder regions, particularly
following the formation of ice, inverse stratification often develops
resulting in winter stagnation. In this condition, the most dense, 4°C
water constitutes the hypolimnion which is overlied by less dense, colder
water between 0°C and 4°C. The difference in density between 0°C and 4"C
is very small, thus inverse stratification results in only a minor density
gradient just below the surface. Hence, the stability of inverse
stratification is low and, unless the lake is covered with ice, is easily
disrupted by wind mixing.
During stratification, the presence of the thermocline suppresses many of
the mass transport phenomena that are otherwise responsible for the ver-
tical transport of water quality constituents within a lake. The aquatic
community is highly dependent on the thermal structure of such stratified
lakes.
Retardation of mass transport between the hypolimnion and the epilimm'on
results in sharply differentiated water quality and biology between the
lake strata. For example, if the magnitude of the dissolved oxygen
transport rate across the thermocline is low relative to. the dissolved
oxygen demand exerted in the hypolimnion, vertical stratification of the
lake will occur with respect to the dissolved oxygen concentration.
Consequently, as ambient dissolved oxygen concentrations in the hypolimnion
decrease, the life functions of many organisms are impaired and the biology
and biologically mediated reactions fundamental to water quality are
altered. Major changes occur if the dissolved oxygen concentration goes to
zero and anaerobic conditions result. Large diurnal fluctuations of
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10
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30
40
50
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10
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25
30
TEMPERATURE, °C
FIGURE II-8.
Vertical temperature profile showing direct
stratification and the lake regions defined
by it (from Cole, 1979).
11-17
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LATE FALL -WINTER
FALL
OVERTURN
'•*'x\ — ~
SPRING
SUMMER
1_| STRATIFICATION
-t 1 I ' 1 1 •—
o * u ! it jo 11 10
ITI-CI
K> IS 20 IS
TCcI i
Figure II-9. Annual Cycle of Thermal Stratification and Overturn in an Impoundment (from Zison et al, 1977)
-------
dissolved oxygen concentrations in the epilimnion can also occur due to
daytime photosynthetic oxygen production superimposed over the continuous
oxygen demand from biotic respiration.
Vertical stratification of a lake with respect to nutrients can also occur.
In the euphotic zone, dissolved nutrients are converted to particulate
organic material through the photosynthetic process. Because the euphotic
zone of an ecologically advanced lake does not extend below the thermo-
cline, this assimilation of the dissolved nutrients lowers the ambient
nutrient concentrations in the epilimnion. Subsequent sedimentation of the
particulate algae and other organic matter then serves to transport the
organically bound nutrients to the hypolimnion where they are released by
decomposition. In addition, the vertical transport of the released
nutrients upward through the thermocline is suppressed by the same
mechanisms that inhibit the downward transport of dissolved oxygen. Thus,
several processes combine to reduce nutrient concentrations in the epilim-
nion while simultaneously enriching the hypolimnion.
In addition to the effect of the temperature structure on the movement of
water quality constituents, the temperature at any point has a more direct
impact on the biology and therefore the water quality structure of an
impoundment. All life processes are temperature dependent. In aquatic
environments, growth, respiration, reproduction, migration, mortality and
decay are strongly influenced by the ambient temperature. According to the
van't Hoff rule, within a certain tolerance range, biological reaction
rates approximately double with a 10°C increase in temperature.
Annual Circulation Pattern and Lake Classification
Lakes can be classified on the basis of their pattern of annual mixing as
described below.
Amixis Amictic lakes never circulate. They are permanently covered
with ice, and are mostly restricted to the Antarctic and very
high mountains.
Holomixis In holomictic lakes, wind-driven circulation mixes the entire
lake from surface to bottom. Several types of holomictic lakes
have been described.
Oligomictic lakes are characterized by circulation that is
unusual, irregular, and in short duration. These are generally
tropical lakes of small to moderate area or lakes of very great
depth. They may circulate only at irregular intervals during
periods of abnormally cold weather.
Monomictic lakes, undergo one regular period of circulation per
year. Cold monbmictic lakes are frozen in the winter (and
therefore stagnant and inversely stratified) and mix throughout
the summer. Cold monomictic lakes are ideally defined as lakes
whose water temperature never exceeds 48C. They are generally
found in the Arctic or at high altitudes. Warm monomictic
lakes circulate in the winter at or above 4"C and stratify
directly during the summer. Warm momomixis is common to warm
11-19
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regions of temperate zones, particularly coastal areas, and to
mountainous areas of subtropical latitudes. Warm monomictic
lakes are prevalent in coastal regions of North America and
northern Europe.
Dimictic lakes circulate freely twice a year in spring and
fall, and are directly stratified in summer and inversely
stratified in winter. Dimixis is the most common type of
annual mixing observed in cool temperate regions of the world.
Most lakes of central and eastern North America are dimictic.
Polymictic lakes circulate frequently or continuously. -Cold
pplymictic lakes circulate continually at temperatures near or
slightly above 4°C. Warm polymictic lakes circulate frequently
at temperatures well above 4"C.These lakes are found in
equatorial regions where air temperatures change very little
throughout the year.
Heromixis Meromictic lakes do not circulate throughout the entire water
column. The lower water stratum is perennially stagnant and is
called the monimolimm'on. The overlying stratum, the mixo-
limnion, circulates periodically, and the two strata are
separated by a severe salinity gradient called the chemocline.
Internal Flow and Lake Classification
Experience with prototype lakes (Roesner, 1969) has revealed that with
respect to internal flow structure there are basically three distinct
classes of lakes. These classes are:
o The strongly-stratified, deep lake which is characterized by
horizontal isotherms.
o The weakly stratified lake characterized by isotherms which are
tilted along the longitudinal axis of the reservoir.
o The nonstratified, completely mixed lake whose isotherms are
essentially vertical.
The single most Important parameter determining which of the above classes
a lake will fall is the densimetric Froude number, F, which can be written
for the lake as:
»
F = (LQ/DV) ( pQ/g0 J1/2 (2)
where
L = lake length, m
Q * volumetric discharge through the lake, nr/s
0 * mean lake depth, m
V » lake volume, nr
P. 3 reference density, taken as 1,000 kg/m 4
3= average density gradient in the lake, kg/m
g = gravitational constant, 9.81 m/sz
11-20
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This number is the ratio of the inertia! force of the horizontal flow to
the gravitational forces within the stratified impoundment; consequently,
it is a measure of the success with which the horizontal flow can alter the
internal density (thermal-) structure of the lake from that of its gravi-
tational static equilibrium state.
In deep lakes, the fact that the isotherms are horizontal indicates that
the inertia of the longitudinal flow is insufficient to disturb the overall
gravitational static equilibrium state of the lake except possibly for
local disturbances in the vicinity of the lake or reservoir outlets and at
points of tributary inflow. Thus, it is expected that F would to be small
for such lakes. In completely mixed lakes, on the other hand, the inertia
of the flow and its attendant turbulence is sufficient to completely upset
the gravitational structure and destratify the reservoir. For lakes of
this class, F will be large. Between these two extremes lies the weakly
stratified lake in which the longitudinal flow possesses enough inertia to
disrupt the reservoir isotherms from their gravitational static equilibrium
state configuration, but not enough to completely mix the lake.
For the purpose of classifying lakes by their Froude number, 0 and p in
equation (2) may be approximated as 10"3 kg/m4 and 1000 kg/m , respec-
tively. Substituting these values and g into equation (2) leads to an
expression for F as:
F = (320) (LQ/DY) (3)
where L and D .have units of meters, Q is in m3/s, and V has units of m.
It is observed from this equation that the principal lake parameters that
determine a lake's classification are its length, depth, and discharge to
volume ratio (Q/V).
In developing some familiarity with the magnitude of F for each of the
three lake classes, it is helpful to note that theoretical and experimental
work in stratified flow indicates that flow separation occurs in a strati-
fied fluid when the Froude number is less than I/*-, i.e., for F < 1/ir, part
of the fluid will be in motion longitudinally while the remainder is
essentially at rest. Furthermore, as F becomes smaller and smaller, the
flowing layer becomes more and more concentrated in the vertical direction.
Thus, in the deep lake it is expected that the longitudinal flow is highly
concentrated at values of F « 1/n- while in the completely mixed case F
must be at least greater than 1/rr since the entire lake is in motion and it
may be expected in general that F » 1/ir. Values of F for the weakly
stratified case would fall between these two limits and might be expected
to be on the order of I/T. As an illustration, five lakes are listed in
Table II-2 with their Froude numbers. It is known that Hungry Horse
Reservoir and Detroit Reservoir are of the deep reservoir class and can be
effectively described with a one-dimensional model along the vertical axis
of the lake. Lake Roosevelt, which has been observed to fall into the
weakly stratified class is seen to have a Froude number on the order of
I/*-, which is considerably larger than F for either Hungry Horse or Detroit
Reservoirs. Finally, Priest Rapids and Wells Dams, which are essentially
completely mixed along their vertical axes, show Froude numbers much larger
than I/*, as expected.
11-21
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TABLE II-2
IMPOUNDMENT FROUDE NUMBERS
RESERVOIR
Hungry Horse
Detroit
Lake Roosevelt
Priest Rapids*
Wells*
LENGTH
(meters)
4.7xl04
l.SxlO4
2.0xl05
2.9xl04
4.6xl04
AVERAGE
DEPTH
(meters)
70
56
70
18
26
DISCHARGE TO
VOLUME RATIO
(sec'1)
1.2xlO'8
3.5xlO'8
S.OxlO"7
4.6xlO"6
6.7xlO"6
F CLASS
0.0026 Deep
0.0030 Deep
0.46 Weakly
Stratified
2.4 Completely
Mi xed
3.8 Completely
Mixed
*R1ver run dams on the Columbia River below Grand Coulee Dam.
SOURCE: Roesner, 1969.
11-22
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Sedimentation in Lakes
One physical process that is particularly important to the aquatic
community is the deposition of sediment which is carried from the
contributing watershed into the body of the lake. Because of the low
velocities through a lake, reservoir or impoundment, sediments transported
by inflowing waters tend to settle to the bottom before they can be carried
through the lake outlets.
Sediment accumulation rates are strongly dependent both on the unique
physiographic characteristics of a specific watershed and upon various
characteristics of the lake. Although sediment accumulation rates can be
transposed from one lake to another, this should be done with a careful
consideration of watershed characteristics (Department of Agriculture,
1975, 1979). Apart from the use of predictive computer models, sediment
accumulation rates may be determined in one of two basic ways: (1) by
periodic sediment surveys on a lake; or (2) by estimates of watershed
erosion and bed load. Watershed erosion and bed load may be translated
into sediment accumulation rate through use of the trap efficiency, defined
as the proportion of the influent pollutant (in this case sediment) load
that is retained in the basin. The second method usually employs the
development of sediment discharge rate as a function of water discharge.
Such a sediment-rating curve is illustrated in Figure 11-10. From such
relationships, annual sediment transport to the lake is developed and
applied to the lake or reservoir trap efficiency functions to develop the
sediment accumulation rates. Trap efficiencies have been developed as a
function of the lake capacity-inflow ratio, as shown in Figure 11-11.
Other methods for predicting trap efficiency are described by Novotny and
Chesters (1981) and Whipple et al. (1983).
Accumulated sediment in lakes can, over many years, reduce the life of the
water body by reducing the water storage capacity. Sediment flow into
lakes also reduces light penetration, eliminates bottom habitat for many
plants and animals, and carries with it adsorbed chemicals and organic
matter which settle to the bottom and can be harmful to the ecology of the
lake. Where sediment accumulation is a major problem, proper watershed
management including erosion and sediment control must be put into effect.
CHEMICAL CHARACTERISTICS
Overview of Physico-Chemical Phenomena in Lakes
Water chemistry phenomena that are characteristic of freshwater have been
discussed in Section III, Technical Support Manual: Water Body Surveys and
Assessments for Conducting""^ Attainability Analyses iu.s. EPA, 1983D).
The material 1n Section III 1s applicable to lakes as well as rivers and
streams. The reader should refer to this Manual for a discussion of hard-
ness, alkalinity, pH and salinity, and for a discussion of a number of
indices of water quality. It would also be helpful to refer to Volume II
of this series, Technical Support Manual; Water Body Surveys and Assess-
ments for Conducting Use Attainability Analyses, Volume II; Estuarine
Systems, for a discussion of eutrophlcation and the importance of aquatic
vegetation. Even though the flora and fauna of estuaries have adapted to
11-23
-------
(0,000
3 i,oco
2, 100
.':! -•• ' •• "Ttfjr i' •'•'
i i
:0 iOO LOCO IC.GOO !00,CGO
Suspended sediment disc.'crge in tons per day
Figure 11-10. Sediment-rating curve for the Powder River at Arvada,
Wyoming (from Fleming, 1969)
(00
90
30
70
30
40
30
10
0
1 it n MU 11 it, ,t
.M i ! ' '
/
! :^^>-^i-----rr-- — : — ..-'• , .
'"•^\^r'i't
^ /£*l£i
•^1 i ! 1 ! : ; i
! ! ' : ' • i
Xl/TX-i : ' : « ' ' ;
! >f /*:
i i S /-A
1 ' ^ £ ' S^~*->—
! */**/\ S
/ # A
/ / / \
/./-/I !
jf Mtaiait cunt far
'' ;a/iaM rtitnair
i ,
Snitiaa* cur ft j far
normal ganatd 'tstrn
i
l '
! i-
normal \
t '' • •
\ a normal aenate' mtrvoiri
\ :a*ranant 'n ir'tct
*. Stim-ntio«)
Figure 11-11. Reservoir trap efficienty as a function of the capacity-
inflow ratio (from Brune, 1953)
11-24
-------
higher salinities than will be found in the lake, many of the interrela-
tionships of biology and nutrient cycling in the estuary have their
counterparts in the lake.
The discussion to follow will be limited to chemical phenomena that are of
particular importance to lakes. This will focus on nutrient cycling and
eutrophication, but will of necessity also be concerned with the effects of
variable pH, dissolved oxygen, and redox potential on lake processes.
Water chemistry in a lake and stages in the annual lake turnover cycle are
closely related. Turnover was discussed in greater detail earlier in this
chapter in the section on physical processes. For the current discussion
on lake water chemistry, we shall refer primarily to the stratified lake
that undergoes the classic lake turnover cycte. Since the patterns of lake
stratification and turnover vary widely, depending upon such factors as
depth, and prevailing climate as characterized by altitude and latitude,
the discussion to follow on water chemistry may not be applicable to all
lakes.
Once a thermocline has formed, the dissolved oxygen (DO) concentration of
the hypolimnion tends to decline. This occurs because the hypolimnion is
isolated from surface waters by the thermocline, and there is no mechanism
for the aeration of the hypolimnion. In addition, the decay of organic
matter in the hypolimnion as well as the oxygen requirements of fish and
other organisms in the hypolimnion serve to deplete DO.
With the depletion of DO, reducing conditions prevail and many compounds
that have accumulated in the sediment by precipitation are released to the
surrounding water. Compounds that are solubilized under, such conditions
include compounds of nitrogen, phosphorus, iron, manganese and calcium.
Phosphorus and nitrogen are of particular concern because of their role in
eutrophication processes in lakes.
Nutrients released from bottom sediments under stratified conditions are
not available to phytoplankton in the epilimnion. However, during overturn
periods, mixing of the hypolimnion and the epilimnion distributes nutrients
throughout the water column, making them available to primary producers
near the surface. This condition of high nutrient availability is short-
lived because the soluble reduced forms are rapidly oxidized to insoluble
forms which reprecipitate. Phosphorus and nitrogen are also deposited
through sorption to particles that settle to the bottom, and are trans-
ported from the epilimnion to the hypolimnion in dead plant material that
is added to sediments.
A special case occurs for ice covered lakes, esepcially when a layer of
snow effectively stops light penetration into the water. Under these
conditions winter algal photosynthesis is curtailed and dissolved oxygen
(DO) concentrations may decline as a result. A declining DO may affect
both the chemistry and the biology of the system. The curtailment of
winter photosynthesis may not pose a problem for a large body of water.
For a small lake, however, respiration and decomposition processes may
deplete available DO enough to result in fish kills.
11-25
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The chemical processes that occur during the course of an annual lake cycle
are rather complex. They are driven by pH, oxidation-reduction potential,
concentration of dissolved oxygen, and by such phenomena as the carbonate
buffering system which serves to regulate pH while providing a source of
inorganic carbon which may contribute to the many precipitation reactions
of the lake. The water chemistry of the lake may be better appreciated
through a detailed review of such references as Butler (1964), and Stumm
and Morgan (1981).
Of the many raw materials required by aquatic plants (phytoplankton and
macrophytes) for growth, carbon, nitrogen and phosphorus are of particular
importance. The relative and absolute abundance of nitrogen and phosphorus
are important to the extent of growth of aquatic plants that may be seen in
a lake. If these nutrients are available in adequate supply, massive algal
and macrophyte blooms may occur with severe consequences for the lake.
The concept of the existence of a limiting nutrient is the crux of Liebig's
"law of the minimum" which basically states that growth is limited by the
essential nutrient that is available in the lowest supply relative to
requirements. This applies to the growth of primary producers and to the
process of eutrophication in lakes where either phosphorus or nitrogen is
usually the limiting nutrient.
Algae require carbon, nitrogen and phosphorus in the approximate atomic
ratio of 100:15:1 (Uttormark, 1979), which corresponds to a 39:7:1 ratio on
a mass basis. The source of carbon is carbon dioxide which exists in
essentially unlimited supply in the water and in the atmosphere. Nitrogen
also 1s abundant in the environment and is not realistically subject to
control. Nitrate is introduced to the water body in rainfall, having been
produced electrochemically by lightening; in runoff to the water body; and
may be produced in the water body itself through the nitrification of
ammonia by sediment bacteria (Hergenrader, 1980). In contrast, many
sources of phosphorus to a lake are anthropogenic.
There are some lakes that are nitrogen limited, for which nitrogen controls
offer a means of controlling eutrophication. This is unusual, however,
and phosphorus limiting situations are much more prevalent than nitrogen
limiting conditions. As stated above, a N:P mass ratio of 7:1 is commonly
assumed to be required for algal growth; a N:P ratio less than 7:1
indicates that nitrogen 1s limiting, while a N:P ratio greater than 7:1
indicates a phosphorus limiting situation.
The growth of aquatic plants is limited when low phosphorus concentrations
prevail in a water body. Adequate control of phosphorus results in
nutrient limiting conditions that will hold the growth of aquatic plants in
check. Most Inputs of phosphorus to a lake are anthropogenic, thus control
of this nutrient offers the best means of regulating the trophic condition
of the lake. The focus of the discussion to follow will be an overview of
the chemistry of phosphorus and its interactions with pH, dissolved
oxygen, carbonates and iron in the water body.
A discussion of phosphorus chemistry may be approached through our under-
standing of the control of phosphorus in wastewater treatment plants by
precipitation reactions. As will be seen in Chapter IV, the principles of
11-26
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phosphorus control in wastewater processes may have application to lakes as
well. The chemistry of phosphorus is very complex and will not be dis-
cussed in great detail in this Manual. The reader who would like further
insight into the fine points of phosphorus chemistry should refer to texts
such as Butler (1964), and Stumm and Morgan (1981).
Phosphorus Removal by Precipitation
Phosphorus removal is discussed in detail in Process Design Manual for
Phosphorus Removal (U.S. EPA, 1976). Chapter 3 of that Manual. "Theory of
Phosphorus Removal by Chemical Precipitation," forms the basis of dis-
cussion for this section.
Ionic forms of aluminum, iron and calcium have proven most useful for the
removal of phosphorus. Calcium in the form of lime is commonly used to
precipitate phosphorus. Hydroxyl ions produced when lime is added to water
also play a role in phosphorus removal. .Because the chemistry of phos-
phorus reactions with metal ions is complex, it will be assumed for the
sake of simplicity that phosphorus reacts in the form of orthophosphate,
P
-------
in the hypolimnion, iron tends to be released from bottom sediments along
with phosphorus that has been tied up in the form of iron and manganese
precipitates.
Both ferrous (Fe^*) and ferric (Fe ) ions may be used to. precipitate
phosphorus. Ferric iron salts are effective for phosphorus removal at pH
4.5 to 5.0 although significant removal of phosphorus may be attained at
higher pH levels. Good phosphorus removal v/ith the ferrous ion is
accomplished at pH 7 to 8.
Lazoff (1983) examined phosphorus and iron sedimentation rates during and
following overturn to evaluate the removal of phosphorus through adsorption
and coprecipitation with iron compounds. At overturn, ferrous iron which
has been released along with phosphorus from the sediment, precipitates as
ferric hydroxides. Iron precipitation at overturn has been observed as the
formation of reddish brown floe particles. Phosphorus is removed from the
water column by these floe particles, either through adsorption or through
coprecipitation and settling. Thus, large amounts of phosphorus may be
removed from the water column and, therefore, become unavailable for
phytoplankton growth.
The removal of phosphorus by this mechanism may be aided by phytoplankton
and other sources of turbidity in the water. To the extent that these limit
light penetration into the water, photosynthesis and phosphorus uptake are
inhibited, thus permitting effective removal by ferric iron (Lazoff, 1983).
Dissolved Oxygen
Lake turnover, and mechanical aeration of bottom waters, leads to re-
oxygenation of the hypolimnion. If the hypolimnion was previously anoxic,
oxygenation will cause a reduction in PO* levels through the formation of
iron and manganese complexes and precipitates (Pastcrok et al., 1981). The
limited ability of iron, manganese and also calcium to tie up phosphorus in
a lake is regulated by 00 levels and by oxidation-reduction (redox)
potential. As the 00 of the hypolimnion falls, the redox potential
decreases and phosphorus is released during the reduction of metal pre-
cipitates that formed when the redox potential was higher. This may not be
a problem while the lake remains stratified, but once stratification ends
and the lake becomes mixed, the soluble phosphorus becomes available to
aquatic plants living near the surface. Lime does not reliably remove
phosphorus from the aquatic system because effective removal occurs at pH
levels greater than those found in natural waters.
Aluminum complexes are much less susceptible to redox changes and, there-
fore, are effective in permanently removing particulate and soluble phos-
phorus from the water column. Removal of phosphorus by aluminum occurs by
precipitation, by sorption of phosphates to the surface of aluminum
hydroxide floe and by the entrapment and sedimentation of phosphorus con-
taining particulates by aluminum hydroxide floe. Once deposited, the floe
of aluminum hydroxide appears to consolidate and phosphorus is apparently
sorbed from interstitial water as it flows through the floe (Cooke, 1981).
Oxygen depletion leads to low redox potentials in the sediment and a net
release of phosphorus into the water column. With aeration, the redox
11-28
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potential increases causing phosphorus to be precipitated and to be sorbed
by the sediment. Low pH values in the hypolimnion may be attributed to high
carbon dioxide associated with decay processes in the sediment. With
oxygenation, C02 levels decrease and pH increases (Fast, 1971).
Eutrophication and Nutrient Cycling
Eutrophication
There are two general ways in which the term "eutrophication" is used. In
the first, eutrophication is defined as the process of nutrient enrichment
in a water body. In the second, "eutrophication" is used to describe the
effects of nutrient enrichment, that is, the uncontrolled growth of plants,
particularly phytoplankton, in a lake or reservoir. The second use also
encompasses changes in the composition of animal communities in the water
body. Both of these uses of the term eutrophication are commonly found in
the literature, and the distinction, if important, must be discerned from
the context of use in a particular article.
Eutrophication is the natural progression, or aging process, undergone by
all lentic water bodies. However, eutrophication is often greatly
accelerated by anthropogenic nutrient enrichment, which has been termed
"cultural eutrophication."
In lakes nutrient enrichment often leads to the increased growth of algae
and/or rooted aquatic plants. For many reasons, however, excessive algal
growth will not necessarily occur under conditions of nutrient enrichment;
thus, the presence of high nutrient levels may not necessarily portend the
problems associated with the second use of the term eutrophication. For
example, the water body may be nitrogen limited or phosphorus limited,
toxics may be present that inhibit the growth of algae, or high turbidity
may inhibit algal photosynthesis despite an abundance of nutrients.
The three basic trophic states that may exist in a lake (or a river or
estuary) may be described in very general terms as follows:
o 01 igotrophie - the water body is low in plant nutrients, and may be
well oxygenated
o Eutrophic - the water body is rich in plant nutrients, and the
hypolimnion may be deficient in DO
o Mesotrophic - the water body is in a state between oligotrophic and
eutrophic.
What specific range of phosphorus or nitrogen concentration to ascribe to
each of these trophic levels is a matter of controversy since the degree of
response of a water body to enrichment may be controlled by factors other
than nutrient concentrations, in effect making the response site specific.
As will be seen in Chapter III, in a discussion of various measures of the
trophic state of a lake, eutrophication is a complex process and whether or
not a water body is eutrophic 1s not always clear, although the consequences
are.
11-29
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Nutrients are transported to lakes from external sources, but once 1n the
lake may be recycled internally. A consideration of attainable uses in a
lake must include an understanding of the sources of nitrogen and phos-
phorus, the significance of internal cycling, especially of phosphorus, and
the changes that might be anticipated if eutrophication could be controlled.
Nutrient Cycling in Lakes
There are many sources of nitrogen in the lake ecosystem. Significant
amounts of this nitrogen stem from natural sources and cannot be controlled.
Many anthropogenic sources, such as agricultural runoff, also are not
readily controlled. This is true in large part because the policy issues
surrounding nitrogen (and phosphorus) control through Best Management
Practices (BMPs) have not been resolved even though technical implementation
of BMPs could appreciably reduce nutrient loadings to a water body. Once in
the aquatic system nitrogen may undergo several bacterially mediated trans-
formations such as nitrification to nitrite and nitrate or denitrification
of nitrate to nitrogen. Proteins undergo ammonification to ammonia which in
turn is oxidized to nitrate. Also, some Cyanophyta (blue-green algae) are
capable of using atmospheric nitrogen. Unlike phosphorus, nitrogen is not
readily removed from a system-by complexation and precipitation reactions.
Whereas nitrogen inputs to a water body are predominantly non-point sources,
phosphorus inputs are predominantly point sources that are more readily
identified and controlled. There are some parts of the country, as in
Florida, where extensive phosphorus deposits are found which could be the
source of significant natural inputs to a lake and its feeder streams. Such
lakes may be nitrogen limited. With the exception of runoff, the anthro-
pogenic sources (particularly the point sources) of phosphorus can be
controlled to a large extent. Control of the external inputs of phosphorus
to a lake may not necessarily end problems of eutrophication, however,
annual fluctuations in DO, pH and other parameters may result in the
recycling of significant amounts of phosphorus within the system.
Uttormark (1979) has noted that most lakes are nutrient traps, on an annual
basis, and that the-trophic status of a lake can be dependent on the degree
of internal nutrient cycling that occurs. There is typically a seasonal
release from and deposition of nutrients to the sediment, and the effect of
this internal nutrient cycling is dependent upon physical characteristics
such as morphology, mixing processes and stratification.
As discussed earlier, phosphorus that has been released from sediments to
anoxic bottom waters under stratified conditions may become temporarily
available to primary producers during overturn periods. This often causes
phytoplankton blooms in spring and fall. During winter and summer,
stratification limits vertical cycling of nutrients and nutrient
availability may limit phytopiankton growth.
Macrophytes derive phosphorus directly from lake sediment or from the water
column. The release of some of this phosphorus to the surrounding water has
been reported for some macrophytes (Landers, 1982). In addition, signifi-
cant amounts of phosphorus and nitrogen are released to the surrounding
water by macrophytes as they die and decompose. Landers has estimated that
about one-fourth of the phosphorus and one-half of the nitrogen within a
11-30
-------
decaying plant will remain as a refractory portion, while the rest is
released to the surrounding water.
In response to soluble phosphorus released by decomposing macrophytes, the
algal biomass (as measured by chlorophyll-^ concentration) may show a
significant increase. When these algae later die, phosphorus will be
returned to the system in soluble form, as precipitates that form with iron,
calcium and manganese, or will be tied up in dead cells that settle to the
bottom to become part of the sediment.
Significance of Chemical Phenomena to Use Attainability
The most critical water quality indicators for aquatic use attainment in a
lake are dissolved oxygen (DO), nutrients, chlorpphyll-a and toxicants.
Dissolved oxygen is an important water quality indicator For all fisheries
uses and, as we have seen above, is an important factor in the internal
cycling of nutrients in a lake. In evaluating use attainability, the
relative importance of three forms of oxygen demand should be considered:
respiratory demand of phytoplankton and macrophytes during non-
photosynthetic periods, water column demand, and benthic demand. If use
Impairment is occurring, assessments of the significance of each oxygen sink
can be useful in evaluating the feasibility of achieving sufficient pol-
lution control, or in implementing the best internal nutrient management
practices to attain a designated use.
Chlorophyll-^ is a good indicator of algal concentrations and of nutrient
overenrichment. Excessive phytoplankton concentrations, as indicated by
high chlorophyll-^ levels, can cause adverse DO impacts such as: (a) wide
diurnal variation 1n surface DO due to daytime photosynthetic oxygen pro-
duction and nighttime oxygen depletion by respiration and (b) depletion of
bottom DO through the decomposition of dead algae and other organic matter.
Excessive algal growth may also result in shading which reduces light
penetration needed by submerged plants.
The nutrients of concern in a lake are nitrogen and phosphorus. Their
sources typically are discharges from industry and from sewage treatment
plants, and runoff from urban and agricultural areas. Increased nutrient
levels may lead to phytoplankton blooms and a subsequent reduction in DO
levels, as discussed above.
Sewage treatment plants are typically the major point source of nutrients.
Agricultural land uses and urban land uses are significant non-point sources
of nutrients. Wastewater treatment facilities often are the major source of
phosphorus loadings while non-point sources tend to be the major con-
tributors of nitrogen. It is important to base control strategies on an
understanding of the sources of each type of nutrient, both in the lake and
in Its feeder streams.
Clearly the levels of both nitrogen and phosphorus can be important deter-
minants of the uses that can be attained in a lake. Because point sources
of nutrients are typically more amenable to control than non-point sources,
and because phosphorus removal for municipal wastewater discharges is
typically less expensive than nitrogen removal, the control of phosphorus
11-31
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discharges is often the method of choice for the prevention or reversal of
use impairment in the lake.
Discussion of the impact of toxicants such as pesticides, herbicides and
heavy metals is beyond the scope of this volume. Nevertheless, the presence
of toxics in sediments or in the water column may prevent the attainment of
uses (particularly those related to fish propagation and maintenance in
water bodies) which would otherwise be supported by water quality criteria
for 00 and other parameters.
TECHNIQUES FOR USE ATTAINABILITY EVALUATIONS
Introduction
In the use attainability analysis, it must initially be determined if the
present aquatic life use of a lake corresponds to the designated use. The
aquatic use of a lake is evaluated in terms of biological measures and
indices. If the designated use is not being achieved, then physical, chem-
ical and biological investigations are carried out to determine the causes
of impairment. Physical and chemical factors are examined to explain the
lack of attainment, and they are used as a guide in determining the highest
use level the system can achieve.
Physical parameters and processes must be characterized so that the study
lake can be compared with a reference lake. Physical parameters to be
considered are average depth, surface area, volume and retention time. The
physical processes of concern include degree of stratification and
importance of circulation patterns. Once a reference lake has been
selected, comparisons can be made with the lake of interest in terms of
water quality differences and differences in biological communities.
Empirical (desktop) and simulation (computer-based mathematical) models can
be used to improve our understanding of how physical and chemical char-
acteristics affect biological communities. Desktop analyses may be used to
obtain an overall picture of lake water quality. These methods are usually
based on average annual conditions. For example, they are used to predict
trophic- state based on annual loading rates of nutrients. They are simple,
inexpensive procedures that provide a useful perspective on lake water
quality and in many cases will provide sufficient information for the use
study. For a more detailed analysis of lake conditions, computer models can
be employed to analyze various aspects of a lake. These models can simulate
the distribution of water quality constituents spatially (at various
locations within the lake) and temporally (at various times of the year).
Desktop calculations and larger simulation models may both be used to
enhance our understanding of existing lake conditions. More importantly,
they can be used to evaluate the lake's response to different conditions
without actually imposing those conditions on the lake. This is of great
benefit in determining the cause of impairment where, for example, the model
can predict the lake response to the removal of point and nonpoint loads to
the lake system. Models can also be used to assess potential uses by simu-
lating the lake's response to various design conditions or restoration
activities. A good discussion of model selection and use is provided by the
U.S. EPA (1983c). '
11-32
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Empirical Models
In contrast to the complex computer models available for the study of lake
processes, there are a number of simple empirical, input/output models that
have proven to be widely applicable to lake studies. Most of these models
consider phosphorus loadings or chlorophyll-^ concentrations in order to
estimate the trophic status of a lake.
Vollenweider Model
Vollenweider (1975) proposed an empirical fit to a simplified phosphorus
mass balance model, using the factor:
9 = 10/2
where
y = specific sedimentation rate, years"1
z = mean lake depth, m
Sedimentation is used by Yollenweider to describe all net internal losses of
phosphorus (Uttormark, 1978) and is extremely difficult to determine
experimentally. Yollenweider derived his value for
-------
10-
O
z
a
.01-
EUTROPHIC
DANGEROUS
PERMISSIBLE
OLIGOTROPHIC
i
10
100
1000
w
Figure II-12a. The Vollenweider Model (from Zison, et al., 1977)
11-34
-------
10
J-
*e
a.
S i
a
z
o
o
-1
-------
An example application of this type of approach is given by Zison, et al.
(1977), where the characteristics of a reservoir are given as:
Bigger Reservoir
Available Data (all values are means):
Length 20 mi = 32.2 km
Width 10 mi = 16.1 km
Depth (z) 200 ft - 61 m
Inflow (Q) 500 cfs
Total phosphorus concentration in inflow 0.8 ppm
Total nitrogen concentration in inflow 10.6 ppm
First determine whether phosphorus is likely to be growth limiting. Since
data are available only for influent water, and since no additional data
are available on impoundment water quality, N:P for influent water will be
used.
M:P = 10.6/0.8 = 13.25
Thus, recalling that a N:P mass ratio of 7:1 is required for algal growth,
Bigger Reservoir is probably phosphorus limited.
Compute the approximate surface area, volume and the hydraulic residence
time.
Volume (V) = (20 mi) (10 mi) (200 ft) (5280 ft/mi)2 =
1.12 x 1012ft3 = 3.16 x 1010m3
Hydraulic residence time (r ) = y/g =
W
1.12 x 1012ft3/500 ft^ec"1 = 2.24 x 109sec = 71 yr
Surface area (A) = (20 mi) (10 mi) (5280 ft/mi)2 =
5-.57 x 109ft2 - 5.18 x 108m2
Next, compute hydraulic loading, q_
q - 61 m/71 yr = 0.86 m yr
Compute annual inflow, Q
Qy = (Q) (3.25 x 107sec yr"1)
Q = 1.58 x 1010ft3 yr"1
Phosphorus concentration in the inflow is 0.8 ppm, or 0.8 mg/1. Loading
(L_) in grams per square meter per year is computed from the phosphorus
concentration (Cp), the annual inflow (Q ), and the surface area (A):
11-36
-------
L0 = V1'JO A ™ " '''""I" '"* r/l<(28.32 im3Hl x 10'3 mg/g)
? (5.18 x 10a nr)
Lp = O./O g/nf-y.'
Referring to the plot in Figure II-12_a, we would expect that Bigger Reser-
voir, with Ln = 0.7 and qs = 0.86, is .eutrophic, possibly with severe
summer algal blooms.
The Vollenweider type of approach has many useful and varied applications.
For example, a phosphorus loading model was used to evaluate three pro-
spective reservoir sites for eutrophication potential (Camp Dresser &
McKee, 1983). Since this evaluation was part of a study to select a future
dam site, and an impoundment did not exist, there was very little infor-
mation, available with which to work. While such an evaluation was not a
use attainability study per se, the application is instructive because in
many cases there may be virtually no data available for use in evaluating
an existing lalca or impoundment for attainable uses. For these cases where
few historical data are available, use of a computer model would require
simulation predictions without the benefit of a calibrated model, unless
considerable resources are available to conduct a sampling program to
characterize the water body from season to season in order to generate the
data required by such a model. There are few options in this case other
than use of an empirical model which, nevertheless, may provide very
instructive results.
In the reservoir site study, phosphorus loading was estimated from water
quality data for the streams that would feed each of the prospective
reservoirs, and from an evaluation of land use practices in the watersheds.
Streamflow data and an analysis of rainfall-runoff relationships provided
an estimate of flow (Q) to each of the three reservoirs, and topographic
maps were used to determine reservoir volume, average depth (z), and
surface area (A).
In the analyses, the quantity l/r may be calculated as:
W
Z>w = ZP= (Y/AMQ/V) = Q/A
where p, the flushing rate, is equal to the reciprocal of r, the hydraulic
residence time.
The quantity Q/A is the hydraulic loading rate—the amount of water added
annually per unit area of lake surface. This may be interpreted to imply
that lakes with the same hydraulic and phosphorus loadings should have the
same 1n-lake phosphorus concentration regardless of differences in flushing
rates (Uttormark and Hutchins, 1978).
The flushing rate is a very important characteristic of a lake, and is an
important determinant of trophic state. If the flushing rate is high, as
11-37
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might be the case in a run-of-river impoundment, algal growth problems may
be much less for a given phosphorus loading than for the same phosphorus
loading to a lake with a low flushing rate. Although hydraulic loading
serves as a surrogate for flushing rate in the Vollenweider model, the
model still represents an important advancement beyond static loading
estimations, such as were presented in Yollenweider in 1968 (Table II-3)
where estimates for trophic state are based solely on mass loading.
Vollenweider-OECD Model
The Organization for Economic Cooperation and Development (OECD)
Eutrophication Study was conducted in the early 1970's to quantify the
relationship between the nutrient (phosphorus) load to a water body (lake,
reservoir, or estuary) and the eutrophication-related water quality
response of the water body to that load. Rast and Lee (1978) applied the
Vollenweider (1975) model to the OECO water bodies in the United States.
The results are plotted in Figure II-12J). It is apparent that the
eutrophic water bodies are clustered in one area of the plot and the
oligotrophic water bodies in another. Between those two zones, the authors
delineated rough boundaries of permissible and excessive phosphorus loading
with respect to eutrophication-related water quality. This model can be
used in the same way as the Vollenweider model discussed previously.
Dillon and Rigler Model
In 1974, Dillon and Rigler (as reported by Uttormark and Hutchins) pub-
lished an empirical model, similar to that of Vollenweider, in which a
phosphorus retention coefficient (R) was proposed to account for phosphorus
retention in the lake.
R = '"in - "cut-in <«1
Incorporation of R into the phosphorus mass balance equation leads to
Equation 7 for the Dillon-Rigler model which is analogous to Equation 5 for
the Yollenweider model.
[P] » Ul-R)/(zP ) (7)
Dillon and Rigler used values of 10 and 20 mg-P/m to define acceptable and
excessive loading values to derive Figure 11-13. Figure 11-13 may be used
to estimate trophic state by plotting the quantity:
L(1-R)/P vs. I
where
L = annual phosphorus loading,
R » retention coefficient, (?..„ - Pni,+)/P
-------
TABLE II-3
SPECIFIC NUTRIENT LOADING LEVELS FOR LAKES
!(EXPRESSED AS TOTAL NITROGEN AND
* TOTAL PHOSPHORUS IN g/mz-yr)*
Mean Depth
Up To:
5 m
10 m
50 mf
100 m
150 m
200 m
Permissible
Loading
Up To:
N
1.0
1.5
4.0
6.0
7.5
9.0
P
0.07
0.10
0.25
0.40
0.50
0.60
Dangerous
Loading in
Excess of:
N
2.0
3.0
8.0
12.0
15.0
18.0
P
0.13
0.20
0.50
0.80
1.00
1.20
*from Yollenweider (1968)
SOURCE: Uttormark and Hutchins, 1978.
11-39
-------
I
-£•
o
N 10
o
_c
cu
•*»
*^
or
8
o_
10
10
EUTROPHIC
OLIGOTROPHIC
M
10
I I
10
10'
Mean depth, z . In meters
Figure 11-13. The Dlllon-Klgler Model (from Dillon and Ritjler, 1974).
-------
The lines of Figure 11-13 represent equal predictive phosphorus concentra-
tions, indicating that the prediction of the trophic state of a lake is
based on a measure of the predictive phosphorus concentration in the lake
rather than on the phosphorus loading (Tapp, 1978).
Larsen and Mercier Model
Larsen and Mercier (as reported in Tapp, 1978) used the phosphorus mass
balance model to describe the relationship between the steady state lake
and mean input phosphorus concentrations. Again using values of 10 and 20
mg/nr (ug/1), Larsen and Mercier developed the curves of Figure 11-14 to
distinguish oligotrophic, mesotrophic and eutrophic conditions. To use
Figure 11-14, one needs to estimate the mean influent lake phosphorus
concentration, P, in g/m , and Rexn, the fraction of phosphorus retained in
the lake. The Larsen and Mercver formula plots mean tributary total
phosphorus concentration against a phosphorus retention coefficient,
thereby addressing the criticism of other models that no distinction is
made between phosphorus increases due to influent flows or concentrations
or both (Hern, et al., 1981). In effect, the Larsen and Mercier model
predicts the mean tributary phosphorus concentration which would cause
eutrophic or mesotrophic conditions.
In a comparative test of these three phosphorus loading models, using data
collected under the National Eutrophication Survey on 23 water bodies (most
in the northeastern and north central United States), it was found that the
Dillon-Rigler and Larsen-Mercier models fit the data much better than the
Yollenweider model (Tapp, 1978). This is probably because the Vollenweider
model considers only total phosphorus loading without regard to in-lake
processes that reduce the effective phosphorus concentration. In a similar
comparison on data from southeastern water bodies, however, all three of
the models generally fit the data.
Of the empirical models, the Vollenweider is the most conservative because
it does not account for phosphorus in the outflow from a lake. This model
should be used in a first level of analysis, in the absence of sufficient
data to establish a phosphorus retention coefficient. If the retention
coefficient can be derived, the Dillon-Rigler or Larsen-Mercier models
would be preferable (Tapp, 1978).
Reckhow (1979) cautions that the application of empirical phosphorus lake
models may not be appropriate for certain conditions or types of lakes.
These include conditions of heavy aquatic weed growth, violation of model
assumptions (for example, no outlet from a lake), or because the lake type
(such as extremely shallow lakes) was not included in the data sets used to
develop each of the models.
Sedimentation rates are apt to differ in a closed lake from sedimentation
in a lake with an outlet. Based on a consideration of the phosphorus mass
balance equation with the outflow term removed, and upon settling rates
discussed by Dillon and Kirchner (1975) and Chapra (1977), Reckhow (1979)
proposed the following expression for predicted phosphorus concentration:
11-41
-------
1000
100
10.
10
I I
EUTROPHIC
/
/
/ ,
/
X /
/
/
X
OUGOTROPHIC
J I ! I
0.1 0.2 0.3 0.4 0.5 0.6 0.7 0.8 0.9 1.0
R
Figure II-14-. The Larsen-Mercier Model (from Tapp, 1978).
11-42
-------
L/(16 + ZP ) < Ptrue < L/13.2 . (8)
Shallow lakes present a problem because the potential for mixing of the
sediments results in phosphorus concentrations that may be more variable
than in deeper lakes. On the other hand, these same conditions may prevent
the development of anaerobic conditions and serve to reduce concentration
variability. Modeling of lakes with heavy weed growth is problematic
because thick growths may restrict mixing, while interacting directly with
the sediment.
Modified Larsen and Mercier Model
Hern, et al. (1981) note the assumption inherent to each of the phosphorus
models discussed above that the relationship of phytoplankton biomass to
phosphorus is the same for all lakes, yet point out that the utilization
and incorporation of phosphorus into phytoplankton biomass varies sig-
nificantly from lake to lake, depending on availability of light; supply of
other nutrients, bioavailability of the various species of phosphorus, and
a number of other factors. They go on to evaluate the factors affecting
the relationship of phytoplankton biomass to phosphorus levels and show how
the phosphorus models may be modified to base trophic state assessments on
chlorophyll-^ rather than phosphorus.
In their analysis of sampling data from a number of lakes, Hern et al.
determined that the response ratio of chlorophyll-^ (CHLA) to high summer
phosphorus concentrations decreases as total phosphorus increases, in
contrast to the findings of other authors (Vollenweider, Dillon, etc.)
whose work is based on data collected in lakes that were free of major
interferences. Hern, et al., indicate a belief that the reason most lakes
do not reach maximum production of chlorophyll-a is because of interference
factors. Factors which may prevent phytopTankton chlorophyll-^ from
achieving maximum theoretical concentrations based on ambient total
phosphorus (TP) levels in a lake include:
1. Availability of light (for example, limitations due to turbidity
or plankton self shading);
2. Limitation of growth by nutrients other than total phosphorus,
e.g., nitrogen, carbon, silica, etc.;
3. Biological availability of the TP components;
4. Domination of the aquatic flora by vascular plants rather than
phytoplankton;
5. Grazing by zooplankton;
6. Temperature;
7. Short hydraulic retention time; and
8. Presence of toxic substances.
11-43
-------
The response ratio (RA) is defined as the amount of chlorophyll-a formed
per unit of total phosphorus. A strong relationship between~CHLA (a
measure of phytoplankton biomass) and TP in lakes has been established by a
number of authors, as discussed by Hern et al. (1981). A log-log trans-
formation of the response ratio and total phosphorus concentration yields a
straight line (Figure 11-15) which provides a basis of comparison between
the theoretical RA and the actual RA at a given phosphorus level. This
relationship was used to modify the Larsen-Mercier model to accompish the
following objectives:
1. Change the trophic classification based on an ambient TP level to
one based on the biological manifestation of nutrients as measured
by chlorophyll-£;
2. Determine the "critical" levels of TP which will result in an un-
acceptable level of CHLA concentration so that the level of TP can
be manipulated to achieve the desired use of a given water body;
and
3. Account for the unique characteristics of a lake or reservoir
which affect the RA.
The Larsen and Mercier (1976) model predicts the mean tributary TP concen-
tration which would cause eutrophic or mesotrophic conditions as follows:
TFr = ETP or (9)
E
TFM = MTP (10)
M
where
= the minimum mean tributary TP concentration in ug/1 which will
cause a lake to be eutrophic at equilibrium,
TP~M = the minimum mean tributary TP concentration in ug/1 which will
cause a lake to be mesotrophic at equilibrium,
ETP = a constant equal to 20, which is the theoretical minimum
ambient ug/1 of TP in a lake resulting in eutrophic conditions
and is the level which if not equaled or exceeded will result
in meso- or oligotrophic conditions,
11-44
-------
-1
-2
DC
a
o
-3
-4
-5
-6
I J_
-4 -2
Log TP In jjg/l
Figure 11-15. The relationship between summer log RA and log TP based
on Jones and Pachmann's (1976) regression equation (from
Hern, et al., 1981).
-------
MTP = a constant equal to 10, which is the theoretical minimum
ambient ug/1 of TP in a lake resulting in mesotrophic condi-
tions and is the level which if not equaled or exceeded will
result in oligotrophic conditions, and
R = fraction of phosphorus retained in the lake.
The Larsen and Mercier equations (i.e., Equations 9 and 10) can be
corrected to account for the RA of a specific lake as follows:
.f = ETP(ERA/AERA) (11)
At
TP~.M = MTP(MRA/AMRA) (12 )
AM - - '
where
= the minimum mean tributary TP concentrations in ug/1 which will
cause a lake to be eutrophic at equilibrium corrected to
account for the lake's RA,
TP... a the minimum mean tributary TP concentrations in ug/1 which will
cause a lake to be mesotrophic at equilibrium corrected to
account for the lake's RA,
ERA * a constant equal to 0.32 which is the RA predicted from 20 ug/1
of ambient TP utilizing Jones and Bachmann's (1976) regression
equation,
MRA s a constant equal to 0.23 which is the RA predicted from 10 ug/1
of ambient TP utilizing Jones and Bachmann's (1976) regression
equation,
AERA a the mean summer RA for the lake corrected to what it would be
at the 20 ug/1 level of TP, i.e., the ambient eutrophic level,
and
AMRA a the mean summer RA for the lake corrected to what it would be
at the 10 ug/1 level of TP, I.e., the ambient mesotrophic
level.
The ERA constant of 0.32 was determined from utilizing the ETP constant of
20 ug/1 of ambient TP in the Jones and Bachmann (1976) regression equation:
log ug/1 CHLA - -1.09 + 1.46 log ug/1 TP (13)
11-46
-------
Substituting 20 ug/1 for TP, log CHLA is equal to 0.81 and CHLA is equal to
6.4. Therefore, the ERA is equal to 6.4/20 or 0.32. Similarly, the MRA
constant of 0.23 was determined utilizing the MTP constant of 10 ug/1 of
ambient TP.
The AERA is determined from the following equation:
*A (14)
where
ORA = the observed summer ambient RA in the lake,
OTP = the observed summer ambient TP in the lake,
A = -4.77 which is the log of the RA determined from Equation 13
utilizing a TP concentration at approximately 0 (since log 0 is
undefined, an extremely low TP concentration, i.e., O.OOOOOC01
ug/1, was used to approximate 0 on the log scale), and
B = -8 which is the log of the TP (i.e., 0..00000001 ug/1, which is
used to approximate 0 in Equation 13).
Substituting into Equation 14:
The AMRA is determined from the following equation:
A (16)
Substituting into Equation 16:
109 AHRA . [°lf^p ?f ] (9) - 4.77 (17)
The constants used in Equations 14 and 16 are used to establish the slope
of a line (Figure 11-15) which begins at -4.77 (log RA) and -3 (log TP).
Using the ORA and the OTP, the RA is adjusted using the relationship shown
in" Figure 11-15, which was determined from the Jones and Bachmann (1976)
regression equation (Equation 13) to one which would cause eutrophic (AERA)
or mesotrophic conditions in the lake (AMRA).
A comparison of trophic state predictions using the Larsen and Mercier
equations (Equations 9 and 10) with the modified equations to account for a
lake's RA (Equations 11 and 12) was made using lake field data (Hern, et
al., 1981). Those data showed that the lake had:
11-47
-------
OTP = 36.3 ug/1,
observed mean summer CHLA (OCHLA) = 6.3 ug/1,
1-R = 0.71,
ORA = 0.17, and
observed mean tributary TP (OTTP) = 57.3 ug/1.
Substituting into Equation 9 (the Larsen-Mercier equation that yields the
minimum mean tributary TP that will cause a lake to be eutrophic), we find:
TPC = 20 = 28.2 ug/1 (9)
L 077T
Since 28.2 ug/1 of TP represents the theoretical minimum mean tributary
concentration which will cause the lake to be eutrophic under steady state
conditions and the OTTP is 57.3 ug/1, the use of Equation 9 would classify
the lake as eutrophic. Substituting into Equation 11 which gives the mean
tributary TP that will cause a lake to be eutrophic, when this TP is
corrected for the lake's response ratio, RA:
TPAC = 20(0.32/0.13) = 69.3 ug/1 (11)
ME DT7I
Since 69.3 ug/1 is greater than 57.3 ug/1, we find if we use the modified
equation which accounts for the lake's RA, the lake could be classified as
mesotrophic and could possibly be oligotrophic. To determine whether it is
mesotrophic or oligotrophic, we substitute into Equation 12 to determine
the mean tributary TP, corrected for the lake's RA, that will support
mesotrophic conditions.
= 10(0.23/0.10) * 32.4 ug/1 (12)
Since 32.4 ug/1 is less than 57.3 ug/1, we would classify the lake as
mesotrophic.
Computer Models
For many lakes, desktop evaluations and the analysis of field data may not
be sufficient for an analysis of attainable uses. When a more sophisti-
cated analysis is indicated, computer-based mathematical models can be used
to simulate physical and water quality parameters, as well as various life
forms and their interrelationships. The model predictions can be used to
determine whether physical and water quality conditions are adequate for
11-48
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use attainment. For example, using the information on biological require-
ments presented later in this manual in conjunction with predicted water
quality conditions, judgments can be made regarding what type of aquatic
life community a lake is likely to be capable of supporting. Computer
models have the great advantage that they can predict the lake's ecological
system rapidly under various design conditions and in addition, many
computer models can simulate dynamic processes in the water body. In
contrast, the phosphorus loading empirical models are suited only to steady
state assumptions about the lake.
Which computer model to select will depend on the level of sophistication
required in the analysis to be conducted. The selection will also depend
highly on the size of the lake and its particular physical characteristics.
For example, a long, narrow lake which is fully mixed horizontally and
vertically can be modeled by a one-dimensional model. Two-dimensional
models may be required where lake currents in a very large, shallow lake
are the dominant factor affecting lake processes. In deep lakes where the
vertical variations in lake conditions are most important, one-dimensional
models in the vertical direction are appropriate.
In many cases lake water quality and ecological models have been developed
to high degrees of sophistication, but these models do not provide the same
degree of sophistication for the mechanisms that describe transport
phenomena in the lake. On the other hand, models developed to simulate the
hydrodynamics of a lake did not include the simulation of an extensive
array of chemical and biological conditions. One of the major weaknesses
in current water quality models as perceived by Shanahan and Harleman
(1982) is the linkage of hydrodynamic and biochemical models.
Hydrodynamic Modeling
Shanahan and Harleman (1982) have described various types of models for
lake circulation studies. They included two major groups: simplified
models and true circulation models.
The simplified models included zero-dimensional models in which a lake is
represented by a fully-mixed tank or continuous-flow stirred tank reactor.
For a larger lake, representation with the zero-dimensional model is accom-
plished by treating different areas of the lake as separate fully mixed
tanks. Simplified models also include longitudinal and vertical one-
dimensional models. These models consider a series of vertical layers or
horizontal segments.
True circulation models are those which employ two- and three-dimensional
analysis. Two-dimensional models have been developed with a single or with
multiple layers where it is assumed that the lake is vertically homogeneous
within a layer. While lake circulation is modeled in each layer, the
interactions between layers must be considered separately. The fully
three-dimensional model, which also handles vertical transport between
layers, is the most complex, and most expensive to set up and run.
Although there are some examples of this type of model in use, Shanahan and
Harleman believe that these models have not reached a point of practical
application.
11-49
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Numerical lake circulation models have been investigated in detail by
Wilbert Lick of the Case Western Reserve University. In a report for the
U.S. Environmental Protection Agency, Lick (1976^) describes his work on
three-dimensional models. The three-dimensional models developed by Lick
include: (1) a steady-state, constant-density model; (2) a time-dependent,
constant-density model; and (3) a time-dependent, variable-density model.
Vertically averaged models are also presented which average the three-
dimensional equations over the depth, thus reducing the model to a two-
dimensional model.
Lake Water Quality Modeling
Many one-, two- or three-dimensional lake water quality models have been
developed for various applications. As part of an EPA technical guidance
manual for performing wasteload allocations (U.S. EPA, 1983^), available
water quality models were reviewed. Information concerning model
capability, model developers, and technical support were presented.
Descriptions of lake models from Book IV - Lakes and Impoundments, Chapter
2 - Eutrophication (U.S. EPA, 1983d are provided in Tables II-4 through
II-8 to present an overview of some of the models that have been developed
for lake studies.
Lake water quality models such as those described in Tables II-4 through
11-8 generally are stand-alone models, however, some lake quality models
have been linked to sophisticated hydrodynamic models. For example, in one
special study for Lake Ontario, Chen and Smith (1979) developed a three-
dimensional ecological-hydrodynamic model. The hydrodynamic model
calculated currents and the temperature regime throughout the lake using a
horizontal grid with eight layers of thickness. The water quality model
included a coarser horizontal grid with seven layers. The hydrodynamic
information was transferred through an interface program to the water
quali ty model.
Much of the focus in water quality models developed for deep lakes and
reservoirs has centered around the prediction of the thermal energy
distribution, and has led to the development of one-dimensional ecological
models such as LAKECO and WQRRS as described in Tables 11-7 and 11 -8,
respectively. This type of model is described in more detail in the
following section.
One-Dimensional Lake Modeling
Development .of LAKECO, WQRRS and other variations of these ecological
models such as EPAECO (Gaume and Duke, 1975) began in the late sixties with
studies on the prediction of thermal energy distribution (Water Resources
Engineers, 1968, 1969). From some of their earlier work, Chen and Orlob
(1972) developed a model of Ecological Simulations for Aquatic Environments
which was used as the basis for many of the subsequent lake and reservoir
models.
One-dimensional lake models assume that mass and energy transfers only
occur along the vertical axis of a lake. To facilitate application of the
necessary mass and energy balance equations, the lake is represented as a
one-dimensional system of horizontal elements with uniform thickness, as
11-50
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TABLE II-4
DESCRIPTION OF WATER ANALYSIS SIMULATION PROGRAM
Name of Model:
Respondent:
Developers:
Year Developed:
Capabilities:
Aval lability:
Applicability:
Support;
Water Analysis Simulation Program (WASP)* -
LAKE1A, ERIE01, and LAKE3
William L. Richardson
U.S. Environmental Protection Agency
Large Lakes Research Station (LLRS)
9311 Groh Road
Grosse Isle, Michigan 48138
(313) 226-7811
Robert V. Thomann, Dominic DiToro, Manhattan College, N.Y.
1975 (LAKE1)
1979 (LAKE3)
Model is one (LAKE1) or three (LAKE3) dimensional and
computes concentration of state variable in each com-
pletely mixed segment given input data for nutrient
loadings, sunlight, temperature, boundary concentration,
and transport coefficients. The kinetic structure in-
cludes linear and non-linear interactions between the
following eight variables: phytoplankton chlorophyll,
herbivorous zooplankton, carniverous zooplankton, non-
living organic nitrogen (particulate plus dissolved),
ammonia nitrogen, nitrate nitrogen, non-living organic
phosphorus (particulate plus dissolved), and- available
phosphorus (usually orthophosphate). Also, a refined
biochemical kinetic structure which incorporates two
groups of phytoplankton, silica and revised recycle
processes is available.
Models are in the public domain
Large Lakes Research Station.
The model is general, however,
specific reflecting past studies.
User's Manual
and are available from
coefficients are site
A user's manual titled "Water Analysis Simulation Program"
(WASP) is available from Large Lakes Research Station.
Technical Assistance
Technicalassistance would be provided if requested in
writing through an EPA Program Office or Regional Office.
*The Advanced Ecosystem Model Program (AESOP) described next is a modified
version of WASP.
SOURCE: U.S. EPA, 1983c.
11-51
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TABLE II-5
DESCRIPTION OF WATER ANALYSIS SIMULATION PROGRAM
AND ADVANCED ECOSYSTEM MODELING PROGRAM
Name of Model
Respondent:
Developers;
Capabilities:
Verification:
Availability;
Applicability;
Water Analysis Simulation Program (WASP)
Advanced Ecosystem Modeling Program (AESOP)
John P. St. John
HydroQual, Inc.
1 Lethbridge Plaza
Mahwah, N.J. 07430
(201) 529-5151
WASP
Dominic M. DiToro, James J. Fitzpatrick, John L. Mancini,
Donald J. 0'Conner, Robert V. Thomann (Hydroscience, Inc.)
(1970)
AESOP
Dominic. DiToro, James J. Fitzpatriclc, Robert V. Thomann
(Hydroscience, Inc.) (1975)
and models
linear kinetics.
The Water Quality Analysis Simulation Program, WASP, may be
applied to one-, two-, and three-dimensional water bodies,
may be structured to include linear and non-
Depending upon the modeling framework the
user formulates, the user may choose, via input options, to
input constant or time variable transport and kinetic
processes, as well as point and non-point waste discharges.
The Model Verification Program, MVP, may be used as an
indicator of "goodness of fit" or adequacy of the model as
a representation of the real world.
AESOP, a modified version of WASP, includes a steady state
option and an improved transport component.
To date WASP has been applied to over twenty water resource
management problems. These applications have included
one-, two-, and three-dimensional water bodies and a number
of different physical, chemical and biological modeling
frameworks, such as BOD-DO, eutrophication, and toxic sub-
stances. Applications include several of the Great Lakes,
Potomac Estuary, Western Delta-Suisun Bay Area of San
Francisco Bay, Upper Mississippi, and New York Harbor.
WASP is in public domain and code is available from USEPA
(Grosse Isle Laboratory and Athens Research Laboratory).
AESOP is proprietary.
Models are general and may be applied to different types of
water bodies and to a variety of water quality problems.
11-52
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TABLE 11-5
DESCRIPTION OF WATER ANALYSIS SIMULATION PROGRAM
AND ADVANCED ECOSYSTEM MODELING PROGRAM (Concluded)
Support: User's Manual
WASP and MVP documentation is available from USEPA (Grosse
Isle Laboratory). AESOP documentation is available from
HydroQual.
Technical Assistance
Technicalasssistance of general nature from advisory to
implementation (model set-up, running, calibration/
verification, and analysis) available on contractura!
basis.
SOURCE: U.S. EPA, 1983c.
II-53
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TABLE 11-6
DESCRIPTION OF CLEAN PROGRAMS
Name of Model;
Respondent:
Developers;
Supporting Agency;
Year Developed;
Capabilities:
CLEAN, CLEANER, MS. CLEANER, MINI. CLEANER
Richard A. Park
Center for Ecological Modeling
Rensselaer Polytechnic Institute
MRC-202, Troy, N.Y. 12181
(518) 270-6494
Park, O'Neill, Bloomfield, Shugart, et al.
Eastern Deciduous Forest Biome
International Biological Program
(RPI, ORNL, and University of Wisconsin)
Thomas 0. Barnwell, Jr.
Technology Development and Application Branch
Environmental Research Laboratory
Environmental Protection Agency
Athens, Georgia 30605
1973 (CLEAN)
1977 (CLEANER)
1980 (MS. CLEANER)
1981 - estimated completion date for MINI. CLEANER
The MINI. CLEANER package represents a complete re-
structuring of the Multi-Segment Comprehensive Lake
Ecosystem Analyzer for Environmental Resources (MS.
CLEANER) in order for it to run in a memory space of
22K bytes. The package includes a series of simula-
tions to represent a variety of distinct environments,
such as well mixed hypereutrophic lakes, stratified
reservoirs, fish ponds and alpine lakes. MINI. CLEANER
has been designed for optimal user application—a turn-
key system that can be used by the most inexperienced
environmental technician, yet can provide the full
range of interactive editing and output manipulation
desired by the experienced professional. Up to 32
state variables can be represented in as many as 12
ecosystem segments simultaneously. State variables
include 4 phytoplankton groups, with or without surplus
intracellular nitrogen and phospho'rus; 5 zooplankton
groups; and 2 oxygen, and dissolved carbon dioxide.
The model has a full set of readily understood commands
and a machine-independent, free-format editor for
efficient usage. Perturbation and sensitivity analysis
can be performed easily. The model has been calibrated
and is being validated. Typical output is provided for
11-54
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TABLE II-6
DESCRIPTION OF CLEAN PROGRAMS (Concluded)
a set of test data. File and overlay structures are
described for -Implementation on virtually any computer
with at least 22K bytes of available memory.
Verification:
Availability:
Applicability:
Support;
The MINI. CLEANER model is being verified with data
from DeGray Lake, Arkansas; Coralvilie Reservoir, Iowa;
Slapy Reservoir, Czechoslovakia; Ovre Heimdalsvatn,
Norway; Vorderer Finstertak See, Austria; Lake Balaton,
Hungary; and Lago Mergozzo, Italy. The phytoplanktpn/
zooplankton submodels were validated for Yorderer
Finstertaler See.
Models are
Richard A.
Athens).
in public domain
Park (RPI) and.
and code is available from
Thomas 0. Barnwell (EPA/
Model is general.
User's Manual
A user's manual for MS. CLEANER is available from
Thomas 0. Barnwell, Jr. A user's manual for MINI.
CLEANER is in preparation.
Technical Assistance
Assistance may be available from the Athens Laboratory;
code and initial support is available for a nominal
service charge from RPI; additional assistance is
negotiable.
SOURCE: U.S. EPA, 1983c.
11-55
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TABLE II-7
DESCRIPTION OF LAKECO AND ONTARIO MODELS
Name of Model;
Respondent:
Developers:
User Developed;
Capabilities:
Verification:
Availability:
Applicability;
Support:
LAKECO*, ONTARIO
Carl W. Chen
Carl W. Chen
Tetra Tech Inc.
3746 Mount Diablo Blvd., Suite 300
Lafayette, California 94596
(415) 283-3771
(Original version developed when Dr. Chen was with Water
Resources Engineers)
1970 (original version)
LAKECO
Model is one-dimensional (assumes lake is horizontally
homogeneous) and calculates temperature, dissolved oxygen,
and nutrient profiles with daily time step for several
years. Four algal species, four zooplankton species, and
three fish types are represented. The model evaluates the
consequences of wasteload reduction, sediment removal, and
reaeration as remedial measures.
ONTARIO
Same as above but in three-dimensions for application to
Great Lakes. . ...
The models have been applied to more than IS lakes by Or.
Chen and to numerous other lakes by other investigators.
The model is in the public domain and the code is avail-
able from the Corps of Engineers (Hydrologic Engineering
Center), EPA and NOAA.
General
Tetra Tech, Corps of
User's Manual
User's manuals are available from
Engineers, EPA and NOAA.
Technical Assistance
Technical assistance is available and would be negotiated
on a case-by-case basis.
*A version of LAKECO, contained in a model referred to as Water Quality for
River Reservoir Systems (WQRSS) and supported by the Corps of Engineers
(Hydrologic Engineering Center), is described separately.
SOURCE: U.S. EPA, 1983c.
11-56
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TABLE II-8
DESCRIPTION OF WATER QUALITY FOR
RIVER RESERVOIR SYSTEMS
Name of Model: Water Quality for River Reservoir Systems (WQRRS)
Respondent:
Developers;
History:
Capabilities:
Verification:
Availability:
Applicability:
Support:
Mr. R.G. Willey
Corps of Engineers
609 Second Street
Davis, California 95616
(916) 440-3292
Carl W. Chen, 6.T. Orlob, W. Norton, D. Smith
Water Resources Engineers, Inc.
1970 (original version of lake eutrophication model)
1978 (initial version of WQRRS package)
1980 (updated version of WQRRS)
See description of LAKECO in Table II-7 (model also can
consider river flow and water quality).
Chattahoochee River (Chattahoochee River Water Quality
Analysis, April 1978, Hydrologic Engineering Center Project
Report)
Model is in public domain and code is available from Corps.
Model is general.
User's Manual
A user's manual is available from Corps.
Technical Assistance
Advisory assistance is available to all users. Actual exe-
cution assistance is available to federal agencies through
an inter-agency funding agreement.
SOURCE: U.S. EPA, 1983c.
11-57
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shown 1n Figure 11-16. Each hydraulic element is treated as a continuous-
flow stirred tank reactor (CFSTR) with completely uniform properties.
The implicit assumption of this geometric structuring of the problem is
that mass concentration and thermal gradients in the horizontal plane are
insignificant in determining the ecological responses and thermal behavior
of the impoundment along the vertical axis. Therefore, simulated results
are interpreted as being average conditions across the lake at a particular
elevation.
These models solve a set of equations representing the water quality of a
lake and the interactions of the lake biota with water quality. In
reality, an aquatic ecosystem exhibits a delicate balance of a mtrlttplicity
of different aquatic organisms and water quality constituents. Of neces-
sity, lake ecological models account only for the more significant inter-
actions in this balance.
An aquatic ecosystem is comprised of water, its chemical impurities, and
various life forms: bacteria, algae, zooplankton, benthos and fish, among
others. The biota responds to nutrients and to other environmental con-
ditions that affect growth, respiration, recruitment, decay, mortality and
predation. Abiotic substances derived from air, soil, tributary waters and
the activities of man, are inputs to the system that exert an influence on
the biotic structure of the lake. Figure 11-17 provides a conceptual
representation of an aquatic* ecosystem.
The fundamentalVtniildlng blocks (nutrients) for all living organisms are
the same: carbon, nitrogen and phosphorous. With solar radiation as the
energy source, these inorganic nutrients are transformed into complex
organic materials by" photosynthetic organisms. The organic products of
photosynthesis serve as food sources for aquatic animals. It is evident
that a natural succession up the food chain occurs whereby inorganic
nutrients are transformed to biomass.
Biological activities generate wastes which include dead cell material and
excreta which initially are suspended but may settle to the bottom to
become part of the sediment. The organic fraction of the bottom sediment
decays with an attendant release of the original abiotic substances. These
transformations are integral parts of the carbon, nitrogen and phosphorous
cycles and result 1n a natural "recycling" of nutrients within an aquatic
ecosystem.
The water quality and biological productivity of a lake vary in both time
and space. Temporal variations are associated with a wide variety of
external influences on a lake. Examples of these influences are. atmos-
pheric energy exchanges, tributary contributions and lake outflows.
Spatial variations occur both in the horizontal plane and with depth.
Variations in the horizontal plane are normally due to local conditions,
such as distance from shoreline, depth of water and circulation patterns.
Many times these variations do not affect the overall ecological balance of
a lake and are not modeled by the one-dimensional lake model.
11-58
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tributary
inflow
evaporation
tributary
inflow
control slice
outflow
Figure 11-16.
Geometric Representation of a Stratified Lake
(from Gaume and Duke, 1975).
11-59
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MAN-INDUCED
WASTE LOADS
NATURAL
INPUTS
DETRITUS
SEDIMENT
BENTHIC
ANIMAL
•
-------
Variations of water quality along the vertical axis of a lake have a more
general effect. The hydrodynamic behavior of a well-stratified lake is
density-dependent and, therefore, is related closely to the vertical tem-
perature structure of the impoundment. The vertical temperature structure,
in turn, is governed by the same external environmental factors as the
temporal variations, i.e., atmospheric energy exchanges, tributary con-
tributions and lake outflows.
EPA Center for Water Quality Modeling
The Center for Water Quality Modeling, located at the Environmental
Research Laboratory in Athens, Georgia, has long been involved in the
development and application of mathematical models that predict the
transport and fate of water contaminants. The Center provides a central
file and distribution point for computer programs and documentation for
selected water quality and pollutant loading models. In addition, the
Center sponsors workshops and seminars that provide both generalized train-
ing in the use of models and specific instruction in the application of
individual simulation techniques.
The water quality model supported by U.S. EPA for well-mixed lakes is the
Stream Water Quality Model QUAL-II (Roesner, et al., 1981). The model
assumes that the major transport mechanisms--advection and dispersion—are
significant only along the main direction of flow (longitudinal axis of the
lake). It allows for multiple waste discharges, withdrawals, tributary
flows, and incremental inflow. Hydraulically, QUAL-II is limited to the
simulation of time periods during which the flows through the lake are
essentially constant. Input waste loads must also be held constant over
time. QUAL-II can be operated as a steady-state.model or a dynamic model.
Dynamic operation makes it possible to study water quality (primarily
dissolved oxygen and temperature) as 1t 1s affected by diurnal variations
in meteorological data.
The Army Corps of Engineers have developed a numerical one-dimensional
model (CE-QUAL-R1), of reservoir water quality (U.S. Army Corps of
Engineers, 1982). The reservoir model is a direct descendant of the
reservoir portion of a model called "Water Quality for River-Reservoir
Systems" (WQRRS) which was assembled for the Hydrologic Engineering Center
of the Corps of Engineers by Water Resources Engineers, Inc. (Camp Dresser
4 McKee). The definitive origin of WQRRS was the work of Chen and Orlob
(1972).
The aquatic ecosystem and geometric representation of this model are sim-
ilar to those discussed in the previous section on one-dimensional lake
modeling. A summary of the model capabilities of CE-QUAL-R1 is given in
Table II-9.
Example Application of Mathematical Modeling
Mathematical modeling of natural phenomena allows planners, engineers,
biologists, and the general public to see the effects on the lake system of
changes in the environment which are planned or predicted to occur in the
future. This insight allows a state to assess the environmental responses
11-61
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TABLE II-9
CE-QUAL-R1 MODEL CAPABILITIES
Factors considered by CE-Q.UAL-R1 Include the following:
a. Physical Factors
(1) Shortwave and longwave solar radiation at the water surface.
(2) Net heat transfer across the air-water Interface.
(3) Convectlve and radiative heat transfer within the water body.
(4) Convectlve mixing due to density Instabilities.
(5) Placement of Inflowing waters at depths with comparable
density.
(6) Withdrawal of outflowing waters from depths Influenced by the
outlet structure and density stratification.
(7) Conservative substance routing.
(8) Suspended sol Ids routing and settling.
b. Chemical and Biological Factors
(1) Accumulation, dispersion, and depletion of dissolved oxygen
through aeration, photosynthesis, respiration, and organic
demand.
(2) Uptake-excretion kinetics and regeneration of nitrogen and
phosphorus and nitrification processes under aerobic condi-
tions.
(3) Carbon cycling and dynamics and alkal1nity-pH-C02 inter-
actions.
(4) Phytopiankton dynamics and trophic relationships.
(5) Transfers through higher trophic levels of the food chain.
(6) Accumulation, dispersion, and decomposition of detritus and
sediment.
»
(7) Coll form bacteria die-off.
(8) Accumulation, dispersion, and reoxidation of manganese, iron,
and sulffde when anaerobic conditions prevail.
SOURCE: U.S. Army Corps of Engineers, 1982.
11-62
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of the lake and help it to analyze alternative plans for protecting the
present use or determining what uses could be attained.
External factors, such as increased nutrients which accelerate the growth
of algae, may destroy the delicate balance of nature, and cause consider-
able harm to the lake and its biology. Therefore, it is important to be
able to predict what the lake response will be to external factors without
actually imposing those conditions on it. The mathematical portrayal of
the lake ecosystem by the computer model helps us toward that end.
As an example, the lake ecological model EPAECO (Gaume and Duke, 1975) pro-
vided a tool to mathematically represent the aquatic ecological system in
the Fort Loudoun Lake, Tennessee. This study was conducted as part of the
208 plan for the Knoxville/Knox County Metropolitan Planning Commission
(Hall, et al., 1976). The 208 study area map is shown in Figure 11-18. In
general, the model EPAECO is designed to simulate the vertical distribution
of the following constituents over an annual cycle:
1. Temperature 10. Total Inorganic Carbon
2. Total Dissolved Solids 11. Carbon Dioxide
3. Alkalinity 12. Hydrogen Ion (pH)
4. Coliforms 13. Dissolved Oxygen
5. Carbonaceous Biochemical 14. Algae (two classes)
Oxygen Demand (CBOD) 15. Zooplankton
6. Ammonia Nitrogen 16. Fish (three classes)
7. Nitrite Nitrogen 17. Benthic Animals
8. Nitr^> Nitrogen 18. Organic Sediment, and
9. Phoi. torus 19. Suspended Detritus.
The general approach to use of the mathematical model EPAECO .is to obtain
data which describe the geometric properties of the lake and its past
history of water quality and hydrodynamics. Data on water quantity and
quality of tributary inputs to the lake (streams and/or waste loads) and
meteorological data are also necessary. Initially, the lake must be
described as a mathematical system of depths, areas, volumes, tributary
inputs and releases. A site-specific model must be developed which
properly describes the environmental community and its interactions for
Fort Loudoun Lake. This is done by a procedure called calibration. A
calibrated model gives the user greater confidence that the simulation
model will react as would the lake itself to changes in external factors
such as increased tributary nutrient concentrations.
Examples of calibration results are shown in Figures 11-19 through 11-21.
Figure 11-19 presents the observed and simulated reservoir elevations for
the year 1971; Figure 11-20 shows the vertical temperature profiles,
observed and simulated, for the months of April, May and July, 1971; and
Figure 11-21 gives the observed and simulated profiles for several water
quality constituents for a single day in September 1971.
One of the main considerations in the study of Fort Loudoun Lake was an
evaluation of present and future trophic states. Lakes which become en-
riched with excessive nutrients may be defined as eutrophic. Eutrophica-
tion produces large algal communities which affect the taste and odor of
the lake's waters. Bacteria which degrade the large amounts of dead
11-63
-------
t
Ok
LEGEND
Knox County (208 Area)
.Fort Loudoun Drainage Area
HEROKEE DAM
FT.
LOUDOUN
0AM
Figure 11-18. 208 Study Area (from Hall et al. 1976)
-------
27.5-1-
27 -J-
26.5--
o
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25J
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245--
24
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100 150 200
JULIAN DAYS, 1971
n
x Q
x O x
00
250
--8IO _
—8C3
— 80S
o
T
300
350
KEYj
O OBSERVED
SIMULATED
Figure 11-19.
Fort Loudoun Reservoir Elevations 1971 Observed vs.
Simulated (from Hall et al, 1976)
11-65
-------
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STATION2(R M 601 2)
• STATION 3(f1.M6IS8|
SIMULATED
PHOSPHORUS:
Q STATION2(RM.6Oi 21
• STATION 3(RM.615.0)
SIMULATED
NITROGEN:
O STATION2(RM.6OJ 2)
• STATlONllRM.6lS.al
SIMULATED
ALGAE.
O STATION 2 (RM.6O1 2)
• STATION 3(a
SIMULATED
NOTE:
•NITROGEN -NtliN»NOi N»N02 N
Figure 11-21- DO and'BODs, Inorganic Phosphorus and Nitrogen, and Algae
September 10, 1971 Fort Loudoun (from Hall et al, 1976)
-------
organic matter in the lake deplete the oxygen supply, which in turn results
in a loss of some types of fish. Excessive aquatic weed growth is also
detrimental to swimming, boating and fishing.
The model EPAECO was used to assess algal growth as a result of various
nutrient loads (high, medium and low) to the lake during the period of May
through September. This type of model application not only quantified the
degree of expected algal growth as a function of the availability of
nutrients but also predicted the algal population and total lake ecology
for future nutrient loads to the lake.
Since phosphorus was the limiting nutrient for algal growth in this lake
study, the total available phosphorus was compared to the maximum seasonal
algal concentrations simulated for the sensitivity study. Figure 11-22
shows this comparison. The curve is derived from the maximum algal con-
centrations resulting from the following sensitivity conditions: high P,
medium P, and low P. This curve represents the maximum algal concentra-
tions reached by a constant inflow concentration of phosphorus during the
algal growing season.
A limited amount of phosphorus is required in the inflows to the stratified
portion of the reservoir to support a desirable algal community without
producing excess growth and thus undesirable conditions. As shown on the
graph in Figure 11-22, fort Loudoun Lake phosphorus concentrations in the
range of 0.013-0.037 mg/1 produced algal concentrations which were suitable
for a well-balanced ecosystem with good water quality as observed in 1971
by the Tennessee Valley Authority.
11-68
-------
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-------
CHAPTER III
BIOLOGICAL CHARACTERISTICS
INTRODUCTION
This chapter contains information about the characteristic plants and
animals found in lakes and provides an overview of the water quality and
the types of habitat that they require. The chapter is divided into major
sections: Plankton, Aquatic Macrophytes, Benthos, and Fish.
Particular emphasis is placed on changes in species composition as lakes
progress from oligotrophy to eutrophy. The biota of lakes is often studied
to assess the trophic state or biological health of the water body. Thus,
indicator organisms are also discussed in this chapter, along with
qualitative and quantitative methods of assessing the biological health of
a lake. The reader is referred to the Techni cal Support Manual: Water
Body Surveys and Use Attainability Analyses (U.S. EPA,1983bJ where an
extensive discussion on species diversity and other measures of community
health will be found.
PLANKTON
Planktonic plants and animals are important members of the lacustrine food
web. Phytoplankton, which comprise pigmented flagellates, green and blue-
green algae, and diatoms, are lowest on the food chain and serve as a
primary food source for higher organisms. Zooplankton may be grazers
(consuming phytoplankton) or predators (feeding on species smaller than
themselves). The zooplankton, in turn, serve as the primary food source
for the young of many fish species. The findings of various authors who
have studied the effects of organic pollution and nutrient enrichment on
the lacustrine plankton are summarized below.
Phytoplankton
The growth of phytoplankton is normally limited by the amount of nitrogen
and/or phosphorus available. When increased quantities of nutrients enter
the lake in runoff or effluents, eutrophication with its attendant
uncontrolled algal growth and its consequences may begin. For example, the
production of toxic substances by some algae may cause human gastrointes-
tinal, skin and respiratory disorders, while blooms of Microcystis and
Nostoc rivulare may poison wild and domestic animals, causing unconscious-
ness, convulsions and sometimes death (Mackenthun, 1969).
Al.gal blooms affect the dissolved oxygen (DO) content of the water.
Diurnal fluctuations of 00 and pH become more pronounced with large algal
populations. In addition, the dissolved oxygen in the hypolimnion is
depleted through algal death and decay, leading to anoxic conditions. Fish
may die because of anaerobic conditions or the production of toxic
substances. Water quality problems caused by algae, such as taste and
odor, are especially troublesome if the water body is used as a source of
drinking water. Finally, scums and mats of the algae destroy the aesthetic
value of the lake.
II-I-l
-------
Since some species are able to compete better than others, Increased
nutrients cause changes In phytopiankton community composition. Thus,
specific algal associations may be indicative of eutrophic conditions.
Indices of trophic state based on phytoplankton taxon are also related to
the degree of eutrophy. The use of phytoplankton as indicators of
eutrophication is discussed below.
Qualitative Response to Environmental Change
The identification of phytoplankton that are commonly found in eutrophic
and oligotrophic lake waters has resulted in lists of pollution tolerant/
intolerant genera and species. Palmer (1969) developed several lists of
pollution tolerant algal genera and species by compiling information in 269
reports by 165 authors. The eight most tolerant genera were Euglena,
Oscillator! a, Chlamydomonas, Scenedesmus. Chlorella, Nitzchia, Navicula,
and Stigeoclonium. The five most tolerant species were Euglena viridis,
NitzcHTapa lea, Oscillator! a limosa, Scenedesmus quadrlcauda^ ami
Oscillator! a tenuisi Palmer used the following method to combine the works
of the various authors: A score of 1 or 2 points was given for each algae
reported by an author as tolerating organic enrichment, the larger figure
being reserved for:the algae that an author emphasized as being typical of
waters with high organic pollution. The compilation by Palmer is presented
in Appendix A, pollution-tolerant genera and pollution-tolerant species.
Palmer's listings have been criticized because the information used to
compile them came from a broad range of sources and geographical areas. In
addition, the compilation 1s restricted to algae tolerating high organic
pollution. Thus, the listing may not be valid for other types of pollu-
tants. Nevertheless, it does provide an indication of relative tolerance
to organic pollution.
Taylor, et al. (1979) studied the environmental conditions associated with
phytoplankton genera. The occurrence of 57 genera was related to total
phosphorus levels, total Kjeldahl nitrogen levels, chlorophyll^a levels,
and N/P ratio values. Most genera were found to occur over extremely wide
ranges or conditions. The seven genera associated with levels of phos-
phorus greater than 200 ug/1 were found to also represent seven of the
eight highest chlorophyll-^ values. Taylor designated this group con-
taining Actinastrum, Anabaenopsis, Schroederia, Raphidiopsis. Chlorogonium,
GolenkioTTj and Lage'rhelmla as the "nutrient rich genera". All seven
genera were summer and fall forms, while Act1nastrum and Lagerheimia also
occur in spring.
The "nutrient-poor" group, containing five genera, were associated with
total phosphorus levels less than 70 ug/1. Asterionella. Dinobryon.
Tabellaria, Peridinium, and Ceratiurn make up this group. Asterionella is
the only genus occurring solely in spring. The other genera occur in
summer and fall; Dinobryon and Tabellaria also occur equally in spring,
summer and fal1.
Taylor, et al. (1979) also noted which genera achieved numerical dominance
most frequently 1n the lakes studied. Melosira was the most dominant
genus, followed by Oscillator!a and LyngbyTI Asterionella was considered
spring dominant, while Stephanodiscus. Synedra and Tabellaria were
III-2
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categorized as spring and summer dominant. Fragilaria occurred equally
throughout the seasons as a dominant, and the remaining genera were summer
and fall dominant. Additional information about the environmental
conditions associated with the presence of the 20 phytoplankton genera most
frequently recorded as dominants is available in Taylor, et al . (1979).
The study by Taylor, et a 1. (1979) concluded the following: (1) Phyto-
plankton genera survive over such a broad range of environmental conditions
that they cannot be used as indicator organisms; (2) No phytoplankton
genera emerged as dependable indicators of any one or combination of the
environmental parameters measured; (3) Preliminary analyses suggest that
phytoplankton community composition shows promise for use in water quality
assessment; (4) Some taxa, e.g., Pediastrum and Euglena, were very frequent
components of phytoplankton communities, but rarely achieved high relative
numerical importance within those communities; (5) Hagellates and diatoms
were the most common springtime plankton genera, while the blue-green and
coccoid green genera were most common in the summer and fall; and (6) Blue-
green algal forms, including se'veral not known to fix elemental nitrogen,
contributed 9 of the 10 genera which attained numerical dominance in water
with a mean inorganic nitrogen/total phosphorus ratio (N/P) of less than 10
(generally suggestive of nitrogen-limitation).
Similarly, Bush and Welch (1972) concluded that phosphorus availability was
most critical to the biomass formation of blue-green algae. They found
that Apham'zomenon and Micnxystis formed mats on the water surface during
warm summer days, and were typical of shallow, hypereutrophic lakes such as
Clear Lake (California), Klamath Lake (Oregon) and Moses Lake (Washington).
Their study showed that the biomass of blue-green algae was related to in-
organic phosphate even when nitrate was low and invariable.
Harris and Yollenweider (1982) noted some diatoms that are characteristic
of oligotrophic lakes. Species of Tabellaria, Fragilaria, and Asterionella
indicated oligctrophic conditions. In sediment cores of Lake Erie, species
of Melosira showed the transition from oligotrophic to eutrophic condi-
tions"!The succession of species was as follows: Melosira distans and M.
italica were present prior to 1850 and are considered indicative of oligo-
trophy; after 1850, _M. distans and ^. italica populations dwindled, and M.
islandica (moderate enrichment) and M. "granulata (eutrophication indicator)
appeared in the core; in the next phase, around 1960, ^. distans disap-
peared and was replaced by ^. binderana.
Quantitative Response to Environmental Change
Because phytoplankton exhibit such a broad range of tolerance to environ-
mental conditions, the presence or absence of a single species is not
necessarily indicative of trophic state. In contrast, indices based on
dominant genera, community composition, cell count, or chlorophyll-a^
provide a useful assessment of lake trophic levels and are better suited to
the classification of lakes than single species evaluations.
Chlorophyll-a. Chlorophyll-a is a widely accepted index of algal biomass.
In lakes ancT reservoirs with retention times greater than 14 days, it is
highly correlated with phosphorus. The correlation does not hold for
III-3
-------
systems with less than 14-day retention times (U.S. EPA, 1979^). Estimates
of chlorophyll-^ values indicative of trophic state are shown in Table
III-l.
Carlson's Trophic State Indices. Carlson (1977) developed three indices of
trophic state, based upon Secchi depth, total phosphorus and chlorophyll-^.
The three indices are defined below:
Carlson's Secchi Depth Index, TSI (SO) = 10(6 - ^p|) (1)
Carlson's Chlorophyll-a. Index, TSI(CHL) = 10(6 - 2tQ4"0^821n CHL) (2)
Carlson's Total Phosphorus Index, TSI(TP) * 10(6 - 1"1^TP) (3)
where
SD = Secchi disc depth, m
CHL = Concentration of chlorophyll-^, ug/1
TP = Concentration of total phosphorus, ug/1.
The scale of values for Carlson's Secchi Depth Index ranges from zero to
greater than 100. A Secchi depth transparency of 64 m, which is greater
than the highest value reported for any lake in the world, yields a value
of zero. A Secchi depth of 32 m corresponds to an Index value of 10. An
Index value of 100 represents a transparency of 0.062 m. Using empirically
determined relationships between total phosphorus and transparency, and
chlorophyll-^ and transparency, Carlson developed equations (1), (2) and
(3). These equations arrive at the same trophic state index value, regard-
less of whether Secchi depth, total phosphorus, or chlorophyll-^ is the
parameter used. However, it is desirable to evaluate all three indices
because of non-nutrient related factors (temperature, inorganic turbidity,
toxics) which may affect productivity and cause disagreement among the
indices.
Based on observations of several lakes, most oligotrophic lakes had TSI
below 40, mesotrophic lakes had TSI between 35 and 45, and most eutrophic
lakes had TSI greater than 45. Hypereutrophic lakes may have values above
60 (Novotny and Chesters, 1981; Uttormark and Hutchins, 1978).
Nygaard's Trophic State Indices. Nygaard (cited by Sullivan and Carpenter,
1982)developedfivephytoplankton indices (myxophycean, chlorophycean,
diatom, euglenophyte, and compound) based on the assumption that certain
algal groups are indicative of various levels of nutrient enrichment. He
assumed that Cyanophyta, Euglenophyta, centric diatoms, and members of
Chlorococcales are typical of eutrophic waters, while desmids and many
pennate diatoms are generally found in oligotrophic waters. Nygaard's
indices are listed 1n Table III-2. In applying these indices, the number
of taxa in each major group is determined from the species list for each
sample (U.S. EPA 1979^).
II1-4
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TABLE III-l
TROPHIC STATE VS. CHLOROPHYLL-a.
Chlorophyll-^ (ug/1)
Trophic
Condition
01 i go trophic
Mesotrophic
Eutrophic
Sakamoto,
1966
0.3-2.5
1-15
5-140
National Academy
of Sciences,
1972
0-4
4-10
>10
Dobson, et al., U^S. EPA,
1974 1974
0-4.3 <7
4.3-8.8 7-12
>8.8 >12
SOURCE: U.S. EPA, 1979a.
III-5
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TABLE II1-2
NYGAARO'S TROPHIC STATE INDICES
Index Calculation Oligotrophic Eutrophic
Myxophycean Myxophyceae 0.0-0.4 0.1-3.0
Oesmideae
Chlorophycean Chlorococcales 0.0-0.7 0.2-9.0
Desmideae
Diatom Centric Diatoms 0.0-0.3 0.0-1.75
Pennate Diatoms
Euglenophyte Euglenophyta ^_^ 0.0-0.2 0.0-1.0
IMyxophyceae + Chlorococcales)
Compound (Myxophyceae + ChlorococcaTes * 0.0-1.0 1.2-25
Centric Diatoms + Euglenophyta)
Desmideae
SOURCE: U.S. EPA, 1979a.
III-6
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Nygaard's ranges show considerable overlap between trophic states.
Sullivan and Carpenter (1982) sampled 27 lakes and reservoirs and found
that Nygaard's indices did not differentiate between trophic states. In
addition, an index value is undefined whenever the denominator is zero.
Palmer's Organic Pollution Indices. Palmer (1969) developed two algal
pollution indices (genus and species) for rating water samples with high
organic pollution. After reviewing reports of 165 authors, Palmer prepared
two lists of organic pollution-tolerant forms, one containing 20 genera
(Table III-3), and the other, 20 species (Table III-4).
In analyzing a water sample, any of the 20 genera or species present in
concentrations of 50/ml or more are recorded. The pollution index numbers
of the algae present are then totaled, giving a genus score (Palmer's Genus
Index) and a species score (Palmer's Species Index). A score of 20 or more
is taken as evidence of high organic pollution, while a score of 15 to 19
is taken as probable evidence of high organic pollution. Lower figures
indicate that the organic pollution of the sample is not high, or that some
substance or factor interfering with algal persistence is present or active
(Palmer, 1969).
Use of Palmer's indices in a study of Indiana lakes and reservoirs showed
that the Genus Index was more sensitive to differences among samples than
the Species Index. The Genus Index was correlated with the degree of
eutrophication, reflecting the abundance of eutrophic indicator genera.
Another advantage of the Genus Index is that genera are easier to identify
than species. However, a study of 250 lakes in the eastern and south-
eastern states showed that Palmer's indices were poorly correlated with
summer mean phosphorus and chlorophyll-^ levels, although the Genus Index
ranked higher (Spearman's rank correlation coefficient) than the Species
Index (U.S. EPA, 1979^).
U.S. EPA Proposed Phytoplankton Indices of Trophic State. Using a test set
of 44 lakes in the eastern and southeastern states, EPA compared the
abilities of several indices to measure trophic state (U.S. EPA, 1979ja).
The same report introduced 10 additional indices that used a combination of
data including total phosphorus, chlorophyll-£, Kjeldahl nitrogen, phyto-
pi ankton genera counts and cell counts/ml.
Each genus was assigned "trophic values" based on mean parameter values
associated with the dominant occurrence of that genus. The data used to
assign trophic values was taken from studies of 250 lakes that were sampled
during spring, summer and fall of 1973. Trophic values used in the general
formulas of the new indices (Table III-5) are presented in Appendix 8,
along with sample problems using the indices.
When the newly developed indices were compared to Nygaard's and Palmer's
indices, they showed a consistently stronger correlation with summer mean
phosphorus levels and chlorophyll-a^ levels. When applied to the dominant
phytopiankton community components, the indices generally had higher
correlations than the analogous indices applied to all phytoplankton
community components, although the differences were small (U.S. EPA 1979aJ .
III-7
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TABLE III-3
TABLE 111-4
VALUES USED IN ALGAL
GENUS POLLUTION INDEX
VALUES USED IN ALGAL
SPECIES POLLUTION INDEX
Genus
Anacystis
Anki strodesmus
Chi amydomonas
Chlorella
Closterium
Cyclotella
Euglena
Gomphonema
Lepocinclis
Melosira
Mi cr actinium
Navieula
Nitzschia
Oscillatoria
Pandorina
Phacus
Phormidium
Scenedesmus
St1 geocl oni urn
Synedra
Pollution
Index
1
2
4
3
1
1
5
1
1
1
1
3
3
5
1
2
1
4
2
2
Species
Pollution
Index
Ankistrodesmus falcatus
Arthrospira jenneri
Chlorella vulgaris
Cyclotella meneghiniana
Euglena gracilis
Euglena viridis
Gomphonema parvulum
Melosira varians
Navicula cryptocephala
Nitzschia acicularis
Nitzschia palea
Oscillatoria chlorina
Oscillatoria limosa
Oscillatoria princeps •
Oscillatoria putrida
Oscillatoria tenuis
Pandorina morum
Scenedesmus quadricauda
Stigeoclonium tenue
Synedra ulna
3
2
2
2
1
6
1
2
1
1
5
2
4
1
1
4
3
4
3
3
SOURCE: Palmer, 1969.
SOURCE: Palmer, 1969.
III-8
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TABLE II1-5
EPA PROPOSED PHYTOPLANKTON INDICES TO TROPHIC STATE
Phytoplankton Trophic State Index (TSI) Calculations Without Cell Counts:
TSI = I
V./n
1
n = number of dominant genera In the sample (Concentration - 10 percent of
the total sample concentration).
Vj* = the trophic value for each dominant genus in the sample; TOTAL? (PO),
CHLA (PD), KJEL (PD), MV (PD); MV = Log TOTALP + Log CHLA + Log KJEL -
Log SECCHI
Phytoplankton Trophic State Index (TSI) Calculations with Cell Counts:
TSI = 25 V c
1-1 1
Total Community:
n = the number of genera in the sample (entire phytoplankton community)
C = the concentration of the genus in the sample (units/ml)
V = the trophic value for each genus;
TOTALP/CONC(P), CHLA/CONC(P), KJEL/CONC(P)
Dominant Community:
n = the total number of dominant genera in the sample
C = the concentration of the genus in the sample (units/ml)
V = the trophic value for each genus;
TOTALP/CONC (P), CHLA/CONC (PD), KJEL/CONC (PO)
*The parameters TOTALP, CHLA, etc. are defined in Appendix B.
SOURCE: U.S. EPA, 1979^.
III-9
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Zooplankton
As lakes become enriched, phytopiankton and (to a large degree) herbivorous
zooplankton populations increase. Changes in species composition also
occur, although it is difficult to classify the trophic state of a water
body on the basis of a list of zooplankton species living in it.
Generally, larger species of zooplankton dominate in oligotrophic waters.
This is probably largely due to predation pressure. In eutrophic waters,
whe^e the fish stock is heavy, the larger zooplankton are eaten first.
Thus, the number of zooplankters that attain a large size is limited.
Species of Bosmina have been commonly accepted as indicators of enrichment.
Hutchinson (1967) observed that Bosmina cpregoni longispina appeared to be
characteristic of larger and less productive lakes, and B. longirpstri; of
smaller and more productive lakes. Studies on the sediments of linsley
Pond, Connecticut (Deevy, 1940), indicated that the disappearance of 8.
coregoni longispina was concurrent with the appearance of B. longirostrTs
as the lake became enriched. However, the collection of T. longirostris
from the epilimnion, and B_. coregoni from the hypo limn ion of another lake
shows the uncertainty of using Sosmina spp. as indicators.
Studies of zooplankton in the Great Lakes showed the following:
1. A decreased significance of calanoids and an increased predomi-
nance of cyclopoids and cladocerans were seen as a general trend
from oligotrophic Lake Superior to eutrophic Lake Erie (Patalas,
1972; Watson, 1974).
2. Larger zooplankton were observed in Lakes Superior and Huron,
although Lake Erie had an increased biomass of zooplankton
(Patalas, 1972; Watson, 1974).
3. In Lake Michigan, Bosmina coregoni has been replaced by B. longi-
rostris, Diaptomus oregonensis has become an important copepod
species, Eurytemora affinis appeared (Beeton, 1969).
4. Diaptpmus siciloides, usually found in eutrophic waters has become
a dominant zooplankton in Lake Erie (Beeton, 1969).
Some rotifers have been considered indicators of eutrophied waters. How-
ever, these organisms (in particular, Srachionus and Keratella quadrata)
have also been collected from oligotrophic lakes. Other zooplankton are
difficult to identify and thus are not practical to use as indicators of
water quality. For example, Cyclops scutifer is principally an oligotro-
phic form while Cyclops scutifer wigrensis lives in meso- and eutrophic
lakes (Ravera, 1980).
Sprules (1977) developed a technique for predicting the limnological
characteristics of a lake which is based on its midsummer limnetic crus-
tacean zooplankton community. The results indicated that northwestern
Ontario lakes characterized by Cyclops bicuspidatus thomasi, and Oiaptomus
minutus are generally large and clear, whereas Tropocyclops prasinus
mexicanus and Diaptpmus minutus are typical of smaller lakes with lower
water clarity.Acidic, small and clear lakes of the Killarney region,
111-10
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Ontario, are dominated by Diaptomus mlnutus, while Diaphanosoma leuch-
tenbergianum, Bosmina Iqngirostris and Mesocyclops edax dominate in lakes
that are less clear, larger and have a higher pH. Finally, in the
Haliburton region of Ontario, small and productive lakes are characterized
by Diaptomus oregonensis, M. edax, and Ceriodaphnia lacustris. Those lakes
with D. minutus, D_. sicilTs, B_. longirostris and Daphnia duba are larger
and less productive.
Thus, the direct effects of nutrient enrichment on the zooplankton are
unclear. Although a few qualitative changes have been mentioned, the only
quantitative information refers obliquely to diversity indices. The
diversity of the zooplankton community generally decreases with increasing
enrichment, as do the other organism communities. Diversity Indices are
discussed in the Technical Support Manual: Water Body Surveys and Assess-
ments for Conducting Use Attainability Analyses (1983bj.
AQUATIC MACROPHYTES
Aquatic plants play several roles in the lake ecosystem. They produce
oxygen through photosynthesis, shade and cool sediments, diminish water
currents and provide habitat for benthic organisms and fish (3oyd, 1971).
Carignan and Kalff (1982) found that water milfoil (Myriophyllum spicatum
L.) was important as physical support for micrcbial communities. Submersed
macrophytes serve as food and nest sites for aquatic insects and fish, and
provide protection from predation. The plants also play a role in nutrient
cycling, especially in the mobilization of phosphorus from sediments.
Barko and Smart (1980) investigated the uptake of phosphorus from five
different sediments by Egeria densa, Hydri11 a verticillata, and Myrio-
phyl1 urn spicatum. The amount of sediment phosphorus mobilization differed
among species and sediments, but it was demonstrated that the plants were
able to. obtain their phosphorus nutrition exclusively from the sediments.
Release of phosphorus from the macrophytes occurred primarily through death
and decay rather than through excretion. Landers (1982) showed that decom-
posing Myriophyllum spicatum supplied significant amounts of nitrogen and
phosphorus to surrounding waters. Nitrogen inputs accounted for less than
2.2 percent of annual allochthonous inputs, but phosphorus recycling from
decaying plants equaled up to 18 percent of the total annual phosphorus
loading for the reservoir studied.
Response of Macrophytes to Environmental Change
Major environmental changes in lakes generally occur in response to nutri-
ent increases (which accelerate eutrophication), suspended sediment, and
sediment deposition. Suspended sediment attenuates1 light penetration,
resulting in reduced photosynthesis by submerged aquatic macrophytes, and a
possible decrease in the coverage by plants. Reed, et al. (1983) noted
that the growth of Chara in a test pond was restricted during years when
the turbidity was high, but luxurious stands developed when the water was
clearer. Sediment deposition smothers some plants. For example, Isoetes
lacustris is not present in areas with rapid silting, but Nitell a and
Juncus often occur instead (Farnworth, 1979). Potamogeton perfoliatus may
also replace Isoetes where silting occurs. The composition of the sub-
strate is important in the growth of macrophytes. Potamogeton perfoliatus,
El odea canadensis, and Myriophyllum spicatum reportedly grew more rapidly
III-ll
-------
in natural sediment than in sand. Lobelia dortmanna grew only in sand
containing organic matter (Farnworth, 1979).
Although aquatic macrophytes are vital to the ecosystem, eutrophication and
the subsequent overgrowth of plants may be detrimental to the water body.
Diurnal DO fluctuations driven by photosynthesis and respiration may be so
extreme that oxygen deficits occur. Oxygen depletion in the hypo!imnion
may also be caused by decaying macrophytes. Low DO may cause fish kills
and eliminate sensitive species (Boyd, 1971).
Although eutrophication is often considered the cause of changes in macro-
phyte composition, management techniques may also be responsible. Nicholson
(1981) argued that techniques such as herbicidal poisoning and mechanized
cutting were primary reasons for the replacement of native Potamogeton
species in Chautagua Lake, New York, by Potamogeton crispus and Myr:To^'
phyllum spicatum.
Preferred Conditions
Certain aquatic plants are able to "out-compete" others and in large popu-
lations become established under eutrophic conditions. Such excessive
growth is usually undesirable, and the plants are considered aquatic weeds.
Aquatic plants that cause difficulty in the United States include Myrio-
phyllum spicatum var. exalbescens (water milfoil), Potamogeton crispus
(curly-leaved pondweed), Eichornia crassipes (waterhyacinth), Pistia
stratioles (water lettuce), Alternanthera phTToxeroides (alligator weed),
Heteranthera dubia (water stargrass), Myriophyllum brasiliense (parrot
feather),M. spicatum var. spicatum (eurasianwatermilfoil), Najas
guadalupensTs (southern naiad), Potamogeton pectlnatus (sago pondweed),
El odea canadensis (el odea), and Phragmites communis (common weed).
Seddon (1972) investigated the environmental tolerances of certain aquatic
macrophytes found in lakes. He grouped the species into the following:
1. Tolerant species-that occur over a wide range of solute concentra-
tions - Potamogeton natans, Nuphar lutea, Nymphaea alba, Glyceria
fluitans. Littorella urn'flora;
2. Highly eutrophic species - Potamogeton pectinatus, Myriophyllum
spicatum;
3. Moderately eutrophic species - Potamogeton crispus, Lemna trisulca;
4. Species tolerant of mesotrophic as well as eutrophic conditions -
Ranunculus circinatus, Lemna minor, Polygonum amphibium, Cera-
tophyllum demersum, Potamogeton obtusifolius;'
5. Species of oligotrophic tolerance - Potamogeton perfoliatus,
Ranunculus aquatilis, Apium inundatum, El odea canadensis, Pota-
mogeton berchtoldii.
Plants occurring only in eutrophic conditions were considered restricted to
such areas by physiological demands. It should be noted that the last
group, although classified as of oligotrophic tolerance, may also be found
111-12
-------
in eutrophic waters. Oligotrophic species, while shown to have a wide
tolerance, are thought to be excluded by competition rather than by
physiological limitation from sites with higher trophic status. The last
group in effect includes those species that can adapt to the relatively
nutrient free conditions of oligotrophic water.
BENTHOS
Benthic macroinvertebrates are often used as indicators of water quality.
Because they are present year-round, are abundant, and are not very motile,
they are well-suited to reflect average conditions at the sampling point.
Many species are sensitive to pollution and die if at any time during their
life cycle they are exposed to environmental conditions outside their
tolerance limits.
There are also disadvantages to basing the evaluation of the biotic integ-
rity of a water body solely on macroinvertebrates. Identification to the
species level is time-consuming and requires taxonomic expertise. Further-
more, the results may be difficult to interpret because life history in-
formation is lacking for many species and groups, and because a history of
pollution episodes in the receiving water may not be available to provide
perspective for the interpretation of results.
Certain organisms and associations of organisms point to various stages of
eutrophy. Decay of organic material often decreases the 00 (dissolved
oxygen) content of the hypolimnion below the tolerance of the inverte-
brates. Attempts to translate the results of studies into meaningful
values have yielded lists (presented later in this section) of tolerant and
intolerant groups of macroinvertebrates. In addition, mathematical for-
mulas have been developed which assign numerical values to various trophic
states depending upon the benthos present. However, factors other than
organic pollution (e.g., substrate, temperature, depth) may also influence
the species composition of benthic populations. Parameters such as these
which govern species distribution are' discussed in Merritt and Cummins
(1978).
Composition of Benthic Communities
The composition of the benthos in littoral and profundal areas of a lake is
mostly dependent upon substrate, but is also influenced by depth, temper-
ature, light penetration and turbidity. The littoral regions of lakes
usually support larger and more diverse populations of benthic inverte-
brates than profundal areas (Moore, 1981). Benthic communities in the
littoral regions consist of a rich fauna with high oxygen demands.
The vegetation and substrate heterogeneity of the littoral zone provide an
abundance of microhabitats occupied by a varied fauna. By contrast, the
profundal zone is more homogeneous, becoming more so as lakes become more
eutrophic (Wetzel, 1975). One of the best illustrations of the differences
of littoral and profundal benthos is seen in studies of Lake Esrom, a
dimictic lake in Denmark (Jonasson, 1970). The bottom fauna found on sub-
surface weeds (depth about 2m) comprises thirty-three groups and species,
totaling 10,810 individuals per square meter. In contrast, only five
species are found in the profundal zone of Lake Esrom, although the density
111-13
-------
is high (20,441 per square meter). The animals in this region burrow into
the bottom instead of living on or near the surface.
The factors mentioned above should be considered in the design of a study
of lake benthos. Because substrates of deep waters generally have finer
sediment particles than substrates of shallow waters, depth should be
considered in quantitative calculations to help compensate for substrate
differences. Adjustments for depth will be discussed in greater detail in
the section on quantitative measures of the effects of pollution on
benthos.
General Response to Environmental Change
The benthos of freshwater is composed largely of larvae and nymphs of
aquatic insects (Arthropoda: Insecta). The benthos also comprises fresh-
water sponges (Porifera: Spongillidae), flatworms (Platyhelminthes:
Tricladida), leeches (Annelida: Hirudinea), aquatic earthworms (Annelida:
Oligochaeta), snails (Mollusca: Gastropoda), clams and mussels (Mollusca:
Bivalvia). Particular groups of insects are most abundant in specific
kinds of freshwater hab.itat. Damsel flies and dragonflies (Qdonata) are
generally found in shallow lakes, but some species occur in running water.
Stoneflies (Plecoptera) and mayflies (Ephemeroptera) are predominantly
running water forms, although certain Ephemeroptera dwell in lakes and
ponds. Caddisflies (Tricoptera) abound in lakes and streams where the
water is well-aerated. The other groups also occur in both streams and
lakes (Edmondson, 1959).
Aquatic insects can be identified by using various keys (Pennak, 1978;
Edmondson, 1959; Needham and Needham, 1962; Merritt and Cummins, 1978).
Merritt and Cummins (1978) also provide lists of the species and habitats
(lentic or lotic) where they are most often found.
The species composition and number of individuals of the benthic community
change in response to increased organic and inorganic loading. Organic.
pollution generally causes a decrease in the number of species of
organisms, but an increase in the number of individuals. Inorganic pol-
lution, such as sediment, causes a decrease in the number of individuals,
as well as a decrease in species. The following sections focus on
qualitative and quantitative changes in freshwater benthic populations that
are indicative of types of pollution and of trophic state in lakes and
reservoirs.
Qualitative Response to Environmental Change
The most sensitive macroinvertebrate species are usually eliminated by
organic pollution. Because decay of organics often depletes oxygen, the
surviving species are those that are more tolerant of low dissolved oxygen
content. The predominant bottom conditions can be inferred by observing
which species are present at a specific site.
Suspended sediment and silt deposition may influence macroinvertebrates by
causing:
(a) Avoidance of adverse conditions by migration and drift;
111-14
-------
(b) Increased mortality due to physiological effects, burial, and
physical destruction;
(c) Reduced reproduction rates because of physiological effects, sub-
strate changes, loss of early life stages;
(d) Modified growth rates because of habitat modification and changes
in food type and availability (Farnworth, et al., 1979).
Indicator Organisms
The macroinvertebrate classes that are most often used as indicator organ-
isms are the Insecta and Annelida. These organisms are illustrated in
Figure III--1. Stonefly nymphs, mayfly naiads, and hellgrammites are
generally considered to be relatively sensitive to environmental changes.
The intermediately tolerant macroinvertebrates include scuds, sowbugs,
blackfly larvae, dragonfly nymphs, damselfly nymphs, and leeches. Blood-
worms (midge larvae) and sludgeworms make up the group of very tolerant
organisms.
Anaerobic environments are tolerated by sewage fly larvae and rat-tailed
maggots. Table II1-6 lists those aquatic insects that have been found at
dissolved oxygen concentrations of less than 4 ppm. The greatest number of
tolerant species are members of the order Diptera.
Sponges are affected by pollution although they are not usually considered
indicator organisms. Of the freshwater sponges, Ephydatia fluviatilis, E.
muelleri, Heteromeyenia tubisperma, and Eunaius fragilis~may be found Tn
eutrophic waters. Also. Ephydatia robusta can survive very low dissolved
oxygen levels and has been col I acted at DO tensions of 1.00 ppm (Harrison,
1974). Of the Mollusca, Unionid clams (Bivalvia) are considered sensitive
to environmental changes. Snails (Gastropoda) commonly occur in moderately
polluted environments. The most resistant species are Physa heterotropha,
P. integra, P_. gyrina, Gyranulus pan/us, Helisoma anceps, and jH. triyolvis.
o~ut almost every common species has been found in polluted areas (Harman,
1974).
Weber (1973) compiled a list of tolerances of freshwater macroinvertebrate
taxa to organic pollution (Appendix C). Organisms that occur in streams
and lakes are included. The tolerances of the organisms listed in the
appendix are based upon classification by various authors.
Trends in macroinvertebrate populations have been shown in studies of
eutrophic lakes. A collection of studies report the following responses of
macrofauna to increasing eutrophication:
o Oligochaetes, chironomids, gastropods and sphaerids increase and
Hexagenia (mayfly nymph) decreases (Carr and Hiltunen, 1965);
o Numbers of Oligochaetes relative to chironomids increase as organic
enrichment increases (Peterka, 1972);
111-15
-------
B
E.
N.
a.
K.
0.
A. Stonefly nymph (Plecoptera)
O U » , * £^ * * •» 4 * 4 *4 I P**C* MM A IMA *%^ A W*1«
«J UUI 1C i i Jr I ijnti p" \.' icwwpw^iu;
Mayfly naiad (Ephemeroptera)
C. Hellgrammite or Dobsonfly
larvae CCorydalidae)
Caddisfly larvae (Trichoptera)
Blackfly larvae (Simuliidae)
Scud (Amphipoda)
-. Aquatic sow bug (Isopoda)
H. Snail (Gastropoda)
J.
K.
L.
M.
N.
0.
P.
Fingernail clam (Sphaeriidae)
Damsel fly nymph (Zygoptera)
Dragonfly nymph (Anisoptera)
Bloodworm or midge fly larvflf1
(Tendipedidae)
Leech (Hirundinea)
Sludgeworm (Tubificidae)
Sewage fly larvae (Psychcda)
Rat-tailed maggot (Tubifera-
Eristalis)
Figure III-l. Representative bottom fauna (from Keup, et al., 1966).
ItI-16
-------
TABLE II1-6
SPECIES FOUND AT DISSOLVED OXYGEN LESS THAN 4 PPM
Odonata - dragonflies and damselflies
Ischnura posita (Hagen)
Pachydiplax longipenm's (Burm.)
Ephemeroptera - mayflies
Paraleptophlebia sp.
Caems sp.
Hemiptera - true bugs
Notonecta irrorata Uhl.
Tropisternus spp,
Machrpnychus glabratus
Stenelmis grossa Sand.
Say
Plea strioTa
Ranatra
Fieb.
austral Is Hung.
Ranatra kirkaldyf Bueno
Pelocoris femoratus P. de B.
Bel ostoma fluminea 'Say
Trepobates sp.
Rhagovel T& obesa Uhl.
Megaloptera - alderflies,
and fishflies
Chauliodes sp.
Coleoptera - beetles
Ha]1 pi us spp.
Peltodytes spp.
Coelambus spp.
LaccophlTus spp.
Hydroporus spp.
Dlneutes spp.
Gyrinus spp.
Lepldoptera - butterflies and moths
Parapoynx sp.
Tnchoptera - caddlsflies
Polycentropus remotus (Banks)
Oecetis eddlestonl Ross
Diptera - true flies
Procladius bell us (Loew)
Cljnotanypus ping'uis (Loew)
Ablabesmyla moni1 is (L.)
Tnchocladlus sp. Roback
Chironomus attenuatus (Walk.)
Chironomus ripariusTMei g.)
dobsonflies, CryptochTFonomus nr. fulvus (Joh.)
picrotendipes nervosus (Staeger)
Harm'schia nr."liboirtiva (Mall.)
MIcrotendTpes pedellus DeGeer
TrTbelos jucUndus (Walk.)
Rheotanytarsus exiguus (Jon.)
Calopsectra nr.~guer1a Roback
Palpomyla~gp. spp.
Tubifera tenax (L.)
SOURCE: Roback, 1974.
111-17
-------
o The smallest insect larvae are characteristic of oligotrophic
waters, and due to a shift in species composition, larval size
increases with increasing eutrophication (Jonasson, 1969);
o Tanytarsini are replaced by Chironomini in positions of dominance
with increasing eutrophication (Paterson and Fernando, 1970).
The study of four reservoirs (Salt Valley Reservoirs) in eastern Nebraska
revealed several trends in macrobenthic communities as eutrophication pro-
gressed. Contrary to the observation frequently reported that oligochaete
populations increase as eutrophication progresses, Hergenrader and Lessig
(198(Db) observed a decrease in Tubifex. They noted, however, that the deep
hypo!imnetic waters of the Salt Valley reservoirs do not become anaerobic,
as is the case in lakes where oligochaetes have increased. The Tanytarsini
(family Chironomidae) present in the less eutrophic reservoirs disappeared
in the most eutrophic. Finally, Sphaerium (order Mollusca) increased
during the early stages of eutrophication but declined as eutrophy pro-
gressed.
Chironomid Communities as Indicators
Instead of using a single organism to indicate water quality, Saether
(1979, 1980) suggests studying chironomid communities. By looking at
profundal, littoral and sublittoral chironomid communities, Saether was
able to delineate 15 characteristic communities found in environments
ranging from oligotrophic to eutrcphic. The communities, 6 in each of the
oligotrophic and eutrophic and 3 in the mesotrophic range, are lettered
from alpha to omikron. The Greek letters emphasize that the 15 sub-
divisions are not trophic level divisions, but are recognizable chironomid
communities. The species found in a lake or part of a lake can be used to
determine the associations and hence the extent of eutrophy. The key to
chironomid associations and the species list noted by Saether are presented
in Appendix 0. By using this system, Saether found significant
correlations between chironomid associations and the ratios of
chlorophyll-a to mean depth (Figure III-2) and total phosphorus to mean
depth (Figure II1-3).
Sediment Effects
The distribution of macroinvertebrates will be much less affected by
currents and drift in a lake than in a river. However, at those points
where rivers enter a lake, or where a river forms at the outlet from a
lake, one might expect to find macroinvertebrate populations that are
similar to the population of the connecting river. The distribution of
macroinvertebrates found in the 1 i-ttoral zone will be less affected by
drift (since rooted plants in the littoral tend to slow currents and
thereby inhibit drift) and more by the physical effects of suspended solids
and sedimentation. As concentrations of suspended and settleable solids
increase, invertebrates tend to release hold of the substrate to be
transported by currents or to migrate elsewhere. Migration from those
areas affected by sediment changes the structure of the benthic community.
The effects of suspended solids on benthic macroinvertebrates are
summarized in Table III-7.
111-18
-------
en
LU
CL
t—
lit
^
<
_l
6
V
M
-
*
v
c
©
h
™
C
-
£
-
6
L>«fJ
North Otoyo«» l.« •^^+ L. Haiartn
^S***\t L£ri«
• L. Hdloron {8lac>««Q
>X^ • L Shaha
>^
• Ctnfra^c Erit
•/I. SammamisJi
Ea«( L Eno
j»l Malartn (Pratttjanitnl
I Qntar*
• / «L Washington
J.M»fttrn
•/•Okanagan L
jKfclaaalka L.
yUL Huron
MjL.Hjota
• Wood
iar«n •• South Oioyooi L
1 > M6S In i • 10 ail
r a 0952 IP < 00011
L. Sup«r.or
0.5
15 20 2.5
Chlorophyll a / z /ig /1 / m
30
Figure III-2. Chlorophyll-a/ Mean Lake Depth in relation to 15 lake
types based on Chironomid Communities (From Saether, 1979)
111-19
-------
CO
UJ
Q.
>-
LU
• Sour* Oioyoot I
L. Malortn (S
North OtOT«ok !..• ^ W«»f L tfit•
^
L Malartn I Btacktnt
«*L Malartn
• C«nt/al I £r»
r LCsnttonc* 119*31
• L innartn
L Hatart
-------
TABLE III-7
SUMMARY OF SUSPENDED SOLIDS EFFECTS ON AQUATIC MACROINVERTEBRATES
I
f\J
OrgarUtm(i)
Mixed Population*
Mixed Population*
Mixed Population*
Mixed Population*
Chrionomui A.
Tubificidae
Cheumaioptyche
(Net spinners)
Tricorylhoide*
Mixed Population*
Mixed Populations
Chiionomidae
Ephemoplcra,
Siinuliidae,
llydracarina
Effect
Lower lummei
populationi
Reduced popula-
liont lo 25%
Densities 11%
of normal
No organisms in the
zone of fettling
Normal fauna re-
placed by
(Speciet Selection)
Number reduced
Number increased
90% increase in
drift
Reduction in
number*
Increased drift with
suspended sediment
Inconsistent drift
response lo added
sediment
Suspended Solid
Concentration
26l-390ppm
(Turbidity)
1000-6000 ppm
>SOOO ppm
(High concen-
trations)
80mg/l
40-200 JTU
Source of
Suspended Solid*
Mining area
Log dragging
Glass manufacturing
Colliery
Limestone Quwry
Limestone Quarry
Limestone Quarry
Manganese
Strip mine
Experimental sediment
addition
Experimental sediment
addition
Comment
Normal populations at
60 ppm
Effect noted 1 3 miles
downstream
Reduction in light re-
duced submerged plants
Suspended solid* as high
a*250mg/l
Due to preference for .
mud or rill
Also caused changes in
density and diversity
SOURCE: Sorenson, et al., 1977.
-------
Deposition of sediment in the profundal zone may provide a stable sub-
strate. In contrast deltas where streams enter the lake or reservoir may
be subject to continuing deposition and erosion. Such areas will support
fewer species and fewer numbers of organisms than the more stable profundal
zone.
•f
Sediment deposition modifies macroinvertebrate habitat and alters the type,
distribution and availability of food. Substrate preference of macro-
invertebrates is related to a variety of factors. In addition to particle
size, the colonization of an area is dependent on the amount and type of
detritus, the presence of vegetation, the degree of compaction and the
amount of periphyton (Farnworth et al., 1979). Sediment preferences may
change with an organism's life history stage, thus compounding the problem
of categorizing associated substrate. Nonetheless, certain groups such as
Chironomidae and Tricorythodes, are recognized as preferring fine sediment.
Quantitative Response to Environmental Change
Quantitative techniques that are used to assess the biological integrity of
lakes include a number of mathematical indices, or focus on the abundance
of certain benthic organisms. These methods are summarized in the fol-
lowing sections. Other measures of community health, such as diversity
indices, are discussed in the Technical Support Manual: Water body Surveys
and Assessments for Conducting Use Attainabl11ty Analyses (U.S. EPA,
1983]>), and in a review by Washington (1984).
Oligochaete Populations
Oligochaetes, particularly members of the family Tubificidae, are present
in large numbers in polluted areas. Aston (1973) found that Limnodrilus
hoffmeisteri and Tubifex tubifex predominate in areas receiving heavy
sewage pollution. In a review of the relationship between tubificids and
water quality, Aston (1973) noted several investigations that have used the
population density of tubificids as an index of pollution. Surber (cited
by Aston, 1973), studied a number of lakes in Michigan and concluded that
areas with an oligochaete density of more than 1,100 per square meter were
truly polluted. Carr and Hiltunen (1965) used the following numbers of
oligochaetes per square meter to indicate pollution in western Lake Erie:
light pollution, 100 to 999; moderate pollution, 1,000 to 5,000; and heavy
pollution, more than 5,000. This means of classification fails to consider
seasonal variation in population density and the organic content and
particle size of the bottom substrate. Since the population density is
likely to vary, this method has limited utility (Aston, 1973).
Wiederholm (1980) noted that a simple depth adjustment could make oligo-
chaete abundance more applicable. By dividing the number of oligochaetes
per square meter by the sampling depth, he found that the correlation with
chlorophyll was increased. This adjustment may account for factors that
are affected by depth such as food supply, predation pressure (which
declines as depth increases), and possible oxygen^deficits.
The relative abundance of oligochaetes may be'-a better indication of
organic pollution than the population density. In a stream study, Good-
night and Whitley (1961) suggested that a population of 80 percent or more
111-22
-------
of oligochaetes in the total macroinvertebrate population indicates a high
degree of organic enrichment. They hypothesized that percentages from 60
to 80 indicate doubtful conditions and below 60 percent, the area is in
good condition. Howmlller and Beeton (1971) used this index in a study of
Green Bay, Lake Michigan, and concluded that in 1967 the lower bay was in a
highly polluted state, and the middle bay had "doubtful conditions."
Brinkhurst (1967) suggested that the relative abundance of the tubifield
Limnodrilus hoffmeisteri (as a percentage of all oligochaetes) may be a
useful measure of organic pollution. Increased percentages of L_. hoff-
meisteri are often indicative of organic pollution. Lower Green Bay (73%
L. hottmeisteri) was identified as being more polluted than middle Green
Bay (50% and 42% L. hoffmeisteri) by reference to the relative abundance of
this oligochaete THowmiller and Scott, 1977).
Oligochaete/Chironomid Ratio
Another proposed indicator uses the ratio of oligochaetes to chironomids.
Generally, the ratio increases as the lake becomes more eutrophic.
Wiederholm (1980) advocates including a depth adjustment (ratfo divided by
sampling depth) when using the oligochaete/chironomid ratio since oligo-
chaetes tend to increase in dominance at greater depths. Studies of
Swedish lakes showed a high correlation between depth-adjusted oligochaete/
chironomid ratios and trophic state, but very little correlation of the
non-adjusted ratio with trophic state. Table III-8 shows that the depth-
adjusted oligochaete/chironomid ratio had low values (from 0-1.5) in
oligotrophic lakes, and progressively higher values for mesotrpphic
(1.5-3.0), eutrophic (3.0-7.4) and hypereutrophic (>18) lakes. Wiederholra
suggests that the oligochaete/chironomid ratio may be used directly when
comparing data from a single site over time or different lakes over time,
but a general application needs some adjustment for depth.
Mathematical Indices
A survey of the literature reveals at least four mathematical indices in
addition to diversity indices that may be applicable in freshwater lake
studies. These indices are described in Table III-9.
Based on their studies of rivers and streams receiving sewage, Kolkwitz and
Marsson (1908, 1909) proposed their sapropic system of zones of organic
enrichment. They suggested that a river receiving a load of organic matter
would purify itself and that it could be divided into saprobic zones
downstream from the outfall, each zone having characteristic biota.
Kolkwitz and Marsson published long lists of the species of plants and
animals that one could expect to be associated with each zone. The zones
were defined as follows:
o Polysaprobic; gross pollution with organic matter of high molecu-
lar weight, very little or no dissolved oxygen and the formation of
sulphides. Bacteria are abundant, and few species of organisms are
present.
111-23
-------
TABLE III-8
BENTHIC COMMUNITY MEASURE
WITH AND WITHOUT ADJUSTMENT FOR DEPTH
Lake
Approximate
Trophic
State3
Ch1orophyl1-a
(ug/1)b ~
OHgochaete/
Chlronomid Ratio
(*)
Vattern, 20-40m
Vattern, 90-110m
Yanern, 40-80 m
Skaren, 10-26m
Innaren, 14-19m
Sommen, 16-49m
Malaren, area C, 30m
Malaren, area C, 45-50m
Malaren, area B, 15m
Hjalmaren, area C, 6-18m
S. Bergundasjon, 3-5m
Yaxjosjon, 3-5m
Hjalmaren, area B, 2-3m
0
0
0
0
M
M
M
M
E
E
HE
HE
HE
1.1
1.1
1.7
2-2.5
2.5-3
3-4
5.5
5.5
17.5
9.4
25-75
50-100
102
wi thout
depth adj.
38.9
90.1
86.0
25.9
. 19.8
44.3
85.5
96.4
69.0
71.9
69.0
87.4
66.8
with
depth adj.c
1.3'
0.9
1.5
1.5
1.2
1.9
2.9
2.0
4.6
7.4
18.5
21.6
34.4
a. 0 3 ollgotrophic, M a mesotrophlc, E 3 eutrophlc, HE = hypereutrophic
b. May-October, 1m
c. 011gochaete/Ch1ronomid ratio divided by sampling depth
SOURCE: Wlederholm, 1980.
111-24
-------
TABLE II1-9
MATHEMATICAL INDICES
Index Name and Description
Saprobic Index
S
Reference
Saether, 1979
2s-h
2 h
s 3 1-4, Oligo - to polysaprobic
h 3 occurrence value; 1, occasional
3, common; 5, mass occurrence.
Benthlc Quality Index
Wlederholm, 1976
Wlederholm, 1980
BQI =
Jo
N
j = based on indicator species of
chironomids, see text
^ = number of individuals of the various groups
N = the total number of indicator species
BQI
C
Saether, 1979
1
T
the constancy of the respective groups
within a sample
Trophic Condition Index
TCI = 2Ni*22N2
2NQ +2Ni +ZN2
2NQ a total number of oligochaete worms
Intolerant of eutrophic conditions
(see Table C)
2N. = total number of organisms characteristics
of mesotrophic areas
2N2 = total number belonging to species
tolerant of extreme eutrophy
Howmlller and Scott, 1977
Saether, 1979
111-25
-------
o Mesosaprobic: simpler organic molecules and increased DO content.
Upper zone (alpha-mesosaprobic) has many bacteria and often fungi,
with more types of animals and lower algae. Lower zone
(beta-mesosaprobic) has conditions suitable for many algae,
tolerant animals and some rooted plants.
o 01 igosaprobic: oxygen content is back to normal and a wide range
of plants and animals occur.
As stated, the saprobic system was designed for rivers and streams.
Nevertheless, the concept could be applied to riverine impoundments that
have a predominant longitudinal flow. More importantly, however, is the
impetus generated by the saprobic 'system for the development of subsequent
biological indices.
Pantle and Buck (1955, cited by Saether, 1979) applied the ideas of Kolk-
witz and Marsson in the Saprobic Index (Table III-9), which was proposed
for use in stream studies. Further extensions of the saprobic system were
made by Sladecek (1965) and these modifications are summarized in Nemerow
(1974).
Wiederholm proposed the Benthic Quality Index (BQI) in 1976 for studies of
Swedish Lakes (cited by Saether, 1979). The value of k.- (Table 111-9)
represents the empirical position of each species in the range from oligo-
trophic to eutrophic conditions. The indicator species used by Wiederholm
were given the following values for k.: 5, Heterotrissocladius subpilosus
(Kieff.); 4, Micropsectra spp. and Paraclaaopeima sppTJspecifically K.
m'gritula (Goetgh.); 3. Phaenpspectra coracina (Zett.) and Stictochironomus
rpsenschoeldi (Zett.); 2, Chlronomus anthracinus (Zett.)l1, Chironomus
plumosus L.; 0, absence of these indicator species. The BQI was related to
total phosphorus/mean lake depth as shown in Figure III-4. The value of
the index approaches 0 as the lakes become more eutrophic, and is nearly 5
in oligotrophic lakes. With the indicator species used here, the BQI
applies to Palearctic lakes (e.g., Europe, Asia north of the Himalayas,
Northern Arabia, Africa north of the Sahara). However, the species used as
indicators may be redefined for Nearctic lake studies (e.g.," lakes in
Greenland, arctic America, northern and mountainous parts of North America)
by using the species lists given in Appendix D.
The Trophic Condition Index (TCI) is the only commonly used index that was
developed in North America specifically for lake studies. This index
(Table III-9) was designed by Brinkhurst (1967) for use on Great Lakes
waters. It is based on oligochaetes which are classified according to the
degree of enrichment of the environments where they are typically found
(Table 111-10). The TCI ranges from 0 to 2, with the higher values associ-
ated with more eutrophic conditions.
In a study of Green Bay, Howmiller and Scott (1977) compared'the TCI with
four other indices. Only the Trophic Condition Index showed a significant
difference between the three areas of Green Bay shown in Figure III-5. The
other indices used were Species Diversity, Oligochaete worms per square
meter, Oligochaete worms (%) and L^ hoffmesiteri (1). As shown in Table
III-11, these indices show no statistical difference between Areas II and
III, and sometimes no significant difference from values for Area I.
111-26
-------
• V«tt«rn
• Vin«rn
\
, Somm«n
• M«l«r«n ( 78 )
\» Innarvn
Figure III-4. Total phosphorus/mean lake depth in
relation to a benthic quality index (BQI) based
on indicator species of chironomids (From Wiederholm, 1980)
111-27 .
-------
TABLE I11-10
A CLASSIFICATION OF OLIGOCHAETE SPECIES
ACCORDING TO THE DEGREE OF ENRICHMENT OF THE ENVIRONMENTS
IN WHICH THEY ARE CHARACTERISTICALLY FOUND
Group 0
Species largely restricted to oligotrophic situations:
Stylodrilus herfngianus
Peloscolex variegatus
P. superiorensis
Limnodrllus profundicola
Tubifex kessleri
Rhyacodrilus coccineus
R. montana
Group 1
Species characteristic of areas which are mestrophic or only slightly
enriched:
Peloscolex ferox
P. freyi
IlyodrHus tempietoni
Potamothrix moldaviensls
P. vejdovskyi
Aulodrilus spp.
Arcteonais lomondi
Dero digitata
Mais elinguis
Slavina appendiculata
Uncinais uncinata
Group 2
Species tolerating extreme enrichment or organic pollution:
Limnodrilus anguistipenis
L. cervix
L. claparedeianus
L. hoffmeisteri
L. maumeensis
L. udekemianus
Peloscolex multisetosus
Tubifex tubifex
SOURCE: Hownlller and Scott, 1977,
111-28
-------
Figure III-5.
Map of Lower and Middle Green Bay showing
location of benthos sampling stations and
areas designated I, II, and III (from Howmiller
and Scott, 1977).
II1-29
-------
TABLE III-ll
AVERAGE VALUES OF FIVE INDICES OF POLLUTION
COMPARED FOR THREE AREAS OF GREEN BAY
Species Diversity
01 i gochaete worms/nr
Oligochaete worms, %
L. hoffmeisteri, %
Trophic index
I
l.CO
1085
63
73
1.92
Area
II
1.62
1672
53
50
1.84
III
1.66
1152
53
42
1.53
NOTE: Values underscored with a common line are not
significantly different from each other.
SOURCE: Howmiller and Scott, 1977.
111-30
-------
FISH
Although fish species in many instances show no preference for either
lacustrine or riverine habitat, certain environmental components (e.g.,
velocity, substrate, dissolved oxygen and temperature) render one habitat
more suitable than another. The following paragraphs highlight the habitat
requirements of certain fish species that are predominantly lacustrine.
Trophic State Effects
Oligotrophic and eutrophic lakes have characteristic fish populations
because of their contrasting habitats. Briefly, oligotrophic lakes are
generally deep and often large in size, and are located in regions where
the substratum is rocky. These lakes usually stratify in summer, but the
cool profundal zone contains sufficient oxygen year-round for fish sur-
vival. Oligotrophic lakes support less than 20 pounds of fish per surface
acre, and characteristic fish are salmons, trouts, chars, ciscoes, and
graylings (Bennett, 1971).
Eutrophic lakes support fish populations of largemouth bass, white bass,
white and black crappies, bluegill and other sunfish, buffalo, channel
catfish, bullheads, carp, and suckers (Bennett, 1971). Such lakes have
shallow to intermediate depths, may have large or small surface areas, and
are located in regions with more fertile soil than oligotrophic lakes.
Hypolimnetic waters of eutrophic lakes frequently exhibit reduced oxygen
levels during summer stratification.
Nutrient enrichment which causes increased production in lakes accelerates
the natural progression of trophic state from oligotrophy to eutrophy.
Initially, eutrophication and the subsequent abundance of food organisms
may cause increased growth of fish. However, undesirable conditions of
temperature and dissolved oxygen in later stages force some fish to leave
the affected area or perish. Fish commonly respond to changes associated
with eutrophication by shifting their horizontal and vertical distribution.
In Lake Erie, whitefish and ciscoes became restricted to the eastern basin
as the environment became more unsuitable (Beeton, 1969). Perch and
whitefish may move from the littoral zone into the pelagic zone, where they
are not usually found (Larkin and Northcote, 1969). The restriction of
coldwater fishes to a thin layer between the oxygen deficient hypolimnion
and the warm epilimnion may lead to mortalities. This may have contributed
to the disappearance of ciscoes from Lake Mendota, Wisconsin.
As eutrophication proceeds, there is a general pattern of change in fish
populations from coregonines to coarse fish. One of the best examples of
population changes is in the Great Lakes. Although factors other than
eutrophication may have contributed to the loss of some species, enrichment
is recognized as being an important cause. Commercial fisheries provide
information on the species composition of catches. In Lake Erie, the major
species in the 1899 catch were lake herring (cisco), blue pike, carp,.
yellow perch, sauger, whitefish and walleye. By 1940, the lake herring and
sauger fisheries had collapsed, and the catch was dominated by blue pike,
whitefish, yellow perch, walleye, sheepshead, carp, and suckers. Blue pike
and whitefish populations have since declined, and the catch has become
111-31
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more concentrated on the warmwater species such as freshwater drum, carp,
yellow perch and smelt (Beeton, 1969; Larkin and Northcote, 1969).
Temperature Effects
Temperature as well as trophic state plays a role in determining the fish
species inhabiting a lake. Trout are generally considered representative
of coldwater species. Rainbow trout and brook trout thrive in water with a
maximum summer temperature1 of about 70°F. Rainbow trout are more tolerant
of higher temperatures than brook trout. Prolonged exposure to tempera-
tures of 77.5°F is lethal to brook trout (Bennett, 1971).
Fish typical of wanner waters include largemouth bass, bluegill, black and
white crappie, and black :and yellow bullhead. These species are fairly
tolerant of high, naturally occurring, water temperatures, and generally
suffer mortality only when additional adverse factors (e.g., anoxic
conditions, toxics, thermal plumes) prevail. Species such as smallmouth
bass, rock bass, walleye, northern pike, and muskellunge are more sensitive
to increased temperatures than the more typical warmwater fish, but are not
as sensitive as trout.
Warmwater fish and coldwater fish may live in the same lake. For example,
a two-tier fishery may exist in a stratified lake, wherein warmwater fish
live in the epilimnion and the metalimnion, while coldwater fish survive in
the cooler waters of the hypolimnion.
Specific Habitat Requirements
Specific habitat requirements for some lake species are published in a
series of documents (Habitat Suitability Index Models) prepared by the Fish
and Wildlife Service and available through the National Technical
Information Service. These publications summarize habitat suitability
information for many lake species including: rainbow trout, longnose
sucker, smallmouth buffalo, bigmouth buffalo, black bullhead, largemouth
bass, yellow perch, green sunfish, and common carp. The following
information on the habitat requirements of these species is contained
within the Fish and Wildlife Service reports.
Rainbow Trout
Rainbow trout prefer cold, deep lakes that are usually oligotrophic. The
size and chemical quality of the lakes may vary. Rainbow trout require
streams with gravel substrate in riffle areas for reproduction. Spawning
takes place in an inlet or outlet stream, and those lakes with no tributary
streams generally do not support reproducing populations of rainbow trout.
The optimal water velocity for rainbow trout redds is between 30 and 70
cm/sec. Juvenile lake rainbow trout migrate from natal streams to a
freshwater lake rearing area.
Adult lake rainbow trout prefer temperatures less than 18°C, and generally
remain at depths below the 18°C isotherm. They require dissolved oxygen
levels greater than 3 mg/1 (Raleigh, et al., 1984).
111-32
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Longnose Sucker
This species is most abundant in cold, oligotrophic lakes that are 34-40 m
deep. These lakes generally have very little littoral area. They are also
capable of inhabiting swift-flowing streams, but longnose suckers in lake
environments enter streams and rivers only to spawn or to overwinter. The
longnose sucker spawns in riffle areas (velocity 0.3-1.0 m/sec), where the
adhesive eggs are broadcast over clean gravel and rocks (Edwards, 1983a_).
Smallmouth Buffalo
Although smallmouth buffalo typically inhabit large rivers, preferring
deep, clear, warm waters with a current, they can do well in large reser-
voirs or lakes. Lake or reservoir populations spawn in embayments or along
recently flooded shorelines. Although small mouth buffalo will spawn over
all bottom types, they prefer to spawn over vegetation -and submerged ob-
jects. Juveniles frequent warm, shallow, vegetated areas with velocities
less than 20 cm/sec. Adults are found in areas with velocities up to 100
cm/sec (Edwards and Twomey, 1982a).
Bigmouth Buffalo
Bigmouth buffalo prefer low velocity areas (0-70 on/sec), and inhabit large
rivers, lowland lakes and oxbows, and reservoirs. Populations in reser-
voirs reside in warm, shallow, protected embayments during, the summer, and
move into deeper water in the fall and winter. Fluctuations of reservoir
water levels reduce buffalo populations due to siltation, erosion and loss
of vegetation (Edwards, 1983J)).
Black Bullhead
Bullheads live in both riverine and lacustrine environments. Optimal
lacustrine habitat has an extensive littoral area (more than 25 percent of
the surface area), with moderate to abundant (more than 20 percent) cover
within this area. Bullhead nests are located in weedy areas at depths of
0.5-1.5 m. Black bullheads are most common in areas of low velocity (less
than 4 cm/sec). They prefer intermediate levels of turbidity (25-100 ppm),
and can withstand low dissolved oxygen levels (as low as 0.2-0.3 mg/1 in
winter, 3.0 mg/1 in summer) (Stuber, 1982).
Largemouth Bass
Largemouth bass prefer lacustrine environments. Optimal habitats are lakes
with extensive shallow areas (more than 25 percent of the surface area less
than 6 m depth) for growth of submergent vegetation, but deep enough (3-15
m) to successfully overwinter bass. Current velocities below 6 cm/sec are
optimal, and velocities above 20 cm/sec are unsuitable. Temperatures from
24-30°C are optimal for growth of adult bass. Largemouth bass will nest on
a variety of substrates, including vegetation, roots, sand, mud, and cob-
ble, but they prefer to spawn on a gravel substrate. Adult bass are con-
sidered intolerant of suspended solids; growth and survival of bass is
greatest in low turbidity waters (less than 25 ppm suspended solids). Bass
show signs of stress at oxygen levels of 5 mg/1, and DO concentrations less
than 1.0 mg are lethal (Stuber, et al., 1982aO.
111-33
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Yellow Perch
Yellow perch prefer areas with sluggish currents or slack water. They
frequent littoral areas in lakes and reservoirs, where there are moderate
amounts of vegetation present. Riverine habitat resembles lacustrine
areas, with pools and slack-water. Perch spawn in depths of 1.0 m to 3.7
m, and in waters of low (less than 5 cm/sec) current velocity. Littoral
areas of lakes and reservoirs provide both spawning habitat and cover
(Krieger, et al., 1983).
Green Sunfish
Green sunfish thrive in both riverine and lacustrine environments. Optimal
lacustrine environments are fertile lakes, ponds, and reservoirs with
extensive littoral areas (more than 25 percent of the surface area).
Preferred environmental parameters are: velocities less than 10 cm/sec,
moderate turbidities (25-100 JTU) and DO levels of more than 5 mg/1 (lethal
levels of 1.5 mg/1) (Stuber, et al., 1982£).
Common Carp
This species prefers areas of slow current. In both riverine and lacus-
trine environments, carp prefer enriched, relatively shallow, warm, slug-
gish and well-vegetated waters with a mud or silty substrate. Adults are
generally found 1n association with abundant vegetation. The common carp
is extremely tolerant of turbidity and its own feeding and spawning
activities over silty bottoms increase turbidity. Adults are also tolerant
of low dissolved oxygen levels, and can gulp surface air when the dissolved
oxygen is less than 0.5 mg/1 (Edwards and Twomey, 1982bJ.
Stocking
The most common fish management technique used is stocking. The purpose of
stocking is to improve the fish population, and certain fish are used more
often than others. The following description is based on information in
Bennett (1971).
Bass and bluegills have often been stocked in the same pond or lake. The
theory behind stocking these species in combination is that both largemouth
bass and bluegills would be available for sport-fishing. The role of the
bluegills is to convert invertebrates into bluegill flesh. The bass then
feed on small bluegills and thereby control the population. Problems may
be caused from an overpopulation of one species, especially since the
bluegills overpopulate more often than the bass. Stocking ratios (numbers
of bass : numbers of bluegills) as discussed by Bennett (1971), influence
the outcome of such stocking endeavors.
Because largemouth, smallmouth, and spotted bass are omnivorous, any of
these three species stocked alone may be fairly successful. They feed on
crayfish, large aquatic insects and their own young. These species do well
in warmwater ponds if they do not have to compete with prolific species
such as bluegills, green sunfish, and black bullheads. Largemouth bass
111-34
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have been stocked in warmwater ponds in combination with minnows, chub-
suckers, red-ear sunfish or war-mouths. These combinations have proved to
be successful.
Walleye stocking reportedly has variable success except in waters devoid of
other fishes. In waters such as new reservoirs and renovated lakes, satis-
factory survival rates for walleye occur. Bennett (1971) noted that,
generally, walleye stocking was unsuccessful in acid or softwater lakes.
111-35
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CHAPTER IV
SYNTHESIS AND INTERPRETATION
INTRODUCTION
The basic physical and chemical processes of the lake were introduced in
Chapter II. Chapter II also includes a discussion of desktop procedures
that might be used to characterize various lake properties, and a dis-
cussion of mathematical models that are suitable for the investigation of
various physical and chemical processes.
The applicability of desktop analyses or mathematical models will depend
upon the level of sophistication desired for a use attainability study.
Case studies were presented to illustrate the use of measured data and
model projections in the use attainability study. The selection of a
reference site is discussed later in Chapter IV.
Chapter II also provides a discussion of chemical phenomena that are of
importance in lake systems. Most important of these are the processes that
control internal phosphorus cycling, and the processes that control dis-
solved oxygen levels in the epilimnion and the hypolimnion of a stratified
lake. Chemical evaluations are also discussed in the earlier Technical
Support Manuals (U.S. EPA, 1983J), 1984).
The biological characteristics of the lake are summarized in Chapter III.
Specific information on plant, fish and macroinvertebrate lake species is
presented to assist the investigator in determining aquatic life uses.
The emphasis in Chapter IV is placed on a synthesis of the physical, chemi-
cal and biological evaluations which will be performed to permit an overall
assessment of aquatic life protection uses in the lake. A large portion of
this discussion is devoted to lake restoration considerations.
Like the two previous Technical Support Manuals (U.S. EPA, 1983_b, 1984),
the purpose of this Manual is not to specifically describe how to conduct a
use attainability analysis. Rather, it is the desire of EPA to allow the
states some latitude in such assessments. This Manual provides technical
support by describing a number of physical, chemical, and biological
evaluations, as well as background information, from which a state may
select assessment tools to be used in a particular use attainability
analysis.
USE CLASSIFICATIONS
There are many use classifications—navigation, recreation, water supply,
the protection of aquatic life—which might be assigned to a water body.
These need not be mutually exclusive. The water body survey as discussed
in this volume is concerned only with aquatic life uses and the protection
of aquatic life in a lake.
IY-1
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The objectives in conducting a use attainability survey are to identify:
1. The aquatic life use currently being achieved in the water body;
2. The potential uses that can be attained, based on the physical,
chemical and biological characteristics of the water body; and
3. The causes of any impairment of uses.
The types of analyses that might be employed to address these three points
are listed in Table IV-1. Most of these are discussed in detail in this
volume, and in the two preceding volumes on estuaries and on rivers and
streams.
Use classification systems vary widely from state to state. Use classes
may be based on salinity, recreation, navigation, water supply (municipal,
agricultural, or industrial), or aquatic life. In some cases geography
serves as the basis for use classifications. Aquatic life use classi-
fications found in state standards generally are rather broad (e.g.,
coldwater fishery, warmwater fishery, fish maintenance, protection of
aquatic life, etc.) and offer little specificity. Clearly, little
information is required to place a water body into such broad categories.
Far more information may be gathered in a water body survey than is needed
simply to assign a classification that is drawn from available state clas-
sifications. The additional data that is gathered is required, neverthe-
less, in order to evaluate management alternatives for the lake and, if
appropriate, to refine state use classification systems for the protection
of aquatic life.
In general, state water quality standards do not address lakes specif-
ically, so one must assume that standards written to cover surface waters
in some states, or rivers and streams in others, are intended to stand for
lakes as well. From the standpoint of aquatic life protection uses this
may be satisfactory since the types of fish found in lakes are also found
in the streams that discharge into lakes. However, the fact that some
lakes stratify and others do not suggests that seasonal aquatic life uses
in a lake could be more complex than in adjacent streams. In highly
stratified lakes, for example, the fish population of the epilimnion might
be substantially different from that of the hypolimnion. That a shallow
lake may become anoxic during summer stratification may have important
implications for the uses of the hypolimnion. That the epilimnion may
become anoxic because of diurnal DO fluctuations due to massive algal
blooms and decay also has implications for the definition of present and
future uses.
Since there may not be an adequate spectrum of aquatic protection use cate-
gories available against which to compare the findings of the biological
survey; and since the objective of the survey is to compare existing uses
with designated uses, and existing uses with potential uses, as seen in the
three points listed above; the investigators may need to develop their own
system of ranking the biological health of a water body (whether quali-
tative or quantitative) in order to satisfy the intent of the water body
survey. Implicit to the use attainability survey is the development of
IV-2
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TABLE IV-1
SUMMARY OF TYPICAL WATER BODY EVALUATIONS
PHYSICAL EVALUATIONS
CHEMICAL EVALUATIONS
BIOLOGICAL EVALUATIONS
o Size (mean width/depth)
o Flow/velocity
o Total volume .
o Reaeration rates
o Temperature
o Suspended solids
o Sedimentation
o Bottom stability
o Substrate composi-
tion and character-
istics
o Sludge/sediment
o Riparian character-
istics
o Downstream
characteristics
o Dissolved oxygen,
o Nutrients
- nitrogen
- phosphorus
o Chlorophyll-a
o Sediment oxygen demand
o Salinity
o Hardness
o Alkalinity
o pH
o Dissolved solids
o Toxics
o Biological inventory
(existing use analysis)
o Fish
o Macroinvertebrates
o Microinvertebrates
o Plants
- phytopi ankton
- macrophytes
o Biological condition/
health analysis
- diversity indices
- primary productivity
- tissue analyses
- Recovery Index
o Biological potentia.1
analysis
o Reference reach
compari son
SOURCE: Adapted from EPA 1982a_, Water Quality Standards Handbook
IV-3
-------
management strategies or alternatives which might result in enhancement of
the biological health of the water body. A clear definition of uses is
necessary to weigh the predicted results of one strategy against another in
cases where the strategies are defined in terms of protection of aquatic
life.
Since one may very well be seeking to define use levels within an existing
use category, rather than describe a shift from one use classification to
another, the existing state use classifications may not be helpful. There-
fore, it may be necessary to develop an internal use classification system
to serve as a yardstick during the course of the water body survey, which
may later be referenced to the legally constituted use categories of the
state.
A scale of biological health classes is presented in Table IV-2 that offers
general categories against which to assess the biology of a lake. A
descriptive scale is found in Table IV-3 that may be used to assess a water
body. This scale was developed by EPA in conjunction with the National
Fisheries Survey.
REFERENCE SITES
Selection
Chapter IV-6 of the Technical Support Manual (U.S. EPA, 1983J>) presents a
detailed discussion on the concept of ecological regions and the selection
of regional reference sites. This process is particularly applicable to
small and medium size lakes. .Use attainability studies for very large
lakes are more likely to be concerned with specific segments of the lake
than with the lake in its entirety. Resource requirements are an important
consideration as well for very large lakes. For example, New York State
may be prepared to investigate uses in Lake Ontario near Buffalo, but may
not be prepared to study the entire lake. A study of this magnitude could
not be done without federal participation, or in the case of Lake Ontario
or Lake Erie, international participation. For the scale of study that a
state may embark upon, reference sites could well be segments of the same
or other large lakes.
The concept of developing ecological regions that are relatively homo-
geneous can be applied to lakes. This concept is based on the assumption
that similar ecosystems occur in definable geographic patterns. Although
the biota of particular lakes in close proximity may vary, it is more
likely to be similar in a given region than in geographically dissimilar
regions.
Within each region various lakes are investigated to determine which sites
have a well balanced ecosystem and to note watershed land use and land
cover characteristics and the effects of man's activities. A major
characteristic to look for in the selection of a reference lake is the
level of disturbance in the watershed that feeds the lake. Good reference
site candidates are lakes located away from heavily populated areas, such
as in protected park land.
IV-4
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TABLE IY-2
BIOLOGICAL HEALTH CLASSES WHICH COULD BE USED
IN WATER BODY ASSESSMENT
Class
Attributes
Excellent
Good
Fair
Poor
Very Poor
Extremely Poor
Comparable to the best situations unaltered by man; all
regionally expected species for the habitat including the
most intolerant forms, are present with full array of age
and sex classes;, balanced trophic structure.
Fish invertebrate and macroinvertebrate species richness
somewhat less than the best expected situation; some
species with less than optimal abundances or size dis-
tribution; trophic structure shows some signs of stress.
Fewer intolerant forms
are present.
of plants, fish and invertebrates
Growth rates and condition factors commonly depressed;
diseased fish may be present. Tolerant macroinvertebrates
are often abundant.
Few fish present, disease, parasites, fin damage, and other
anomalies regular. Only tolerant forms of macroinverte-
brates are present.
No fish,
life.
very tolerant macroinvertebrates, or no aquatic
SOURCE: Modified from Karr, 1981
IV-5
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TABLE IV-3
AQUATIC LIFE SURVEY RATING SYSTEM
A water body that is rated a five has:
- A fish community that is well balanced among the different levels of the food
chain.
- An age structure for most species that is stable, neither progressive (leading to
an increase in population) or regressive (leading to a decrease in population).
- A sensitive sport fish species or species of special concern always present.
- Habitat which will support all fish species at every stage of their life cycle.
- Individuals that are reaching their potential for growth.
- Fewer individuals of each species.
- All available niches filled.
A water body that is rated a four has;
- Many of the above characteristics but some of them are not exhibited to the full
potential. For example, the water body has a well balanced fish community; the age
structure is good; sensitive species are present; but the fish are not up to their
full growth potential and may be present in higher numbers; an aspect of the
habitat is less than perfect (i.e., occasional high temperatures that do not have
an acute effect on the fish); and not all food organisms are available or they are
available in fewer numbers.
A water body that is a three has:
- A community is not well balanced, one or two trophic levels dominate.
- The age structure for many species is not stable, exhibiting regressive or
progressive characteristics.
- Total number of fish is high, but individuals are small.
- A sensitive species may be present, but is not flourishing.
- Other less sensitive species make up the majority of the biomass.
- Anadromous sport fish infrequently use these waters as a migration route.
A water body that is rated a two has:
- Few sensitive sport fish are present, nonsport fish species are more common theu>
sport fish species.
- Species are more common than abundant.
- Age structures may be very unstable for any species.
- The composition of the fish population and dominant species is very changeable.
- Anadromous fish rarely use these waters as a migration route.
- A small percent of the reach provides sport fish habitat.
A water body that is a one has:
- The ability to support only nonsport fish. An occasional sport fish may be found
as a transient.
A water body that is rated a zero has:
- No ability to support a fish of any sort, an occasional fish may be found as &
transient.
IY-6
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For the selection of a reference lake, it is important to seek compara-
bility in physical parameters such as surface area, volume, and mean depth,
and in physical processes such as degree of stratification and sedimenta-
tion characteristics.,_j*It will be important also to seek comparability in
detention time, which plays a role in determining the chemical and
biological characteristics of the lake. Detention time is determined by
lake volume and rate of flow into the lake from both point and nonpoint
sources.
The selection of a candidate reference lake could be based on an analysis
of existing data. Data for many lakes throughout the country are available
from the National Eutrophication Survey conducted by the U.S. EPA in
cooperation with state and local agencies. National computerized data
bases such as WATSTORE and STORE! can provide flow and water quality data.
Many states and counties have their own water qu'ality and biological
monitoring programs which should be used to obtain the most up-to-date
information on the lake.
In addition to the historical data that may be available through WATSTORE
or the National Eutrojj?.hication Survey, it is very important to obtain
current information on a lake in order to evaluate its present character-
istics. One must be careful to note trends that may have occurred over
time so as to fully understand the extent to which the reference lake
represents natural conditions.
Comparison
The reference site will have been selected on the basis of physical simi-
larity with the study area, and upon the determination that it reflects
natural conditions or conditions as close to natural as can be found.
Subsequent comparisons for the purpose of describing attainable uses will
be based on comparisons of the chemical and biological properties of the
two water bodies. Similarities and differences in chemical and biological
characteristics can be examined to identify causes of use impairment, and
potential- uses can be determined from an analysis of the lake's response to
the abatement of the identified causes of impairment.
Comparisons of individual chemical and biological parameters can be made by
using simple statistics such as mean values and ranges for the entire data
base or that part of the data base which is considered appropriate to re-
flect present conditions. Seasonal and monthly statistics can also be used
for lakes which demonstrate major changes throughout the year.
In addition to individual parameters, water quality and biological indices
are useful for comparisons. Water quality indices summarize a number of
water quality characteristics into a single numerical value which can be
compared to standard values that are indicative of a range of conditions.
The National Sanitation Foundation index, the Dinius water quality index,
and the Harkins/Kendall water quality index, each of which may provide
insight into the study site, are discussed in Chapter III of the Technical
Support Manual (U.S. EPA, 1983^).
Biological indices to be considered include: diversity indices which
evaluate richness and composition of species; community comparison indices
IY-7
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which measure similarities or dissimilarities between entire communities;
recovery indices which indicate the ability of an ecosystem to recover from
pollutant stress; and the Fish and Wildlife Service Habitat Suitability
Index which examines species habitat requirements. These indices are dis-
cussed in detail in Chapter IV of the Technical Support Manual (U.S. EPA,
1983])). Another useful tool which is described in that Manual is cluster
analysis, which is a technique for grouping similar sitesor sampling
stations on the basis of the resemblance of their attributes (e.g., number
of taxa and number of individuals).
Statistical tests can be used to 'determine whether water quality or any
other use attainment indicator at the study site is significantly different
from conditions at the reference site or sites. Several of these tests are
described in Volumes I and II of the Technical Support Manual (U.S. EPA,
1983J), 1984).
CURRENT AQUATIC LIFE PROTECTION USES
The actual aquatic life protection uses of a water body are defined by the
resident flora and fauna. The prevailing chemical and physical attributes
will determine what biota may be present, but little need be known of these
attributes to describe current uses. The raw findings of a biological sur-
vey may be subjected to various measurements and assessments, as discussed
in Section IV (Biological Evaluations) of the Technical Support Manual
(U.S. EPA, 1983J)). After performing an inventory of the flora and fauna
(preferably an historical inventory to reflect seasonal changes) and
considering diversity indices or other measures of biological health, one
should be able to adequately describe the condition of the aquatic life in
the lake.
CAUSES OF IMPAIRMENT OF AQUATIC LIFE PROTECTION USES
If the biological evaluations indicate that the biological health of the
system is impaired relative to a "healthy" reference aquatic ecosystem (as
might be determined by reference site comparisons), then the physical and
chemical evalutions can be used to pinpoint the causes of that impairment.
Figure IV-1 shows some of the physical and chemical parameters that may be
affected by various causes of change in a water body. The analysis of such
parameters will help clarify the magnitude of impairments to attaining
other uses, and will also be important to the third step in which potential
uses are examined.
ATTAINABLE AQUATIC LIFE PROTECTION USES
A third element to be considered is the assessment of potential uses of the
water body. This assessment would be based on the findings of the physi-
cal, chemical and biological information which has been gathered, but
additional study may also be necessary. A reference site comparison will
be particularly important. In addition to establishing a comparative
baseline community, the reference site provides insight into the aquatic
life that could potentially exist if the sources of impairment were
mitigated or removed.
IV-8
-------
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The analysis of all information that has been assembled may lead to the
definition of alternative strategies for the management of the lake at
hand. Each such strategy corresponds to a unique level of protection of
aquatic life, or aquatic life protection use. If it is determined that an
array of uses is attainable, further analysis which is beyond the scope of
the water body survey would be required to select a management program for
the lake.
One must be able to separate the effects of human intervention from natural
variability. Dissolved oxygen, for example, may vary seasonally over a
wide range in some areas even without anthropogenic effects, but it may be
difficult to separate the two in order to predict whether removal of the
anthropogenic cause will have a real effect. The impact of extreme storms
on a water body, such as the effect of Hurricane Agnes on Pennsylvania
lakes and streams in 1972, may completely confound our ability to
distinguish the relative impact of anthropogenic and natural influences on
immediate effects and long term trends. In many cases the investigator can
only provide an informed guess.
If a lake and stream system does not support an anadromous fishery because
of dams and diversions which have been built for water supply and recre-
ational purposes, it is unlikely that a concensus could be reached to
restore the fishery by removing the physical barriers—the dams--which
impede the migration of fish. However, it may be practical to install fish
ladders to allow upstream and downstream migration. Another example might
be a situation in which dredging to remove toxic sediments may pose a much
greater threat to aquatic life than to do nothing. Under the do nothing
alternative, the toxics may remain in the sediment in a biologically-
unavailable form, whereas dredging might resuspend the toxic fraction,
making it biologically available while facilitating wider distribution in
the water body.
The points touched, upon above are presented to suggest some of the phenom-
ena which may be of importance in a water body survey, and to suggest the
need to recognize whether or not they may realistically be manipulated.
Those which cannot be manipulated essentially define the limits of the
highest potential use that might be realized in the water body. Those that
can be manipulated define the levels of improvement that are attainable,
ranging from the current aquatic life uses to those that are possible
within the limitations imposed by factors that cannot be manipulated.
PREVENTIVE AND REMEDIAL TECHNIQUES
Uses that have been impaired or lost can only be restored if the conditions
responsible for the impairment are corrected. In most cases, impairment in
a lake can be attributed to toxic pollution or nutrient overenrichment.
Uses may also be lost through such activities as the disposal^of dredge and
fill materials which smother plant and animal communities, through
overfishing which may deplete natural populations, and the destruction of
freshwater spawning habitat which will cause the demise of various fish
species. One might expect losses due to natural phenomena to be temporary
although man-made alterations of the environment may preclude restoration
by natural processes.
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Assuming that the factors responsible for the loss of species have been
identified and corrected, efforts may be directed toward the restoration of
habitat followed by natural repopulation, stocking of species if habitat
has not been harmed, or both. Many techniques for the improvement of
substrate composition in streams have been developed which might find
application in lakes as well. Further discussion on the importance of
substrate composition will be found in the Technical Support Manual (U.S.
EPA, November 1983bj.
The U.S. EPA National Eutrophication Study and companion National Eutro-
phication Research Program resulted in the development and testing of a
number of lake restoration techniques. In the material to follow, an
overview is provided of a number of projects sponsored by the U.S. EPA in
which these techniques were applied. This is an overview that is not
intended to' be exhaustive in detail. For further information, the reader
is referred to a manual on lake restoration techniques that is currently in
preparation by U.S. EPA and the North American Lake Management Society.
Dredging
Introduction
Dredging to remove sediments from lakes has several objectives: to deepen
the lake, to remove nutrients associated with sediment, to remove toxics
trapped in bottom sediment, and to remove rooted aquatic plants. Dredged
lakes generally show improved aesthetics, and often enjoy improved fish
habitat as shown by increased growth of fish (Peterson, 1981). The
following sections summarize the objectives of lake dredging programs, the
environmental concerns associated with sediment removal, and the methods
used in implementing dredging projects.
Lake Conditions Most Suitable for Sediment Removal. Dredging to improve
lake conditions is better suited for some lak-es than others. Obviously, a
lake with a sediment-filled basin is a prime candidate for dredging. Other
considerations are lake size, the presence of toxics in the sediment,
dredging cost, and sedimentation rate. Toxics are of concern because they
may be released to the water column during the dredging operation. Because
of dredging costs, the dredging of large areas is not feasible. Lakes that
have been dredged in whole or in part range in size from 2 hectares (ha) to
1,050 ha (Peterson, 1981).
The practicality of sediment removal as a lake restoration technique also
depends on the depth of sediment to be removed. Lakes with surface sedi-
ment that is highly enriched relative to underlying sediment are best
suited for dredging projects. Dredging will not be cost effective in lakes
with high sedimentation, rates. The effect of sediment removal lasts longer
in water bodies with smaller ratios of watershed area to lake surface area
(Peterson, 1981). One other consideration in dredging projects is the dis-
posal of the dredged material. "Clean" sediment may be sold as landfill to
offset the cost of dredging. However, the disposal of contaminated
sediment may add considerably to the overall cost of the restoration
program.
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Purpose
Lakes in colder sections of the United States require a mean depth of about
4.5 m or greater to avoid winter fish kills; thus, lake deepening projects
may help assure fish survival (Peterson, 1981). Removal of sediment con-
taining high concentrations of nutrients helps to control algal growth.
The resultant decreased algal growth is also beneficial for fish popula-
tions. These purposes are explained in greater detail in the following
sections. Examples of lakes that have been dredged for the aforementioned
purposes are summarized in a separate section, Case Histories.
Removal of Nutrients. The primary nutrient of concern in dredging opera-
tions is phosphorus. Removal of enriched sediment reduces the internal
phosphorus load, as internal phosphorus cycling can amount to a major
portion of the total loading. Peterson (1981) cited these examples of
lakes in which a large percentage of the total phosphorus was attributed to
internal sources:
(1) Linsley Pond, Connecticut—internal phosphorus was about 45 per-
cent of the total phosphorus loading (Livingston and Boykin,
1962);
(2) Long Lake, Washington—phosphorus loading from sediment was 25-50
percent of the external loading (Welch, et al., 1979); and
(3) White Lake, Michigan—about 40 percent of the total phosphorus
loading was contributed by sediment phosphorus regeneration (Jones
and Bowser, 1978).
Because such large amounts of phosphorus are found within the sediments,
dredging may be a feasible means by which to greatly reduce internal
loading.
Lake Deepening. Summer stratification and vertical mixing characteristics
change with increasing depth. In addition, a larger volume of hypolimnetic
water, and a larger quantity of dissolved oxygen, are present in deeper
lakes (Stefan and Hanson, 1981). Therefore, assuming identical rates of
benthic oxygen uptake per unit area, the hypolimnion of a shallow lake will
be depleted sooner than the hypolimnion of a deeper lake. Summer overturn
due to wind-induced mixing may be frequent in shallow lakes. Therefore,
dredging to increase depth may help to reduce the frequency of overturn.
Increased lake volume may also help reduce water temperature. Reduced
water temperature increases oxygen solubility and decreases metabolic rates
of organisms. Therefore, algal growth rates and hypolimnetic oxygen deple-
tion may be slowed (Stefan and Hanson, 1981);
Removal of Toxics. The bottom sediment may be a sink for toxic and hazard-
ous materials as well as nutrients. Toxics in sediments pose a potentially
serious problem, although there is a paucity of information concerning the
direct effects of contaminated sediment on organisms. Another major con-
cern about sediments containing toxics is the possible introduction of
toxics into the food web, and the bioaccumulation and biomagnification of
toxics that may follow.
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Macrophyte Removal. Rooted aquatic macrophytes can be removed by dredging.
Aquatic plants are most often removed for reasons of aesthetics or inter-
ference with recreational uses. However, the role of macrophytes in
internal nutrient cycling also justifies their removal. Barko and Smart
(1980) demonstrated that Egeria densa, Hydrill a vertici11ata, and Myrio-
phyllum spicatum could obtain their phosphorus nutrition exclusively from
the sediments. When the plants die and decompose, nutrients in soluble
form may be released to the water column, or be returned to the sediments
as particulate matter.
Some researchers contend that healthy aquatic macrophytes obtain nutrients
from the sediment and excrete them to the surrounding water (Twilley, et
al., 1977; Carignan and Kalff, 1980). There is considerable evidence to
show that large quantities of nutrients are recycled to the lake when
plants die and decay (Barko and Smart, 1980; Landers, 1982). Landers
(1982) found that senescing stands of Myriophyllum spicatum contained up to
1$ percent of the annual total phosphorus loading in an Indiana reservoir.
Because aquatic macrophytes cause mobilization of nutrients from the soil,
their removal is a key to reducing the internal phosphorus load.
Environmental Concerns of Lake Dredging
Many of the environmental problems caused by dredging are associated with
resuspension of fine particulates. Increased turbidity reduces light
penetration; consequently, photosynthesis and phytoplankton production are
inhibited. Suspended sediments absorb radiation from the sun and transform
it into heat, thereby increasing the water temperature. Increases in
temperature affect the metabolic rate of organisms, in addition to reducing
the oxygen-holding capacity of the water. Dredging may also cause
increased nutrient levels in the water column, and potentially favorable
conditions for algal blooms (Peterson, 1981).
Toxic substances may also be liberated during dredging operations. For
example, the aldrin concentration in Vancouver Lake, Washington, was 0.012
mg/1 prior to dredging and increased by three times at one site and ten
times at another site during dredging (Peterson, 1979). Return flow from
settling ponds reached even higher concentrations, at times up to 0.336
mg/1.
Resuspended organic matter may present a different type of problem. Rapid
decomposition may deplete the available dissolved oxygen. This may be
especially important since the organic content of lake sediments can reach
80 percent on a dry weight basis (Wetzel, 1975). Although Peterson (1981)
noted that no lake dredging projects have caused this problem, the poten-
tial should be recognized.
Implementation of Lake Dredging Projects
Sediment Removal Depth. After it has been determined that sediment removal
is a viable lake restoration technique, a removal depth and method must be
selected. Sediment removal depth has been determined by several different
methods. The following paragraphs briefly describe two methods by which to
determine removal depth.
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Sediment Characterization. Studies of chemical and physical character-
istics of a lake bottom may show distinct stratification of sediment. The
greatest concentration of nutrients may be in a single layer, so that
removal of the layer will significantly affect the internal nutrient
loading. The sediment removal depth may be determined on the basis of
nutrient content and release rates for the layers of sediment.
For example, sediment in Lake Trummen,. Sweden, was characterized chemically
and physically, horizontally and vertically. The study showed a definite
layer of FeS-colored (black) fine sediment deposited on a brjpwn layer.
Based on aerobic and anaerobic release rates of PO,~-P and NH, -N, it was
decided that the black layer would be removed (Peterson, 1981). Born
(1979) noted that the ecosystem of Lake Trummen was restored following
dredging.
Lake Simulation. Another approach to determining sediment removal depth
uses a lake model to predict the lake depth necessary to prevent summer
destratification (Stefan and Hanson, 1980). This method of computation is
generally used for shallow lakes.
Stefan and Hanson (1981) modeled the Fairmont Lakes, Minnesota, to
determine the lake depth that would be required to prevent phosphorus
redrculation from the sediments. Using air temperature, dew point
temperature, wind direction, solar radiation, and wind speed, plus a
consideration of lake morphology, the model predicts temperature with
depth. Lake simulation helps determine the appropriate temperature and,
therefore, minimum depth for stable seasonal stratification. This method
of determining removal depth is based on the concept that shallow eutrophic
lakes can be dredged to such a depth that a stable system is formed. In
theory, phosphorus released from the sediment into the hypolimnion will be
recycled to the photic zone with diminished frequency. By controlling and
reducing the phosphorus concentration of the epilimnion, the standing crop
of algae will be decreased. The simulation results agreed with the
hypothesis of phosphorus release and recycling and the anticipated effects
of dredging (Stefan and Hanson, 1981).
The method of lake simulation does not consider sediment release rates.
Removal of the upper sediment layer may reduce nutrient levels in the
overlying water even though stratification is not stable. Therefore,
sediment release rates should also be examined along with the modeling
approach (Peterson, 1981).
Dredging Equipment. Barnard (1978) and Peterson (1979) describe various
dredges including the Mud Cat, the Bucket Wheel, and others, and their
advantages and disadvantages. The reader should refer to these sources,
especially Barnard (1978), for more detailed information.
The typical dredges are grab, bucket, and clamshell dredges which are
generally operated from a barge-mounted crane. These systems remove
sediment at nearly its in-s1te density, but removal volumes are limited to
less than 200,000 m . Turbidity is created due to bottom impact of the
bucket, the bucket pulling free from the bottom, bucket overflow and
leakage both below and above the water surface, and the intentional over-
flow of water from receiving barges to increase the solids content.
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Gutter-head dredges are the most commonly used in the United States. The
cutterhead dredge removes material in a slurry that is 10 to 20 percent
solids. These hydraulic dredges can remove larger volumes of sediment than
bucket dredges. Turbidity from hydraulic dredges is largely dependent on
pumping techniques and cutterhead configuration, size and operation.
Sediment Disposal. Dredged material disposal must also be considered in
sediment removal projects. Fill permits are required for the filling of
low-lying areas when the area exceeds 4.0 ha (10 acres) {Section 404,
Public Law 92-500).
Upland disposal sites, which do not require Federal permits, commonly
employ dikes to retain dredged material. Dike failure and underdesigned
capacity are two major problems with upland disposal areas.
Several documents prepared by the U.S. Army Corps of Engineers contain
useful information about dredged material disposal. They include: Treat-
ment of Contaminated Dredged Material (Barnard and Hand, 1978), Evaluation
of Dredged Material Pollution Potential (Brannon, 1978), Confined Disposal
Area Effluent and Leachate Control (Chen, et al., 1978), Disposal Alterna-
tives for Contaminated Dredged Material as a Management Tool to Minimize
Adverse Environmental Effects (Gambrell, et al., 1978), Upland and Wetland
Habitat Development with Dredged Material: Ecological Considerations
(Lunz, et al., 1978), Guidelines for Designing, Operating, and Managing
Dredged Material Containment Areas (Palermo, et al., 1978), and Productive
Land Use of Dredged Material Containment Areas (Walsh and Malkasain, 1978).
Lake Dredging Case Studies
Peterson (1981) lists 64 sediment removal projects in the United States
that are in various stages of implementation. Several of these projects
will be considered in more detail in the following section.
Lilly Lake, Wisconsin. Lilly Lake has a surface area of 35.6 ha, a maximum
depth of 1.8 m and a mean depth of 1.4 m. The main problem in Lilly Lake
was excessive macrophyte growth,. resulting in an accumulation of organic
detritus and bottom sediment. Macrophytes also curtailed recreational
activities such as boating and fishing. Winter fish kills were common in
Lilly Lake.
Dredging began in July 1978 and continued through October of the same year.
During dredging operations, the 5-day BOD increased by 1-2 mg 0?/liter, and
turbidity rose by 1-3 formazin units. Ammonia concentration increased from
0.01 mg/liter to a high of 5.5 mg/liter when dredging was halted in Octo-
ber. Prior to dredging, chlorophyll-£ levels averaged 2.5 ug/liter to 3.0
ug/liter. Immediately after dredging commenced, chlorophyll-a reached a
concentration of 27 ug/liter, and then decreased to levefs of 12-18
ug/liter. . Productivity also increased from pre-dredging levels of about
200 mg C/mJ/d to an average of 750 mg C/m /d in 1978 (Peterson, 1981).
Dredging began again in May 1979 and was completed by September. Maximum
depth was increased to 6.5 m following dredging. The water quality in 1980
was improved over previous?years, and the macrophyte biomass was reduced
from 200-300 g dry weight/nr to nearly zero.
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Steinmetz Lake. New York. Steinmetz Lake is 1.2 ha in area, and has a mean
depth of 1.5 m and a maximum depth of 2.1 m. Weed growth, algal growth and
highly turbid water were the major concerns.
Restoration included complete drawdown, sediment removal and stormwater
drainage diversion. The removed sediment was then replaced with clean
quarry sand. This method does not increase lake depth, but produces a new,
clean substrate.
Short term results of the restoration project were: increased Secchi disc
readings (from 1.25 m to the maximum lake depth), decreased chlorophyll-^
levels (from 10.4 ug/liter to 0.1 ugVliter), and reduced aquatic macrophyte
biomass (from 30-50 g wet weight/m to virtually zero) (Peterson, 1981).
After the treatment, plants grew where tracked vehicles forced organic
sediment through the sand cover. The number of people using the lake for
recreational purposes increased from almost none to over 3,000.
Lake Herman, South Dakota. Lake Herman has a surface area of 526 ha, a
maximum deptJi af 2.4 m and a mean depth of 1.7 m. The basin has a volume
of 8.9 x 10 m (2,642 million gallons). Farming practices in the water-
shed surrounding the lake have caused high nutrient concentrations and
excessive sedimentation. Lake Herman is primarily nitrogen limited and
nitrogen frequently declines to zero during algal blooms.
The dredging project was implemented to deepen the lake and remove the
nutrients, associated with the sediment. Hydraulic dredging removed about
48,000 m of silt from the lake, increasing the mean depth from 1.7 m to
about 3.4 m. Dredged material was deposited in an area adjacent to the
lake. Shortly after the dredging operation commenced, orthophosphorus
concentrations increased from 0.13 mg P/liter to more than 0.56 mg P/liter
(Peterson, 1981). Phytoplankton blooms did not accompany the increased
phosphorus concentrations because the lake is nitrogen limited. Although
no major increase in phytoplankton productivity was observed, the high
phosphorus concentrations attributable to phosphorus released to the water
column during dredging points out a potentially -serious problem that may
accompany hydraulic dredging operations.
Nutrient Precipitation and Inactivation
Introduction
Many eutrophic lakes respond slowly following nutrient diversion because of
poor flushing rates that facilitate sedimentation, and because of continued
internal phosphorus recycling. Phosphorus recycling is controlled by
precipitation and inactivation techniques generally used to remove
phosphorus from the water column and control its release from bottom
sediments. Chemical precipitants used for this purpose include salts of
aluminum, iron, and calcium. Calcium (II) has limited use in lakes because
it is ineffective below pH 9. Iron salts are not suitable inactivants for
long-term phosphorus control, since anoxic conditions reduce iron
complexes. This releases phosphorus and iron in the soluble state (Fe III
- Fe II). Therefore, aluminum compounds such as aluminum sulfate and
sodium aluminate are the most widely used. Zirconium and lanthanum (rare
earth elements) have proved effective in phosphorus removal, but more
IY-16
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research is needed on direct toxicity and general health effects before
this technique receives large-scale use.
Suitable Lake Types. Certain lake types are better suited to nutrient
precipitation and inactivation than others. Lakes should have moderate to
high retention times (several months or longer), since the treatment will
not be effective if there is a rapid flow-through of water. A water-
phosphorus budget is useful in assessing the significance of retention
time.
Nutrient precipitation and inactivation is generally implemented following
nutrient diversion, but this method of lake restoration will not be effec-
tive if the diversion is insufficient. Lakes with low alkalinity will
exhibit excessive pH shifts unless the lake is buffered or a mixture of
alum and sodium aluminate is used as precipitant. Finally, in lakes with
large littoral areas, phosphorus that is derived from groundwater, trans-
located from sediments by macrophytes, or resuspended by some activity that
stirs up sediment deposits may cause higher phosphorus concentrations than
expected.
Purpose
Phosphorus precipitation and inactivation techniques are used in water
bodies with high concentrations of phosphorus in the water column and the
sediment. Such a condition is generally indicated by nuisance algal
blooms. Immediate results of phosphorus precipitation include decreased
turbidity and algal growth. Application of aluminum compounds, primarily
aluminum sulfate and sodium aluminate, may also effectively control the
release of phosphorus from the sediment.
Environmental Concerns of Nutrient Precipitation
One immediate response of phosphorus precipitation is a reduction in tur-
bidity. The increased light penetration could stimulate increases in
rooted plant biomass. Other undesirable side-effects include reduced
planktonic microcrustacean species diversity and toxic effects of residual
dissolved aluminum (RDA) on aquatic biota. Laboratory research is cur-
rently underway to enlarge the aquatic toxicity data base available for the
U.S. EPA to develop water quality criteria for aluminum for the protection
of aquatic life. Aluminum toxicity is pH dependent and it becomes
extremely toxic below pH 5. Cooke and Kennedy (1981) cited the following
laboratory studies regarding the possible toxic effects on the biota of
phosphorus precipitation using aluminum compounds:
(1) Daphnia magna had a 16 percent reproductive impairment at 320 ug
Al/1 (Biesinger and Christian, 1972);
(2) A few weeks exposure to 5,200 ug Al/1 seriously disturbed rainbow
trout tested in flow through bioassays (Everhart and Freeman,
1973);
(3) No obvious effect on rainbow trout after long-term exposure to 52
ug Al/1 (Kennedy, 1978; Cooke, et al., 1978);
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(4) Daphnla magna survival was reduced 60 percent in 96-hr tests of
concentrations to 80 ug Al/1 (Peterson, et al., 1974, 1976); and
(5) No negative effects on fish (Kennedy and Cooke, 1974; Bandow,
1974; Sanvllle, et al., 1976) or benthlc invertebrates (Narf,
1978) after full-scale lake treatments. Cooke and Kennedy (1981)
noted that there were no toxic effects on fish as long as the pH
remains in an acceptable range and the RDA is less than about 50
ug Al/1.
Implementation of Nutrient Precipitation Projects
The following factors should be considered for phosphorus precipitation/
inactivation through chemical application: dose, choice of dry or liquid
chemical, depth of application, application procedure, and season (Cooke
and Kennedy, 1981).
Dose Determination. Cooke and Kennedy (1980) and Cooke and Kennedy (1981)
describe some methods for determining dose. A dose of aluminum that re-
duces pH to 6.0 is considered "optimal." The residual dissolved aluminum
should remain below 50 ug Al/1, the level at which aluminum begins to
elicit toxic effects. A simplified method for dose determination is
outlined below (Cooke and Kennedy, 1980).
Procedure:
(1) Obtain representative water samples from the lake to be treated.
Care should be exercised in selecting sampling stations and depths
since significant heterogeneities,- both vertical and horizontal,
commonly occur in lakes. Samples should be collected as close to
the anticipated treatment date as possible.
(2) Determine the total alkalinity and pH of each sample. Total
alkalinity, an approximate measure of the buffering capacity of
lake water, will dictate the amount of aluminum sulfate (or
aluminum) required to- achieve pH 6 and thus optimum dose. Addi-
tional chemical analyses can be performed, depending on the
specific needs of the Investigator. For example, phosphorus
analyses before and after laboratory treatment would allow
estimation of anticipated phosphorus removal effectiveness.
(3) Determine the optimum dose for each sample. Initial estimates of
this dose, based on pH and alkalinity, can be obtained from Figure
IV-2. More accurate estimates should be made by titrating samples
with fresh stock solutions of aluminum sulfate of known aluminum
concentration using a standard burette or.graduated pipette. The
concentration of stock aluminum solutions should be such that pH 6
can be reached with additions of 5 to 10 mill niters per liter of
sample. Samples must be mixed (about 2 minutes) using an overhead
stirring motor and pH changes monitored continuously using a pH
meter. Optimum dose for each sample will be the amount of
aluminum, which when added, produces a stable pH of 6.0.
TV-18
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ALUMINUM DOSE (mg Al/l) TO OBTAIN pH 6.0
3
O
ca
O
O)
H-
o
I-
250
200
150
100 -
Figure IV-2.
Estimated aluminum sulfate dose (mg/1) required to
obtain pH 6 in treated water of varying initial alkalinity
and pH (from Cooke and Kennedy, 1980).
IV-19
-------
(4) The relationship between total alkalinity and optimum dose can be
determined using information from each of the above titrations by
plotting optimum dose as a function of alkalinity. This relation-
ship will allow determination of dose at any alkalinity with the
range tested.
Liquid alum and liquid sodium aluminate generally form a better floe and
are more effective than the dry forms (Cooke and Kennedy, 1981). If only
dry alum is available, it can be mixed in tanks to form a slurry before
application.
Depth of Application. Aluminum salts can be applied to surface water, or
at predetermined depth(s), depending upon treatment objectives. A surface
application is generally needed to remove phosphorus from the water column,
whereas hypolimnetic treatment controls the release of phosphorus from
sediments.
Time of Application. Both particulate and dissolved forms of phosphorus
are efficiently removed by the aluminum floe as it settles to the bottom.
Whether there is an optimum season for the application of aluminum salts
for the removal of various forms of phosphorus is debatable, as discussed
by Cooke and Kennedy (1981).
Nutrient Precipitation Case Studies
Although at least 28 lakes have been reported in the literature that have
been treated by the phosphorus inactivation/precipitation technique, there
is a paucity of information regarding post-treatment effects. The
following sections summarize five case histories that are representative of
different approaches, have long-term monitoring, or illustrate strengths
and shortcomings of this technique. Information concerning dose, method of
application, cost, and long-term effects on additional restoration projects
employing inactivation/precipitation techniques is found in Cooke and
Kennedy (1981).
Horseshoe Lake, Wisconsin. Horseshoe Lake has a surface area of 8.9 ha, a
maximum depth of 16.7 m, and a mean depth of 4.0 m. It is the first
reported full scale in-lake inactivation experiment in the United States
(Funk and Gibbons, 1979). Prior to treatment, the lake exhibited algal
blooms, dissolved oxygen depletions and fish kills. High nutrient levels
were attributed to agricultural and natural drainage, and to waste dis-
charges from a cheese-butter factory prior to its closing in 1965.
Alum was applied, just below the water surface, in May 1970. No decrease
in phosphorus level was observed until after fall circulation, when con-
centrations decreased substantially. Reduced phosphorus concentrations
were observed in both the epilimnion and the hypolimnion. Although hypo-
limnetic phosphorus increased slightly every year following treatment, it
was controlled for about 8 years. Secchi disc transparency also increased
and no fish kills have occurred since the alum application. Additional
information about the restoration of Horseshoe Lake is provided by
Peterson, et al. (1973).
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Medical Lake, Washington. Medical Lake covers an area of 64 ha. It has a
maximum depth of 18 m and a mean depth of 10 m. Prior to treatment, the
lake exhibited nuisance algal blooms, summer anoxia and high nutrient con-
centrations, primarily because of internal nutrient cycling. Treatment
with alum was chosen as the best method for inactivating phosphorus in
Medical Lake.
Alum was applied at the surface or at 4.5 meters, depending upon whether
the area was shallow or deep. Application began in August 1977 and con-
tinued over a 5-week period.
Water quality monitoring through June 1980 showed that alum treatment
successfully reduced phosphorus levels, eliminated algal blooms and in-
creased water clarity. Total and orthophosphorus levels prior to alum
treatment were 0.47 mg/liter and 0.32 mg/liter, respectively. These levels
decreased about 87 and 97 percent, respectively. Chlorophyll-a_ decreased
from-a mean monthly value of 25.2 mg/m prior to alum treatment, to 3.2
mg/m following treatment. Secchi disc transparency improved from a mean
depth of 2.4 meters to 4.9 meters. Whereas the lake did not support a
fishery prior to treatment, a rainbow trout population flourished after
phosphorus precipitation/inactivation. -No negative impacts on biota were
observed although the concentration of dissolved aluminum increased to 700
ug Al/1 during treatment. Post-treatment levels fell to 30-50 ug/1 (Cooke
and Kennedy, 1981). Detailed results of water quality monitoring following
phosphorus precipitation/inactivation treatment are presented in Gasperino,
et al. (1980aJ and Gasperino, et al. (1980b_).
Annabessacook Lake, Maine. Annabessacook Lake, located in central Maine,
covers an area of about 575 ha, and has a hypolimnetic area of 130 ha. The
mean lake depth is 5.3 m and the maximum depth is 14.9 m. High levels of
phosphorus in the water column and sediments were believed to be respon-
sible for blue-green algal blooms. Industrial and municipal wastewater
inputs contributed to high phosphorus levels prior to 1972, and internal
nutrient cycling caused continued high nutrient levels in the lake
(Dominie, 1980).
Annabessacook Lake underwent an extensive lake restoration program, in-
cluding nutrient diversion, agricultural waste management and in-lake
nutrient inactivation. Point sources were diverted from the lake and
agricultural waste management plans were implemented. Laboratory testing
showed that aluminum treatment was a feasible alternative for lake res-
toration. Because the lake water has a low alkalinity, a combination of
aluminum sulfate and sodium aluminate was used to provide sufficient buf-
fering capacity to moderate potential pH shifts.
After the aluminum application and commencement of waste management pro-
grams, the following changes were observed (Dominie, 1980):
o Total phosphorus mass in the lake was reduced from over 2,200
kilograms (kg) in 1977 to 1,030 kg in 1978.
o Internal recyclable phosphorus was reduced 65 percent from 1,800
kg in 1977 to 625 kg in 1979.
IY-21
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o The average June chlorophyll-a concentration decreased from 11.5
ug/1 (1977) to 6.2 ug/1 (197817"
o Secchi disc depth for June (monthly mean) increased from 2.0 m
(1977) to 3.1 m (1978).
Additional information on the restoration of Annabessacook Lake is found in
Dominie (1980), Gordon (1980), Cooke and Kennedy (1981), and U.S. EPA
(1982).
Liberty Lake, Washington. Liberty Lake, in Spokane County, has a surface
area of 316 ha. 'The lake has a mean depth of 7.0 m, and a maximum depth of
9.1 m. A combination of septic tank drainage, urban runoff, and poor solid
waste disposal practices caused excessive nutrient levels and heavy blooms
of blue-green algae in the lake.
In 1974, Liberty Lake was treated with aluminum sulfate to precipitate and
inactivate phosphorus. Jar tests and in situ tests were made to determine
dosage. The alum slurry was applied to the surface. After application of
aluminum sulfate, total phosphorus was reduced from 0.026 mg/1 to less than
0.015 mg/1. Water clarity increased following the treatment. Although
alkalinity and pH dropped, the effect was short lived and these parameters
returned to pretreatment levels within 24 to 48 hours (Funk and Gibbons,
1979).
The -treatment effectively controlled algal blooms from 1974 to 1977. Heavy
blooms equivalent to those prior to treatment did occur in the fall of
1977.
Dollar Lake and West Twin Lake, Ohio. Dollar Lake has a surface area of
2.22 ha, a mean depth of 3.89 m and a maximum depth of 7.5 m. West Twin
Lake, which is adjacent to Dollar Lake, is larger, with a surface area of
34.02 ha, a mean depth of 4.34 m and a maximum depth of 7.50 m. Septic
tank drainage was largely responsible for eutrophic conditions. Although
septic effluent was diverted in 1971-72, algal blooms continued, partly
because of internal cycling of phosphorus.
Aluminum sulfate was applied to the hypolimnion of the lakes to inactivate
and precipitate phosphorus. Following the alum application, both lakes
showed decreased phosphorus content in the water column and improved water
transparency. Blue-green algae dominance in West Twin Lake was reduced by
80 percent (Funk and Gibbons, 1979; Cooke and Kennedy, 1981). Zooplankton
populations were affected, and the dominant species shifted from Cladocera
to Copepoda. Hypolimnetic phosphorus concentration in Dollar and West Twin
Lakes remained low for four years after treatment.
Aeration/Circulation
Introduction
Aeration/circulation is a potentially useful technique for treating
symptoms of eutrophication. The range of aeration/circulation techniques
can be divided into two major groups: artificial circulation and hypo-
limnetic aeration. Both of these techniques increase the dissolved oxygen
IY-22
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concentration of hypolimnetic waters. The two techniques differ in that
hypolimnetic aeration aerates hypolimnetic waters without mixing them with
surface waters while artificial circulation breaks down stratification by
mixing the upper and lower strata of the water column. These techniques
can be used to enhance the habitat of aquatic biota and improve water
quality by alleviating problems created by stratification and deoxygenation
of the hypolimnion.
Both techniques restore oxygen to anaerobic bottom waters. These res-
toration procedures lead to habitat expansion for zooplankton, benthos and
fish. Destratification is usually benefical for warmwater fish, promoting
an increase in the depth distribution. Howev-er, complete mixing may
eliminate coldwater habitats and fish such as salmonids may disappear from
the lake.
Lakes Best Suited for Aeration/Circulation. Anaerobic bottom waters of a
stratifiedlakecanbe. oxygenated5yaeration/circulation techniques.
Either method may be implemented when the primary purpose of treatment is
to alleviate "taste and odor" problems resulting from high concentrations
of Fe, Mn, ^S and other chemicals in an anoxic hypolimnion. Both methods
expand or improve habitat for zooplankton, benthos, and warmwater fish.
However, artificial circulation and hypolimnetic aeration do not produce
the same effects in lakes.
Artificial aeration may cause the replacement of blue-green algae com-
munities by more desirable communities of green algae, while hypolimn.etic
aeration generally does not have an effect on phytoplankton. Since
hypolimnetic aeration does not effect mixing of surface and hypolimnetic
waters, nutrient concentrations in the euphotic zone are basically
unaffected when this technique is employed. Consequently, hypolimnetic
aeration generally does not affect the phytoplankton community. In
contrast, artificial circulation vertically mixes the water column and can
increase nutrient concentrations in the euphotic zone. In a series of
experiments, Shapiro (1973) showed that natural populations of blue-green
algae were replaced by green algae after enrichment with phosphorus and
nitrogen when carbon dioxide was added or pH was lowered. These results
indicate that green algae can outcompete blue-green algae under enriched
nutrient conditions as long as C02 is abundantly available.
When control of algal blooms is not a prime consideration and a coldwater
supply is necessary, the preferred method is hypolimnetic aeration. A cold
hypolimnion is needed for survival of coldwater fish, and thus hypolimnetic
aeration is recommended when improvement of fisheries is the only con-
sideration. In southern lakes, high water temperatures in the epilimnion
and metalimnion often preclude survival of coldwater fish; therefore, it is
necessary to preserve the integrity of the water layers, including the
colder hypolimnion, and artificial destratification would not be appro-
priate.
Artificial circulation is preferred when limitation of algal biomass is
desired, oxygenation of the metalimnion is needed, or a completely mixed
water column is acceptable. Artificial circulation is also suitable for
northern lakes where the temperature of surface waters does not exceed 22 °C
during the summer (Pastorak, et al., 1981).
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Purpose
Artificial Circulation. Anaerobic conditions in the hypolimnion of a
stratifiedTikerestrict the vertical distribution of fish, eliminate
certain benthic organisms, and may cause the release of nutrients and toxic
substances to the overlying water. Artificial circulation alleviates these
problems by destratifying and oxygenating bottom waters of the lake. The
water becomes oxygenated primarily through atmospheric exchange at the
water surface. Except in very deep lakes, the transfer of oxygen from air
bubbles of diffused air systems is relatively small.
By aerating and destratifying lakes, artificial circulation improves water
quality, decreases algal growth, and improves fish habitat. These effects
are described below.
Elimination of Taste and Odor Problems. Generally, artificial destrati-
fication oxygenates anaerobic hypolimnetic waters. Anaerobic conditions
near the lake bottom cause the release of reduced chemical species from
sediments to the water column. Water supply utilities experience water
quality control problems resulting from, the accumulation of iron (Fe),
manganese (Mn), carbon dioxide (CO?), hydrogen sulfide (F^S), ammonium ions
(NH.+) and other chemicals in the nypolimnion. As hypolimnetic waters are
brought to the lake surface during artificial circulation, gases such as
C02, H2S and NH3 are released to the atmosphere. Artificial circulation
increases hypolimnetic oxygen, and raises the redox potential near the lake
bottom. The result is decreased concentrations of reduced chemical
species, thereby eliminating taste and odor problems..
Decreased Algal Growth. In some cases, algal production is reduced through
artificial circulation. Pastorak, et al. (1981) cited Fast (1975) for
several mechanisms that cause reduced algal growth. Internal nutrient
loading may be reduced through the elimination of anaerobic conditions that
cause nutrient regeneration. Artificial circulation also increases the
mixed depth of the algae, thereby reducing algal growth through light
limitation. When mixing is induced during an algal bloom, the algae are
distributed through a greater water volume, and lake water transparency
will increase immediately. In addition, as water is pumped to destratify
the lake, rapid changes in hydrostatic pressure and turbulence serve to
destroy phytopiankton.
Artificial circulation does not consistently decrease algal populations,
and may cause increased algal biomass in some instances. Pastorak, et al.
(1981) surveyed the literature covering 40 experiments in which destratifi-
cation was relatively complete. Only 26 experiments exhibited significant
changes in phytopiankton biomass, and of these, about 30 percent exhibited
increases in algae.
Forsberg and Shapiro (1981) found that changes in algal species composition
during artificial aeration depend primarily on the mixing rate. With slow
mixing rates, surface levels of total phosphorus and pH generally in-
creased, and the relative abundance of blue-green species such as Anabaena
circulinus and Microcystis aureginosis increased.
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The abundance of green algae and diatoms increased when faster mixing rates
were used. Complete chemical destratification caused by high mixing rates
was accompanied by large increases in surface total phosphorus and CCL
concentration. The green algae Sphaerocystis schroederi, Ankistrodesmus
falcatus and Scenedesmus spp., and the diatoms Nitzchia spp., Synedra spp.,
and Melosira spp. grew particularly well under these conditions (Forsberg
and Shapiro, 1981).
Benefits to Fish Populations. Artificial circulation may enhance fish
habitat and food supply, thereby potentially improving growth of fish,
environmental carrying capacity, and overall yield.
Low oxygen levels in the hypolimnion may prevent fish from using the entire
potential habitat. Oestratification and aeration of bottom waters may
allow fish to inhabit a greater portion, of the water column, expanding the
vertical distribution of warmwater fish.
Salmonids in particular may be restricted to a layer of metalimnetic
habitat, with warm water above and anaerobic conditions below. If surface
water temperatures remain below 22°C throughout the summer, as in northern
lakes, artificial circulation should increase habitat for cold-water fish.
In addition, summer-kill of fish due to anoxic conditions and toxic gases
may be prevented by artificial circulation.
Artificial circulation has also proved to be an effective method of
preventing over-winter mortality of salmonids. Whereas natural oxygen
concentrations may be depleted during the winter, aeration prior to ice
formation can provide sufficient oxygen for fish survival. Winter mor-
talities of fish in Corbett Lake, British Columbia, were prevented in this
way (Pastorak, et al., 1981).
Hypolimnetic Aeration and Oxygenation. Hypolimnetic aeration and oxygen-
ation add dissolved oxygen to the bottom waters without destratifying the
lake. Aeration of the hypolimnion occurs through oxygen transfer between
air bubbles and water, and oxygenation occurs more slowly than with
artificial circulation.
Major goals of programs employing hypo!imnetic aeration and oxygenation are
to improve water quality and provide habitat for coldwater fish. Unlike
artificial circulation, there is no evidence that hypolimnetic aeration
will control algal blooms.
Improvement of Water Quality. Hypolimnetic aeration minimizes taste, odor
and corrosion problems by oxygenating bottom waters, which raises the pH
and lowers concentrations of reduced compounds. Although artificial
circulation aerates the water column more rapidly, hypolimnetic aeration
maintains stratification, thereby retaining a coldwater resource.
Improvement of Fisheries. Hypolimnetic aeration creates habitat for cold-
water fish by oxygenating the cold bottom layers of a lake. Because the
lake does not become completely mixed as a result of hypolimnetic aeration,
a two-story fishery can develop. Aeration also enhances fish food supply,
since.the distribution and abundance of macroinvertebrates increases.
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Planktivorous fish may also find an increased food supply following hypo-
limnetic aeration. While phytoplankton abundance is generally unaffected,
zooplankton populations may expand their vertical range after treatment.
Fast (1971) found a significant increase in the population of Daphnia pulex
following aeration of Hemlock Lake, Michigan. He attributed the population
change to an expanded habitat, which allowed Daphnia to inhabit dimly lit
depths of the lake and avoid predation by troufT
Environmental Concerns of Aeration/Circulation
Most of the environmental concerns are associated with the use of arti-
ficial destratification systems, whereas very few adverse impacts of
hypolimnetic aeration are known. Hypolimnetic aeration has very little
influence on depth of mixing, pH of the water, sediment resuspension, and
algal densities. Adverse impacts of aeration/circulation, including
effects on water quality, nuisance algae, macrophytes and fisheries, are
described in the following sections. Examples of impacts of aeration/
circulation on lakes are presented later in a section on Case Histories.
The purpose of the present discussion of environmental concerns is to point
out adverse consequences that might occur as a result of artificial
destratification. Although these effects will not necessarily be seen, it
is instructive to recognize the potential problems that could arise, on a
site-specific basis.
Water Quality. Artificial circulation may cause several chemical and
physical changes that adversely affect water quality. The mixing of
nutrient rich hypolimnetic water could increase the concentrations of
nutrients in the upper water layers. Heightened concentrations of the
gases NH^ and HgS may also occur in surface water.
Turbulence due to mixing and aeration systems may further affect water
quality by resuspending silt, thereby increasing turbidity. Decreases in
water transparency after mixing may also be associated with surface algal
blooms (Pastorak, et al., 1980).
Nuisance Algae. Artificial circulation/destratification may produce un-
desirable changes in phytoplankton communities. For example, temporary
algal blooms may occur because of recycling of hypolimnetic nutrients and
elevation of total phosphorus. Such a rise in algal biomass may favor
blue-green algae by depleting C02 and keeping pH levels high.
Macrophytes. Improved water transparency following artificial circulation
may allow increased macrophyte growth. Rooted aquatic plants could expand
to nuisance levels, especially in lakes with shallow littoral shelves.
Fisheries. Where coldwater fish exist in the metalimnetic region, artifi-
cial circulation and the subsequent warming of bottom waters may eliminate
habitat for certain species. The surface temperatures of northern lakes
generally remain below 22°C, and thus the bottom waters will not be warmed
(as might occur in southern lakes), and habitat for coldwater fish will be
enhanced during circulation. Destratification and mixing can also lead to
dissolved oxygen decreases in the whole lake. In this instance, resus-
pension of bottom detritus increases the biochemical oxygen demand (BOD)
beyond the rate of reaeration (Pastorak, et al., 1981). Extensive
IY-26
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depletion of dissolved oxygen may be responsible for fish mortalities.
Aeration of Stewart Lake initially caused a decline in bluegill population,
presumably because of reduced dissolved oxygen (Pastorale, et al., 1981).
Fish kills may also be caused by supersaturated concentrations of nitrogen,
which may result from circulation or hypolimnetic aeration. In spring, M-
levels generally equilibrate at 100 percent saturation with respect to
surface temperature and pressure. Warming of the hypolimnion during the
summer results in supersaturation of N2 relative to surface temperature and
ambient temperature at depth. This fupersaturation of N- may induce gas
bubble disease in fish, causing stress or mortality .(Pfstorak, et al.,
1981). Although this has not been documented in lakes, dissolved nitrogen
concentrations of 115-120 percent saturation induced salmonid mortalities
in rivers (Rucker, 1972).
Implementation of Aeration/Circulation Projects
Aeration/circulation is a relatively inexpensive and efficient restoration
technique. The following sections briefly describe methods and equipment
used tn restoration projects employing artificial circulation or hypo-
limnetic aeration.
Artificial Circulation. Lake circulation techniques can be broadly classi-
fied in the categories of diffused air systems or mechanical mixing systems
(Lorenzen and Fast, 1977). Diffused air systems employ the "air-lift"
principle, as water is upwelled by a plume of rising air bubbles. Mechan-
ical systems move water by using diaphragm pumps, fan blades, or water
jets. Lorenzen and Fast (1977) reviewed the design and field performance
of various circulation techniques, and concluded that diffused air systems
are less expensive and easier to operate than mechanical mixing systems.
Diffused Air Systems. Diffused air systems inject compressed air into the
lake through a perforated pipe or other simple diffusers. Johnson and
Davis (1980) reviewed submerged jetted inlets and perforated pipe air-
mixing systems used in reservoirs. Hypolimnetic water is upwelled by the
rising air bubbles. Upon reaching the surface, this water flows out
horizontally and sinks, mixing with the warm surface water in the process.
The amount of water flow induced by the rising bubbles is a function of air
release depth and air flow rate. Artificial circulation is generally most
effective if air is injected at the maximum depth possible (Pastorak, et
al., 1981). In a thermally stratified lake, mixing will normally be in-
duced only above the air release depth. However, while an aerator located
near the surface of the lake may be unsuitable for destratifying a lake, it
may effectively prevent the onset of stratification (Pastorak, et al.,
1981).
Mechanical Mixing. Mechanical mixing devices such as pumps, fans'and water
jets are employed less frequently than diffused air systems. Pastorak, et
al. (1981) notes several instances in which mechanical mixing devices have
been successfully employed:
IV-27
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(1) Stewart Hollow .Reservoir and Vesuvius Reservoir, Ohio—a pumping
rate of 10.9 m /min was sufficient to destratify the reservoirs
within 8 days (Irwin, et al., 1966);
(2) Ham's Lake, Oklahoma—an axial-flow pump with a capacity of 102
m /min completely destratified the lake, which has a mean depth of
2.9 m, after 3 days of operation (Toetz, 1977).
On the other hand, mechanical mixing may not always be successful:
(1) West Lost Lake—a pumping capacity of 1.3 m /min over a period of
10.1 days was not sufficient to completely mix the lake (Hooper,
et al., 1953);
(2) Arbuckle. Lake, Oklahoma—an array of 16 pumps (total capacity
1,600 m /min) did not completely mix the lake, which has a mean
depth of 9.5 m (Toetz, 1979).
Artificial circulation techniques should be started before full development
of thermal stratification, because nutrients that become trapped in the
hypolimnion and then are recycled may cause, increased algal growth.
Lorenzen and Fast (1977) recopnerul about 9.2 m /min of air per 10 m of
lake surface (= 30 SCFM per 10 ft ) to adequately mix and aerate the water
column.
Hypolimnetic Aeration. Fast and Lorenzen (1976) reviewed designs of hypo-
limnetic aerators, and proposed the following divisions: mechanical agi-
tation systems, pure oxygen injection, and air injection systems (which
include full air-lift designs, partial air-lift designs, and downflow air
injection systems). Hypolimnetic aeration systems generally remove water
from the hypolimnion, aerate and oxygenate it, and then return the water to
the hypolimnion.
Mechanical Agitation. Mechanical agitation systems generally draw hypo-
limnetic water up a tube and aerate it at the surface through mechanical
agitation. Fast and Lorenzen (1976) noted that a surface agitator design
is most efficient for hypolimnetic aeration of shallow lakes where water
depth is insufficient to provide a large driving force for gas dissolution.
Oxygen Injection Systems. As in other hypolimnetic aeration systems, water
is removed from and returned to the hypolimnion. In oxygen injection
systems, nearly pure oxygen becomes almost completely dissolved when it is
returned to the hypolimnion (Fast and Lorenzen, 1976).
Air Injection Systems. The full air lift design is the least costly system
to construct, install and operate (Fast and Lorenzen, 1976; Fast, et al.,
1976; Pastorak, et al., 1981). In these systems, compressed air is in-
jected near the bottom of the aerator, and the air/water mixture rises. At
the water surface, air separates from the mixture and water is returned to
the hypolimnion.
Partial air lift designs are less efficient than full air lift designs.
Partial air lift systems aerate and circulate hypolimnetic water by an air
injection system, but the air/water mixture does not upwell to the surface.
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Air and water separate below the lake's surface and air rises to the
atmosphere while water returns to the hypolimnion (Fast and Lorenzen,
1976).
Aeration/Circulation Case Studies
Three case studies are presented in this section to summarize the effects
of artificial circulation on lakes.
Parvin Lake, Colorado. Parvin Lake is a 19 ha mesotrophic reservoir, with
a maximum depth of 10 m and a mean depth of 4.4 m. Summer surface tempera-
tures remain less than 21"C year-round.
The effects of artificial circulation on Parvin Lake were studied for two
years (Lackey, 1973). November 1968 to October 1969 was the control period
during which phytoplankton were sampled to provide baseline information.
The treatment year, when the destratification system operated continuously,
extended from November 1969 to October 1970.
Phytoplankton in Parvin Lake were affected in the following ways (Lackey,
1973):
o Abundance of green algae significantly decreased during treatment;
o Anabaena, a nuisance blue-green algae, followed a similar pattern
of abundance during both control and treatment years;
o Planktonic diatoms decreased in abundance during the treatment
winter.
Ham's Lake, Oklahoma. Pastorak.'et al. (1981) summarized the effects of
artificial destratification on Ham's Lake, Oklahoma. The lake, which has a
maximum depth of 10 m, and a mean depth of 2.9 m, covers an area of 40 ha.
Following destratification, the lake showed an increase in Seechi disc
depth, dissolved oxygen concentration, and phosphate concentration. Both
the density and the diversity of benthic organisms increased. Decreases i-n
concentrations of ammonium, nitrate, iron and manganese in the water column
were noted. Mo changes in algal density, chlorophyll-a, green algae,
blue-green algae, or the ratio of green algae/blue-green algae was
observed.
Kezar Lake, New Hampshire. Kezar Lake has an area of 73 ha, a maximum
depth of 8.4 m, arm a mean depth of 2.8 m. Artificial circulation was
imposed from July 16 to September 12, 1968, and became completely de-
stratified (Haynes, 1973). The responses of the lake to artificial
circulation were:
o Increases in Secchi disc depth, pH, dissolved oxygen concentra-
tion, phosphate, and total phosphorus;
o Decreases in ammonium, iron and manganese concentrations;
o Reductions in algal density, algal standing biomass, and blue-
green algae;
IY-29
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o Increases in green algae, and the ratio of green algae/blue-green
algae; and
o No change in mean chlorophyll-_a concentration.
Ottoville Quarry. Ohio. Ottoville Quarry is a small (0.73 ha) water-filled
quarry, with a maximum depth of 18 m. Prior to treatment, rainbow trout
(Salmo gairdneri) were unable to survive the summer because of high water
temperature and oxygen depletion. A program employing hypolimnetic oxy-
genation was implemented in 1973 (from July to September), and increased
summer dissolved oxygen concentrations from nearly zero to 8 mg/1 (Over-
hoi tz, et al., 1977). Aeration from May to October, 1974, caused dissolved
oxygen concentrations in the hypdlimnion to exceed 20 mg/1 by September.
Overholtz, et al. (1977) found that hypolimnetic aeration created an
environment suitable for rainbow trout survival while maintaining thermal
stratification in the quarry.
Lake Drawdown
Introduction
The primary purpose in restoration programs employing lake drawdown is to
control the growth of nuisance aquatic macrophytes. In general, the water
level in a lake is lowered sufficiently to expose the nuisance plants while
retaining an adequate amount of water in the lake to protect desirable fish
populations: This technique is effective for short-term control (1-2
years) of susceptible aquatic macrophytes. Secondary objectives include
turbidity control by sediment consolidation, reduction of nutrient release
from sediments (through sediment consolidation or removal), management of
fish populations and waterfowl habitats, repair of shoreline structures and
simultaneous use of other restoration methods such as covering sediment
with new clean material (Cooke, 1980£, 1980t>). Sediment consolidation may
also cause a slight increase in lake depth. The following sections expand
upon the technique of lake drawdown, including methods and case studies.
Lake Conditions Most Suitable for Lake Drawdown. Drawdown and sediment
consolidation may be feasible for the restoration of shallow lakes if two
conditions are met. The lake basin should have a shallow slope, so that a
small vertical decline in water level exposes a large part of lake bottom,
and the source of water must be controlled (Dooris, et al., 1982).
The nature of the lake sediment is particularly important to the success of
drawdown projects. The sediment that will be exposed must be able to dry
and consolidate quickly so that a prolonged dewatering period is not re-
quired, and the dried and compacted sediment should not rehydrate signifi-
cantly after the refilling of the lake basin. However, the sediment should
be of a consistency which would allow colonization by desirable plants and
benthic organisms (Dooris, et al., 1982).
Purpose
The main objective of lake level drawdown is to manage nuisance macrophytes
by destroying seeds and vegetative reproductive structures through exposure
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to drying and/or freezing conditions. In addition, dewatering and consoli-
dation of sediments alters the substrate, thereby eliminating conditions
required for the growth of certain aquatic plants. Sediment consolidation
also helps control turbidity, reduces nutrient release from sediments and
causes a slight deepening of the lake.
Lake drawdown can be used to enhance fisheries and waterfowl habitats. The
simultaneous use of other restoration techniques, such as sediment covering
or removal, will be even more effective for control of vegetation. The
period of dewatering may also be used to repair shoreline structures, such
as dams, docks and swimming beaches.
Environmental Concerns: of Lake Drawdown
There may be negative impacts of lake drawdown as well as desirable
effects. Negative environmental changes that may occur following drawdown
include establishment of resistant macrophytes, algal blooms, fish kills,
changes in littoral fauna, failure to refill, and decline in attractiveness
to waterfowl.
Algal blooms that occur after reflooding may be one of the undesirable
effects of drawdown. Geiger (1983) observed increases in total nitrogen,
total phosphorus, and chl orophyll-_a following drawdown of Blue Lake,
Oregon. The cause of such increases is unclear although it is postulated
that drawdown and exposure of sediments, and the subsequent aeration and
oxidation bring about nutrient release when the basin is reflooded. The
released nutrients are then available for algal growth.
Fish kills may be caused by drawdown, especially if the water level is
lowered during the summer. The warmer temperatures cause increased rates
of metabolism and heighten the sediment oxygen demand. However, Cooke
(1980a_) noted that a 2 m summer drawdown of Long Lake, Washington (maximum
depth 3.5 m) did not cause fish kills, and the dissolved oxygen remained
above 5 mg/1.
Drawdown and reflooding may cause changes in the diversity and density of
benthic fauna. Increases in invertebrate density, but decreases in species
diversity, have been observed following drawdown and reflooding (Cooke,
1980£). Summer drawdown and subsequent hardening of littoral soils may
reduce repopulation by insects. These changes may be detrimental to fish
and waterfowl.
The basin may not refill because of an insufficient watershed drainage
area, unexpected drought and, in the case of reservoirs, failure to close
the dam at the proper time. Failure to refill may have a great impact on
the aquatic biota, interrupting the life cycles of those species dependent
at some time upon littoral areas.
While drawdown brings about short-term control of most rooted species, some
species are strongly resistant to exposure and may even be stimulated by
it. Those species that are strongly resistant to drawdown and exposure
include Myriophyllum spicatum, Ceratophyll urn demersum, Lemna minor, Najas
flexilis, and Potamogeton pectinatus. Cook"e~( 1980a_) compiled the fol lowing
list of responses of some common nuisance aquatic macrophytes to drawdown:
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o Increased: Alternanthera philoxeroides (alligatorweed)
Najas fiexilis (naiad)
Potamogeton spp. (pondweed)
o Decreased: Chara vulgan's (muskgrass)
Eichorma crassipes (water hyacinth)
Nuphar spp. (water 1i 1 y)
o No clear response or change: Cabomba caroliniana (fanwort)
El odea canadensis (elodea)
Myriophyllum spp." (milfoil)
Utrlcularla vulgaris (bladderwort)
Information on the responses of 63 aquatic plants to drawdown is available
in Cooke (198(3a).
Additional negative effects of drawdown may include lowered levels in
potable water wells, and the loss of open water or access to open water for
recreation.
Implementation of Drawdown Projects
Lake drawdown should not be considered without first conducting a number of
laboratory and other investigations to determine the feasibility of the
technique. These investigations should include simulations of lake draw-
down, and laboratory studies of nutrient solubilization. Lake drawdown is
applicable only to lakes in which water input and output may be controlled.
The extent of macrophyte growth is important in specifying the depth to
which the lake level will be lowered.
Laboratory Experiments. Drawdown simulations are performed to determine
the extent to which sediments will dry and consolidate. Containers that
have been used in lake simulations range in size from Plexiglass tubes that
are 4.45 cm (ID) and 0.3 m high (Dooris, et al., 1982), to columns 0.3 m
(ID) and 1.2 m high (Fox, et al., 1977). Fox, et al. (1977) also used
plastic swimming pools (2.4 m in diameter, 45 cm deep) in lake simulation
experiments. The containers of sediment are exposed to air and light for a
period of time, during which sediment shrinkage and water loss are meas-
ured. The drying rate of the sediment can then be determined.
The container of dried sediment should be refilled, and the orthophosphate,
total phosphorus and total nitrogen levels measured. Ideally, only small
amounts of nitrogen and phosphorus compounds should be released from the
consolidated sediment. Large releases of nutrients may presage algal
blooms that may occur when the lake basin is refilled following drawdown.
Drawdown. The level of the lake should be lowered sufficiently to expose
most of the nuisance macrophytes, but to allow enough water for fish sur-
vival (if desired). It may be advantageous to combine drawdown with other
restoration techniques such as sediment removal and sediment covering.
Certain species of aquatic macrophytes may be more susceptible to drawdown
during one season than another. The decision to employ summer or winter
drawdown should be based upon the severity of the climate in a particular
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area, and upon consideration of lake uses and secondary management objec-
tives. For example, winter drawdown is advantageous because there will be
no invasion by terrestrial plants nor development of aquatic emergents, and
little interference with lake recreational uses. In addition, water bodies
drawn down in winter can usually be refilled in spring. In contrast, re-
filling in the autumn after a summer drawdown, may not be possible.
Complete dewatering of sediment is problematic during the winter, espe-
cially in regions of heavy snow or frequent winter rain. Winter drawdown
may also defeat other objectives such as the establishment of emergent
vegetation for waterfowl habitat, since these species may be susceptible to
the cold.
Lake Drawdown Case Studies
Lake level drawdown is a multipurpose improvement technique. The major
objective is generally to control the growth-of rooted aquatic vegetation,
with secondary objectives of fish management, sediment consolidation, and
turbidity control. The following case.histories exemplify the effects of
drawdown on lake biota.
Murphy Flowage, Wisconsin. Murphy Flowage (303 ha) was drawn down for two
consecutive winters in an effort to control the macrophyte species
Potamogeton robbinsii (Robbin's pondweed), Ceratophyllum demersum
(coontail), Nuphar sp. (water lily), Potamoqeton natansffToating-leaf
pondweed), and Myriophyl1 urn sp. (water milfoil). In 1967 and 1968, the
water level of the Flowage was lowered 1.5 m from November to March, and
restored in April. There was an 89 percent reduction in area covered by
macrophytes following the first drawdown, and an additional 3 percent
reduction occurred following the second drawdown. The species, that had
been dominant were controlled or nearly eliminated. No fish kills occurred
during drawdown. Following the second drawdown, resistant species such as
Megalondonta beckii (bur marigold), Najas flexilis (naiad), and Potamogeton
diversifolius (pondweed) began to spread.The extent to which resistant
species may have spread is unknown, because a flood destroyed the Flowage
in 1970 and evaluations were ended (Cooke, 1980aJ.
Blue Lake, Oregon. Blue Lake is an oxbow lake with a surface area of 26.3
ha, a maximum depth of 7.3 m, and a mean depth of 3.4 m. Prior to draw-
down, Eurasian water milfoil, Myriophyl 1 urn spicatum, dominated the littoral
areas of the lake. During the winter of 1981-1982, the lake level was
dropped 2.7 m to the base of most of the milfoil beds.
Drawdown reduced the standing crop biomass by 47 percent at depths less
than 1.2 m, and by 57 percent at depths from 2.4-3.7 m. The death of
shoots by drying and freezing during drawdown served to reduce milfoil
biomass. .However, drawdown alone did not eliminate the milfoil, and re-
growth from surviving rootcrowns was widespread. The herbicide 2,4'-D was
applied in 1982 to reduce milfoil growth.
Water quality effects that may be seen following reflooding include a
decrease in Secchi disc transparency and an increase in total suspended
solids, turbidity, chlorophyll-_a and total nitrogen and total phosphorus
concentrations (Geiger, 1983).
IV-3'3
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Additional In-Lake Treatment Techniques
Several additional methods of lake restoration are available, but have not
been applied as widely as the techniques noted in the previous sections.
The techniques that will be discussed in this section include dilution/
flushing, techniques to control nuisance aquatic vegetation (chemical
applications, harvesting, habitat manipulation and biological controls),
and liming of acidified water bodies.
Dilution/Flushing
Dilution/flushing improves lake water quality by reducing the concentration
of the limiting nutrient and increasing the water exchange rate in the
lake. The result is a reduction in the biomass of planktonic algae because
the loss rate exceeds algal growth rate. The technique is implemented by
adding low-nutrient water to the lake in order to reduce the concentration
of the limiting nutrient and thereby reduce algal growth. In addition,
nutrients and algal biomass are washed from the lake because the water
exchange rate is increased (Welch, 1979, 1981a, 1981b).
The purpose of dilution, as suggested earlier, is to deter blue-green algal
blooms by decreasing total phosphorus and total nitrogen, and by elimi-
nating biomass at a greater rate than the growth rate can supply new cells.
The reduction of allelopathic substances excreted by blue-green algae may
also contribute to the increased abundance of diatoms and green algae
(Welch and Tomasek, 1980).
Use of the dilution/flushing method is most feasible when large quantities
of low-nutrient water are available for transport to the lake that is to be
restored. This condition was met in the instances of Moses and Green Lakes
in Washington State. Case histories of these two lakes are discussed
below.
Moses Lake, Washington. Moses Lake has an area of 2,753 ha and a mean
depth of 5.6 m. Prior to restoration by dilution/flushing, the lake was
eutrophic and experienced blue-green algal blooms because of high nutrient
concentrations. Inflowing water (Crab Creek, [P]=92 ug/1) was diluted with
low nutrient water from the Columbia River ([P]=30 ug/1) with about a 3:1
dilution of Crab Creek. Following dilution/flushing, Secchi disc depth in
the lake increased from 0.5 m to 1.1 m (April-July values). Total phos-
phorus, which had a mean value of 142 ug/1 prior to dilution, was reduced
to 53 ug/1. Chlorophyll-a also decreased from 55 ug/1 (mean values for
April-July) to 9 ug/1 (April-July mean).
Green Lake, Washington. Green Lake, which is located in King County,
Washington State, has a surface area of 104 ha, a mean depth of 3.8 m, and
a maximum depth of 8.8 m. Prior to dilution, Green Lake had a high level
of blue-green algal production, and high nutrient levels caused by sub-
surface seepage (U.S. EPA, 1982).
Dilution began in 1962 with the Seattle city water supply as the source of
low nutrient water. The technique applied to Green Lake was one of long-
term dilution at a relatively low rate. Post-dilution monitoring did not
begin until three years after dilution was begun, and only one pre-dilution
IV-34
-------
measurement was made. The data available showed that Secchi disc depth
increased from 1 m to 4 m and chlorophyll-a decreased over 90 percent (from
45 ug/1 to 20 ug/1). Total phosphorus in the lake water declined from a
summer mean of 65 ug/1 to 20 ug/1 (Welch, 1979, 1981a, 1981bj U.S. EPA,
1982).
Control of Nuisance Aquatic Vegetation
Management practices for the control of aquatic weeds include chemical con-
trol, mechanical control (dredging and harvesting), habitat manipulation
(use of shades, dyes, bottom coverings, lake drawdown) and biological con-
trol (fish, shellfish, insects, disease, competitive plants).
Chemical Control. Aquatic weeds can be controlled by a variety of chemi-
cals,including 2,4-0, Oiquat, Endothal, Simazine, Fenac, Dichlobenil,
Floridone, acrolein, and copper compounds. Combinations of Diquat and
copper sulfate (CuSO,,) and Endothal and copper sulfate have been shown to
be effective for weea control, using lower concentrations of. herbicide than
that required for the herbicide alone (Nichols and Shaw, 1983). Herbicides
are most effective in water with low turbidity, at water temperatures of
158C to 18°C, and on young plants. Effectiveness is also increased in
waters with high calcium concentrations, and when herbicides are applied
before weeds develop seeds.
Harvesting. Harvesting is commonly practiced in the Northeast, Upper
Midwest, and West coastal regions to control aquatic weeds. The efficacy
of harvesting depends upon the biology of the particular species. For
example, more than one harvest is needed to control milfoil regrowth over
the growing season. The major positive effects of harvesting are (Nichols
and Shaw, 1983):
o Organic material removed by harvesting is no longer available to
deplete oxygen supplies upon decay;
o Nutrients are not available for recycling upon decay of the plant;
and
o Foreign material of a chemical or biological nature is not being
introduced into the system.
The negative impacts include (Nichols and Shaw, 1983):
o Temporary increase in turbidity;
o Loss of animal habitat;
o Potential of plant spread by vegetative means;
o Increased growth following removal of canopy;
o Harvesting of animal material;
o Release of nutrients from cut stumps.
XY-35
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Habitat Manipulation. Dredging may be used to mechanically remove the
whole plant from shallow waters, or it may be used to increase the depth to
a point below which plants are unable to grow. Dredging may also remove
sediment nutrient sources for aquatic plant growth.
Shades, dyes, bottom coverings and drawdown are also included in habitat
manipulation techniques to control aquatic weeds. Black plastic sheeting
that floats on the water surface has reportedly controlled growth of
Myri ophyl1 urn spicatum (Nichols and Shaw, 1983). Following four weeks of
shading, £h~e plants were brown and dead, and there was little or no re-
growth during the rest of the summer. Cooke (1980j>) reviewed the various
methods that are encompassed by the general category of covering bottom
sediments. Included within these techniques are sheeting and screening,
and smothering with sand or fly ash. Cooke (1980JD) concluded:
o Plastic sheeting appears to be effective in retarding macrophyte
growth, but there are problems with application methods and in
anchoring the material;
o Fiberglass screens hold promise as effective means of controlling
macrophytes, but further evaluation is recommended;
o Sand is apparently not effective if enriched sediment is not first
removed because the sand particles sink into flocculent sediments;
and ,
o Fly ash was not recommended because of the negative water quality
effects (elevated pH, low dissolved oxygen, high concentrations of
heavy metals) and subsequent effects on the biota.
The aniline dye nigrosine has been used in attempts to control macrophytes.
Although the toxicity of aniline dyes to other organisms is not known, they
are very toxic to humans. Other considerations associated with the use of
dyes include aesthetics, loss of effect through dilution, loss of dye
through plant uptake and loss by sorption to suspended solids and sediment.
Biological Controls. Biological controls include the use of fish, shell-
fish, insects, and disease. Some fish that have been suggested for control
of aquatic weeds are the common carp (Cyprinus carpio), roach (Rutilus
rutilus), rudd (Scardinus erythopthalmus), some species of tilapia (Tilapia
zillii, T. mossambica), silver dollar fish (Metynnis roosevelti, Mylossoma
argenteunT), white amur (Ctenopharyngodon idell a) and hybrids of the white
amur (Mulligan, 1969; Nichols and Shaw, 198TTIt should be noted that the
introduction of exotic species is strictly regulated in many states.
Carp are not primarily herbivores, but they serve to decrease plant growth
by uprooting plants when searching for benthic organisms or when spawning,
and by increasing turbidity in the water. Although carp have been shown to
effectively control el odea and curly-leaved pondweed, they cause water
quality problems (suspended sediment, turbidity) which can lead to the
demise of sportfish populations (Nichols and Shaw, 1983).
Herbivorous fish can be used to control certain species of aquatic weeds.
For example, roach and rudd prefer elodea over milfoil. Milfoil is also
IY-36
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the least preferred food of Tilapia spp. The introduction of grass carp at
Red Haw Lake, Iowa, resulted in control of El odea, Potamogeton, Cerato-
phylluin and Najas. The biomass of aquatic, macrophytes in the lake
decreased from 2,438 g/m in 1973 to 211 g/nr in 1976 (Mitzner, 1978).
Since milfoil is not the preferred food of herbivorous fish, there is a
possibility that persistent monocultures of Myriophyllum spicatum will
develop.
Herbivorous snails have been suggested as potential controls for macro-
phytes. Although native snail species in temperate regions do not eat
macrophytes, two South American species (Marisa cornuarielisi L. and
Pomacea australialis) are macrophyte herbivores that may potentially be
used to control pest species. The crayfish Orconectes causeyi, which
consumes both El odea canadensis and Myriophy11 urn exalbeseems, has also been
suggested as a means of biologicalcontrolo? macrophytes (Nichols and
Shaw, 1983). . •
Several insects have also been investigated as predators on Eurasian water
milfoil. Some of the promising species noted are Parapoynx stratiota, P.
allionealis, Acentria nivea, Litodactylus leucogasteran~d all aquatic
moths. However, most of these insects are not specific to milfoil. Dis-
eases that may cause declines in milfoil populations include "Lake Venice"
disease and "Northeast" disease. The causes of these two diseases are not
known nor are the long-term consequences of artificial introduction of
disease. Thus, the use of pathogens to control milfoil is not recommended
(Nichols and Shaw, 1983).
Neutralization of Acidified Lakes
Causes of Acidity and Problem Definition. Acidity of surface waters is
largely caused by two nonpoint sources: acid mine drainage and acid
precipitation. Acid mine drainage results when mine water comes in contact
witn sulfur-containing minerals. Acid precipitation is caused by atmos-
pheric sulfur that is released by electric utilities and urban and in-
dustrial operations that use sulfur-containing fuel. Oxidation+of sulfuric
compounds produces sulfuric acid, which dissociates to form H and SO.
ions in surface or atmospheric water (Novotny and Chesters, 1981).
Acid mine drainage and acid precipitation cause undesirable "oligo-
trophication" (a severe loss of productivity caused by the low pH condi-
tions), including loss of natural fish populations. Salmonid fisheries,
particularly lake trout, are susceptible to acidification (Goodchild and
Hamilton, 1983).
The ability of surface waters to neutralize acidic inputs is largely a
function of the chemical composition and solubility of the surrounding
soils and underlying rocks. For example, limestones (CaCO,) and dolomites
(CaMg(C03)2) yield infinite acid neutralizing capacity, whereas hard rocks
such as granites (i.e., quartz - Si02, feldspar - KAISi^Og) and related
igneous rocks, crystalline metamorphi
-------
and Olem, 1983). Areas of the United States where lakes are highly sensi-
tive to acidification are in New England, the Adirondack Mountains of New
York, the Appalachians, and the Rockies.
Neutralization. Several materials have been considered for use in neu-
tralizing acid lakes. These include lime (CaO, Ca(OH)2), limestone
(CaC03), dolomite, lime slags, basic flyash, soda ash, and priosphorus. Of
these; lime and limestone are the most widely employed to neutralize sur-
face waters (Driscoll, et al., 1982). Dolomite, dolomitic hydrated lime,
and dolomite quicklime (each exceeding a 35 percent magnesium content) may
also be used. However, limestones containing more than 10 percent mag-
nesium carbonate dissolve slowly and are not practical for use in neutral-
izing surface waters. Agricultural limestone, while not as effective as
quicklime or hydrated lime, has several advantages: it is noncaustic,
relatively inexpensive, relatively free of harmful contaminants', and does
not produce harmful alkaline conditions (Britt and Fraser, 1983).
Application. Techniques for lime application in lakes include using trucks
(blowers), boats (blowers, slurries, bags), aircraft, and sediment injec-
tion systems. The proper time and place to apply neutralizing agents
depends upon two main factors: the time and location of acidic episodic
events (e.g., snowmelt, autumnal rains); and relationships between such
events and the critical life stages of aquatic biota. For example, in
dimictic lakes, mixing and distribution of lime is enhanced when it is
applied during the spring overturn. However, spring acidic snowmelt
creates two problems. First, neutralization may occur too late to prevent
fish embryo and fry mortality that is caused by acidic snowmelt. Second,
the colder snowmelt water may be less dense than deeper lake water, and
mixing with neutralized water may be inhibited (Britt and Fraser, 1983).
Liming the entire lake area is desirable, but may not be feasible because
of time and other resource constraints. Alternatively, application of lime
over the deepest part of the lake allows the particles of CaC03 more time
to react within the water column. Another alternative may be to distribute
limestone in shallow littoral zones where wave action enhances dissolution
(Britt and Fraser, 1983). An alternative liming strategy involves
chemically treating watersheds, thereby neutralizing the associated aquatic
ecosystem. Methods to estimate lime requirements are found in Boyd (1982)
and Driscoll, et al. (1982).
Liming Effects. The biological consequences of liming have been summarized
by Hultberg and Andersson (1982) and Britt and Fraser (1983). Case histo-
ries of limed lakes show the following changes in lake biota:
o Decreases in acidophilic algae and mosses, with concurrent in-
creases in diversity of planktonic algae;
o Predominance of cladocerans shifts to a predominance of copepods
after neutralization;
o Reduction in benthic biomass after liming, but eventual recovery
with repopulation of less acid tolerant species;
IY-38
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o Most fish species respond positively, with enhanced .survival due
to successful spawning and hatching.
Some chemical changes caused by neutralization may be of concern. Toxicity
changes of metals, especially aluminum, may have serious environmental con-
sequences. Aluminum toxicity varies with pH changes; gill damage to fish
may be caused when aluminum reacts with hydroxides from pH 4.4 to 5.2,
while other studies indicate that aluminum is most toxic to fish from pH
5.2 to 5.4 (Britt and Fraser, 1983). The sediments of a limed lake may
become sinks for aluminum and other toxic metals as pH is raised and the
metals are removed from the water column. If the lake is allowed to re-
acidify after several years of treatment, the remobilization of metals may
cause serious biological problems.
Watershed Management
The quality of a lake's water is often a direct manifestation of the number
and types of pollution sources in the surrounding watershed. Agricultural
practices such as tillage, the use of fertilizers, and operations of con-
fined animal feedlots may potentially increase the loss of sediments and
nutrients from the land and accelerate the natural process of lake
eutrophication. In urban areas, many pollutants are carried to lakes in
stormwater runoff, via combined sewers, storm sewers and direct surface
runoff.
The effectiveness of in-lake restoration techniques would be short-lived if
the cause of eutrophication (high nutrient input) was not corrected.
Watershed pollution control techniques are important corrective and often
preventive measures. The following sections highlight watershed management
techniques that help control nonpoint sources of pollution from agricul-
tural and urban areas.
Agricultural Pollution Control
Control of Sediment Input and Associated Nutrients. One of the most impor-
tant water pollutants that resultsfrom agricultural activities is the
sediment input from eroding croplands. Sediment itself is a physical pol-
lutant, and in addition serves as a vehicle to transport nutrients,
pesticides, toxic chemicals, organic matter, and inorganic matter to water
bodies. Techniques to reduce soil loss from agricultural lands have been
discussed in the U.S. Environmental Protection Agency publication entitled
Effectiveness of Soil and Water Conservation Practices for Pollution
Control (1979b) and in a publication by Stewart, et al.(1975).Several
Soil and Water Conservation Practices (SWCP) will be discussed in the
following paragraphs.
No-Till Planting. Planting is accomplished by placing seeds in the soil
without tillage, using a fluted coulter that leaves the vegetative cover
virtually undisturbed. Chemical herbicides are used to control weeds and
previously planted crops. No-till planting can reduce soil loss to less
than 5 percent as compared to conventional plowing and planting practices
(Novotny and Chesters, 1981). However, this method requires a greater use
of herbicides, and lower yields may be expected on some soils. Because
vegetative cover is left to decompose on the surface, the loss of soluble
IY-39
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plant nutrients is greater in runoff from no-till than from conventionally-
tilled plots (U.S. EPA, 1982).
In summary, no-till farming reduces runoff and erosion losses. Therefore,
losses of strongly adsorbed and solid phase pollutants (total phosphorus
and organic nitrogen) are decreased. Losses of weakly adsorbed pesticides
and plant nutrients (dissolved phosphorus) may increase; but overall the
no-till technique is effective in reducing losses of both phosphorus and
nitrogen.
Conservation Tillage. This technique replaces conventional plowing with a
form of noninversion tillage that retains some of the plant residue on the
surface. A chisel, field cultivator, or disk .can be used for tilling. The
organic residue cover protects the soil surface from erosion and decreases
the volume and velocity of runoff (U.S. EPA, 1979bj. Because runoff volume
and soil loss are reduced, losses of strongly adsorbed organic phosphorus,
organic nitrogen and insecticides are decreased.
Sod-Based Rotations. This system involves the periodic rotation of row
crops and a sod crop such as alfalfa, other legumes, or grasses. Plowing
the sod improves filtration and reduces credibility. Increased soil
porosity helps decrease surface runoff, and the reduction in runoff can
continue for several years of continuous row crops after the sod crop is
plowed under (U.S. EPA, 1982).
An additional benefit of sod-based rotations is that crop rotations lessen
the need for applications of fertilizers and pesticides by increasing soil
organic matter and species diversity. Also, legumes help restore nitrogen
to soils through fixation of atmospheric nitrogen.
Cover Crops. Shredded stalks of corn or sorghum can be left on fields
during the non-growing season, thereby reducing runoff and soil loss from
normally fallow fields. More protection from surface runoff is provided
from the cover crop that is left in place than by late-seeded small-grain
winter cover on plowed fields (Novotny and Chesters, 1981).
Terraces. Terraces divide the field into segments with lesser or near-
horizontal slopes, thereby reducing the slope effect on erosion rates.
Generally, terraces consist of an embankment or a combination of an embank-
ment and a channel that diverts or stores surface runoff.
Terraces are more effective in reducing erosion than in decreasing surface
runoff. Consequently, terraces are most effective in reducing strongly
adsorbed substances such as total phosphorus and paraquat (Smith, et al.,
1979). Impoundment terraces, which retain runoff in surface storage areas,
reduce both runoff volume and sediment loss, but the eventual percolation
of the stored water may increase the nitrogen loading to the groundwater.
Other Methods to Prevent Sediment and Nutrient Losses. Contouring, ridge
planting, contour listing, and strip cropping are methods that are designed
to create barriers perpendicular to the natural direction of flow. Runoff
volume and water velocity are thus decreased. In the technique of contour
plowing, crop rows and plowing follow the natural contour of the land.
This practice provides excellent erosion control for moderate rainstorms
IY-40
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(Novotny and Chester*, 1981). Ridge planting involves planting crops on
preformed ridges that follow the natural contours of the field. Crop
residues are pushed into the furrows between rows, further deterring runoff
and erosion (U.S. EPA, 1982).
A special plow (lister) is required to form alternating ridges and furrows
for contour listing. Row crops are then planted either in the bottom
furrows or the ridge tops. Contour strip cropping is accomplished by
alternating the cultivated crops with strips- of grass or close growing
crops.
The principal erosion control practices for use on croplands are summarized
in Table IY-4.
Waste Management Planning. The p.lanning of a waste management system helps
prevent the owner from investing in unnecessary components. Evaluations
include estimations of liquid and solid waste sources on a farm and devel-
opment of a complete system to manage them without degrading' air, soil or
water resources. An operation plan, which provides specific details for
operation of the system, should include:
1. Timing, rates, volumes, and locations for applications of waste
and, if appropriate, approximate number of trips for hauling
equipment and an estimate of the time required..
2. Minimum and maximum operation levels for storage and treatment
practices and other operations specific to the practice, such as
estimated frequency of solids removal.
3. Safety warnings, particularly where there is danger of drowning or
exposure to poisonous or explosive gases.
4. Maintenance requirements for each of the practices.
Waste Storage Ponds. The purpose of waste storage ponds is to temporarily
store liquid and solid wastes, wastewater, and polluted runoff until it can
be applied to land without polluting surface or ground water. Common uses
of waste storage ponds are storage of milkhouse wastes and manure and
storage of polluted runoff from feedlots and barnyards.
Diversions or dikes are usually combined with systems employing waste
storage ponds. Clear water diversion systems direct water from upland
watersheds away from feedlots or barnyards. Polluted runoff may be
collected and directed to storage ponds by constructing a system of curbs,
gutters or terraces. Design of waste storage ponds should consider the
maximum period of time between emptying, which varies according to
precipitation, runoff, and waste volume.
Waste Storage Structures. Waste storage structures such as storage tanks
and manure stacking facilities serve the same purposes as waste storage
ponds, and while storage structures are more expensive they offer several
advantages. Advantages include preservation of nutrient content of stored
wastes, minimization of odors, management flexibility and improved
aesthetics.
IY-41
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TABLE IV-4
PRINCIPAL TYPES OF CROPLAND EROSION CONTROL PRACTICES AND THEIR HIGHLIGHTS (Continued)
£9 Contouring
EIO Giaded rows
Ell Contour strip cropping
£12 Terraces
i
-r=»
r\i
£13 Grassed outlets
£14 Kidge planting
EIS Contour listing
£16 Change in land use
£17 Olhei practices
Can reduce average soil loss by 50% on moderate slopes, but less on sleep slopes; loses effectiveness
if rows break over; must be supported by terraces on long slopes; soil, climatic, and topographic
limitations; not compatible with use of large farming equipment on many topographies. Docs not
affect fertilizer and pesticide rales.
Similar to contouring but less susceptible to tow breakovers.
Kowcrop and hay in alternate SO- lo 100-fl strips reduce soil loss to about 50% of thai with the same
rotation contoured only; fall seeded grain in lieu of meadow about half as effective; alternating corn
and spring grain not'effective; area must be suitable for across-slope farming and establishment of
rotation meadows; favorable and unfavorable features similar to £3 and £9.
Support contouring and agronomic practices by reducing effective slope length and runoff concentra-
tion; reduce erosion and conserve soil moisture; facilitate more intensive cropping; conventional
gradient terraces often incompatible with use of large equipment, but new designs have alleviated this
problem; substantial initial cost and some maintenance costs.
I actfilale drainage of graded rows and terrace channels with minimal erosion; involve establishment
and maintenance costs and may interfere with use of large implements.
Earlier warming and drying of row zone; reduces erosion by concentrating runoff How in mulch-
covered furrows; most effective when rows are across slope.
Minimizes row breakover; can reduce annual soil loss by 50%; loses effectiveness with poslernergence
corn cultivation; disadvantages same as £9.
Sometimes the only solution. Well managed permanent grass or woodland effective where other
control practices arc inadequate, lost acreage can be compensated for by more intensive use of less
erodible land.
Contour furrows, diversions, subsurface drainage; land forming, closer row spacing, etc.
SOURCE: Stewart, et al., 1975
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TABLE IV-4
PRINCIPAL TYPES OF CROPLAND EROSION CONTROL PRACTICES AND THEIR HIGHLIGHTS
CO
1:2
ftuosion Control Practice
Ucnuflls and Impact
No III! plant in prior-flop residues
Conservation tillage
1:3 Sod-hascd rotations
1:4 Meadowless rotations
1:5 Winlci cover crops
1:6 Improved soil I'uiiility
17 Timing of field operations
I'low-planl systems
Most effective in dormant grass or small grain; highly effective in crop residues; ininirni/es spring
sediment surges and provides year-round control; reduces man. machine, and fuel requirements;
delays soil warming and drying; requires more pesticides and nitrogen; limits fertilizer- and pesticide
placement options; some climatic and soil restrictions.
Includes a variety of no-plow systems that retain some of the residues on the surface; more widely
adaptable Inil somewhat less effective than I I; advantages and disadvantages generally same as I I
l>ul to lesser degree.
(iood meadows lose virtually no soil and reduce erosion from succeeding crops; total soil loss greallv
reduced hut losses unequally distributed over rotation cycle; aid in control of some diseases and
jicsls; more fertilizer-placement options; less realized income from hay years; greater potential trans-
port of water-soluble I*, some climatic reslricltons.
Aid in disease and pest control; may provide more continuous soil protection than one-crop systems;
much less effective than Ii3.
Keduce winter erosion where corn stover has been removed and after low-residue crops; provide good
base lor slot-planting next crop; usually no advantage over heavy cover ol chop|ied stalks or straw;
may reduce leaching of niliale; water use by winter cover may reduce yield of cash crop.
Can substantially reduce ciosion hazards as well us inciease crop yields
I all plowing facilitates more timely planting in wet spiings, but it greatly increases winter and early
spring eiosiun hazards; optimum liming of spiing operations can reduce erosion and increase yields.
Uough. ilnddy surface increases infiltration and reduce;, erosion; much less effective than II and 1:2
when long rain periods occur; seedling stands may be pool when moisture conditions are less than
optimum. Mulch effect is lost by plowing.
-------
Waste Treatment Lagoons. Treatment lagoons may be designed as anaerobic,
aerobic, or aerated lagoons. They are used principally to treat liquid
wastes.
Anaerobic lagoons are the most commonly used. They require less area than
aerobic lagoons, and do not need require electricity for operation, as do
aerated systems. Treated wastes may be lower in nitrogen due to ammonia
volatilization; therefore, the waste may be applied over a smaller land
area.
Aerobic lagoons are used for weak agricultural wastes, such as those
originating from milk centers. They require large surface areas, and the
effluent is rarely suitable for discharge to surface water.
Filter Strips. In.this method, runoff from feedlots and barnyards flows
over grassy strips. The strips help reduce the volume and pollution
content by soil percolation, the filtration capability of the grass, and
volatilization.
Waste Utilization. Waste utilization refers to where and when manure
should be applied to land. Its purpose is to use the wastes as fertilizer
for crops, forage and fiber production, to prevent erosion, to improve or
maintain soil structure, to produce energy, and to safeguard water
resources.
Factors to be considered include the land areas available, and the crops
that will be grown. Other factors that should be considered are the timing
of application, nutrient release rates, soil types, and climate.
Urban Runoff Pollution Control
Lakes in urban areas are subject to pollution from stormwater runoff which
enters lakes via combined sewers, storm sewers, and direct surface runoff.
The runoff contains high concentrations of sediment, nutrients, heavy
metals and toxic chemicals.
During storm events, the capacity of combined sewer lines may be exceeded,
and overflow structures at sewage treatment plants or in the sewerage
system are designed to discharge the excess into surface water bodies. The
"first flush effect" refers to the phenomenon in combined sewer overflow
samples whereby the highest concentrations of BODS, suspended solids,
grease and other pollutants are found during the earliest part of a storm
event. Accumulated solid deposits that contain organic matter undergoing
decay in combined, sanitary and storm sewers may increase BODr concen-
trations to levels greater than those of normal untreated dry-weather
wastewater (Lager and Smith, 1974). Long periods between rainfall, low
sewer slopes, infrequent cleaning, and failure to block off or clean catch
basins magnify pollutant concentrations in combined sewer overflows, and
(to a lesser extent) storm sewer discharges.
Several management alternatives are available to alleviate problems caused
by urban stormwater. Techniques may be grouped into three categories:
land management, collection system modifications, and storage. While
descriptions of urban runoff control measures are beyond the scope
IY-44
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of this manual, several components of each category will be briefly sum-
marized in the following paragraphs.
Land Management. Land management practices include those measures designed
to reduce urban and construction site stormwater runoff at the source, by
employing Best Management Practices (BMPs). Qn-site measures can be
further divided into low structural or non-structural controls.
Low structural control measures require physical modifications in a
construction or urbanizing area. The most common on-site control is
storage. Storage attenuates peak runoff flows, treats runoff (detention/
sedimentation), or contains the flow in combination with another treatment
process such as retention/percolation (Lynard, et al., 1980).
Non-structural control measures include surface sanitation, chemical use
control, use of natural drainage, and certain erosion/sedimentation control
practices (Field, et al., 1977). Surface sanitation (street sweeping
operations) may have a significant impact on the quantity of pollutants
washed off by stormwater. Certain street cleaning techniques are able to
remove 93 percent of the dry weight solids, which make up a significant
portion of the overall pollution potential (Field, et al., 1977; Lager and
Smith, 1974). A frequently overlooked measure for reducing the pollution
potential from urban areas is reduction in the use of fertilizers, pesti-
cides and deicing materials. Suggestions for methods to reduce such inputs
can be found in Lager and Smith (1974) and Field, et al. (1977).
Construction in urbanized areas replaces areas of natural infiltration and
drainage wfth impervious areas. The result is increased runoff and
fiowrates, and decreased infiltration to the groundwater. Use of natural
drainage helps reduce drainage costs and pollution, while it enhances
groundwater supplies and flood protection (Field, et al., 1977).
Non-structural erosion/sedimentation controls include cropping (seeding and
sodding), use of mulch blankets, nettings, chemical soil stabilizers and
earthen berms. These measures are described in Lager and Smith (1974),
Field, et al. (1977), and Lynard, et al. (1980).
Col 1ection System Controls. Collection system controls include sewer
separation,inflow control, flushing and polymer injections, regulators,
and remote flow monitoring and control. Several of these alternatives are
briefly described below.
Sewer Separation. Sewer separation refers to the conversion of a combined
sewer system into separate sanitary and storm sewer systems. The practice
of sewer separation has been used for many years, but Lager and Smith
(1974) note two main reasons for Devaluating sewer separation. The first
reason stems from changes in physical conditions and quality standards from
the past, which include: (1) increases in urban impervious areas and
municipal water usage, causing overflows of increased duration and quan-
tity; (2) rapid industrial expansion, causing increased quantities 'of
industrial wastewaters in the overflows; (3) increasing environmental
concern for better water quality; and (4) the realization that the total
amount of available fresh water is limited and that complete reclamation of
substantial portions of the flow may be necessary in the future. The
IY-45
-------
second reason includes: (1) separated storm sewer discharges contain pol-
lutants that affect the receiving water and create new problems; and (2)
storm sewer discharges occur more frequently and last longer than combined
sewer overflows because combined sewer regulators prevent overflows during
minor events.
Lager and Smith (1974) concluded that in many cases the separation of
existing combined sewer systems is not practically or economically feasible
to resolve combined sewer problems. A feasibility study including, the cost
of alternative methods would indicate the practicality of each option.
Infiltration/Inflow Control. Problems result from infiltration into sewers
from groundwater sources, and high inflow rates through direct connections
from sources other than those which the sewers .are intended to serve.
Examples of infiltration are the volumes of water that enter the sewer
system through manhole walls, cracks, defective joints, and illegal
connections.
Remote Flow Monitoring and Control. Computerized collection system control
can be applied to upgrade combined sewer systems. Control systems are
intended to assist in routing .and storing combined sewer flows to
effectively use interceptor and line capacities (Lager and Smith, 1974).
The control system is able to sense and report minute-torminute system
status, including flow levels, quantities, treatment rates, pumping rates,
gate (regulator) positions, and characteristics at significant locations in
the system. Such observations may assist in determining where necessary
overflows can be discharged with the least impact. The control system also
provides a means for manipulating the system to maximum advantage.
Storage. Storage of runoff effectively prevents or reduces stormwater
runoff from entry into combined sewers and surface water bodies. Storage
facilities can provide complete or short-term retention of stormwater
flows. Retention facilities may incorporate infiltration systems such as
gravel bottoms or tile drains.
Detention basins are capable of reducing peak flow volumes from storms, and
providing a sediment trap for suspended solids. The gradual release of
stormwater lessens impacts caused by flooding, erosion, and disruption of
aquatic habitats (U.S. EPA, 1982).
Stormwater flows to treatment plants, and subsequent overflows, may be
controlled by in-line or off-line storage facilities. Storage facilities
have several advantages: they are basically simple in design and opera-
tion, they respond without difficulty to intermittent and random storm
behavior, they are relatively unaffected by flow and quality changes, and
they are capable of providing flow equalization (Lager and Smith, 1974).
Drawbacks of storage basins include their large size (real estate require-
ments and therefore cost), visual impact and the need to provide for solids
dewatering and disposal.
Storage facilities may be in-line, in which regulators and pumping stations
are used to store stormwater runoff in areas of the sewer system with extra
capacity, or off-line, which may be concrete vaults, or storage basins such
IV-46
-------
as described earlier. Detailed information concerning storage facilities
is available in Lager and Smith (1974), Field (1977), and Lynard, et al.
(1980).
IV-47
-------
CHAPTER Y
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V-15
-------
APPENDIX A
PALMER'S LISTS OF POLLUTION TOLERANT ALGAE
Source: Palmer, 1969
A-l
-------
APPENDIX A
PALMER'S LISTS OF POLLUTION TOLERANT ALGAE
TABLE A-l
POLLUTION-TOLERANT GENERA OF ALGAE
LIST OF THE 60 MOST TOLERANT GENERA,
IN ORDER OF DECREASING EMPHASIS BY 165 AUTHORITIES
No.
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
. 31
32
33
34
35
36
37
38
Genus
Euglena
Oscillatoria
Chlamydomonas
Scenedesmus
Chlorella
Nitzchia
Navicula
Stigeoc Ionium
Snynedra
Ankistrodesmus
Phacus
Phornridium
Melosira
Gomphonema
Cyclotella
Closterium
Micractinium
Pandorina
Anacystis
Lepocinclis
Spirogyra
Anabaena
Cryptomonas
Pediastrum
Arthrospira
Trachelomonas
Carteria
Chlorogoniura
Fragilaria
Ulothrix
Surlrella
Stephanodiscus
Eudorina
Lyngbya
Oocysti s
Agnjenellum
Splrulina
Pyrobotrys
Group3
F
8
F
G
G
D
D
G
D
G
F
B
D
0
D
G
G
F
B
F
G
B
F
G
B
F
F
F
D
G
D
D
F
B
G
B
B
F
No.
authors
97
93
68 '
70
60
58
61
50
44
36
39
37
37
35
35
34
27
32
28
25
26
27
27
28
18
26
21
23
24
25
27
22
23
17
20
19
17
16
Total .
Points
172
161
115
112
103
98
92
69
58
57
57
52
51
48
47
45
44
42
39
38
37
36
36
35
34
34
33
33
33
33
33
32
30
28
28
27
25
24
A-2
-------
TABLE A-l (CONTINUED)
No.
39
40
41
42
43
44
45
46
47
48
49
50
51
52
53
54
55
56
57
53
59
60
Genus
Cymbella
Ac tina strum
Coelastruni
Cladophora
Hantzschia
Diatoma
Spondylomorum
Golenkinia
Achnanthes
Synura
Pinnularia
Chlorococcum
Asterionella
Cocconeis
Cosmarium
Gonium
Tribonema
Stauroneis
Selenastrum
Dlctyosphaerium
Cymatopleura
Crucigenia
Group3
D
G
G
G
D
D
F
G
D
F
0
G
D
0
G
F
G
D
G
G '
D
G
No.
authors
19
20
21
22
18
19
16
14
16
14
15
13
14
14
14
15
10
14
13
11
13
13
Total
Points
24
24
24
24
23
22
21
19
19
18
18
17
17
17
17
17
16
16
15
14
14
14
Groups: B, blue-green; D, diatom; F, flagellate; G, green,
SOURCE: Palmer, 1969.
A-3
-------
TABLE A-2
POLLUTION-TOLERANT GENERA OF ALGAE
LIST OF THE 80 MOST TOLERANT SPECIES,
IN ORDER OF DECREASING EMPHASIS BY 165 AUTHORITIES
No.
1
2
3
4
5
6
7
8
9
10
11 •
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
35
36
37
38
39
40
41
42
Genus
Euglena viridis
Nitzschia palea
Oscillatoria limosa
Scenedesmus quadricauda
Oscillatoria tenuis
Stigeoc Ionium tenue
Synedra ulna
Ankistrodesmus falcatus
Pandorina morum
Oscillatoria chlorina
Chlorella vulgaris
Arthrospira jenneri
Melosira varians
Cyclotella meneghiniana
Euglena gracilis
Nitzschia acicularis
Navicula cryptocephala
Oscillatoria princeps
Osclllatoria putrida
Gomphonema parvulum
Hantzschia amphioxys
Oscillatoria chalybea
Stephanodiscus hantzschii
Euglena oxyuris
Closterium acerosum
Scenedesmus obliquus
Chlorella pyrenoidosa
Cryptomonas erosa
Eudorina elegans
Euglena acus
Surirella ovata
Lepocinclis ovum
Oscillatoria formosa
Oscillator! a splendida
Phacus pyrum
Micractinium pusillum
Agmenellum quadriduplicatum
Melosira granulata
Pediastrum boryanum
Diatoma vulgare
Lepocinclis texta
Euglena deses
Group3
F
D
B
G
B
G
D
G
F
B
G
B
D
D
F
D
D
B
B
0
D
B
D
F
G
G
G
F
F
F
D
F
B
B
F
G
B
D
G
0
F
F
No.
authors
50
45
29
26
26
25
25
21
23
17
19
15
22
20
18
18
19
16
13
14
18
14
16
15
16
16
11
15
16
16
16
14
14
14
11
12
13
14
15
17
12
13
Total
Points
93
69
42
41
40
34
33
32
30
29
29
28
28
27
26
26
25
24
23
23
23
22
22
21
21
21
20
20
20
20
20
19
19
19
18
18
18
18
18
18
17
17
A-4
-------
TABLE A-2 (CONTINUED)
NO.
43
44
45
46
47
48
49
50
51
52
53
54
55
55
57
58
59
60
61
62 .
63
64
65
66
67
68
69
70
71
72
73
74
75
76
77
78
79
80
Genus /
Spondylomorum quaternarium
Phormidium uncinatum
Cnlamydomonas reinhardii
Chlorogonium euchlorum
Euglena polymorpha
Phacus pleuronectes
Navicula viridula
Phormidiurn autumnal e
Oscillator! a lauterborm'i
Anabaena constn'cta
Euglena pisciformis
Actinastrum hantzschii
Synedra acus
Chlorogonium elongatum
Synura uvella
Cocconeis placentula
Mitzschia sigmoidea
Coelastrum microporum
Achnanthes minutissima
Cymatopleura solea
Scenedesmus dimorphus
Fragilaria crotonensis
Anacystis cyanea
Navicula cuspidata
Scenedesmus acumi natus
Euglena intermedia
Pediastrum duplex
Closterium leibleinii
Oscillatoria brevis
Trachelomonas volvocina
Dictyosphaerium pulchellum
Fragilaria capucina
Cladophora glomerata
Cryptomonas ovata
Gonium pectorale
Euglena proxima
Py robotrys graci 1 i s
Tetraedron muticuin
Group3
F
8
F
F
F
F
D
B
B
3
F
G
D
F
F
0
D
G
0 .
0
G
0
B
0
G
F
G
G
B
F '
G
0
/»
a
F
F
F
F
G
No.
authors
13
• 15
10
10
11
11
13
13
3
9
11
13
9
10
11
12
12
13
10
12
8
9
10
10
10
11
11
8
8
8
9
9
10
10
10
7
7
7
Total
Points
17
17
16
16
16
16
16
16
15
15
15
15
14
14
14
14
14
. 14
13
13
12
12
12
12
12
12
12
11
11
11
11
11
11
11
11
10
10
10
aGroups: B, blue-green; D, diatom; F, flagellate; G, green.
SOURCE: Palmer, 1969.
A-5
-------
APPENDIX B
U.S..ENVIRONMENTAL PROTECTION AGENCY'S PHYTOPLANKTON TROPIC INDICES
Source: U.S. EPA, 1979 a.
8-1
-------
All qenus-trophic-values used in formulating the phytoplankton trophic
indices are presented in Table B-l. The genus-trophic-values, total
ho phorul(TOTALP), chlorophyll-a (CHLA), and total Kjeldahl nitrogen
(KJEL) in Table B-l are simply mean photic zone values associated with
the domnant occurrences of each genus. TOTALP/CONC, CHLA/CONC, and
SL/cSlc were calculated by dividing the TOTALP, CHLA, and KJEL values
by the corresponding mean cell count. Also given in Table B-l is a
genus-trophic-multivariate-value (MV) calculated for each genus using
the following formula:
MV = Log TOTALP + Log CHLA + Log KJEL - Log SECCHI
B-2
-------
TABLE B-l
TROPHIC VALUES OF SELECTED GENERA BASED UPON MEAN PARAMETER VALUES ASSOCIATED WITH THEIR OCCURRENCES
AS DOMINANTS.
GENUS
Aohnanthea
Aatinaetrum
Anabaena
Anabaenopaia
Ankiatrodeemua
Anomoeoneia
Aplianizomenon
Aphanoaapaa
Aphanotheoe
Arthroapira
Aaterionella
At they a
Binualearia
Bo tryoaoaauB
Carteria
Ceratiwn
Chlamydomonaa
Chlorella
Chromu I ina
CkrooGoccus
Chroomonaa
Cliryaocapea
OwyaooooouB
Cloaterium
CoeldBtrum
Coe loaphaerium
CoBcinodiecua
CoeanariiM
Cruaigenia
DOMINANT
OCCURRENCES
6
2
33
7
9
3
41
4
3
2
36
1
1
2
2
2
4
3
1
19
1
1
2
4
6
6
3
3
2
TOTAL P
29
56
183
70
75
10
147
242
65
51
36
70
42
56
509
140
847
70
8
163
116
10
1580
20
60
44 '
138
14
361
CHLA
11.5
3.5
19.7
32.9
17.9
5.4
37.6
21.1
32.4
21.0
9.6
1.4
6.7
10.3
44.5
5.2
55.1
53.1
10.0
46.6
32.9
7.9
75.0
19.8
13.4
11.7
62.7
9.9
11.8
KJEL
734
594
1015
1393
573
364
1437
1427
1493
1227
491
473
425
1049
1513
1046
3143
991
340
1630
1421
261
4631
698
1208
888
1267
586
1048
TOTAL P
CONC
.027
.142
.098
.008
.082
.005
.058
.034
.009
.022
.023
1.892
.038
.013
.176
3.784
.162
.015
.008
.028
.084
.015
.197
.007
.077
.097
.053
.003
.696
CHLA
CONC
.001
.009
.011
.004
.020
.002
.015
.003
.004
.009
.006
.038
.006
.002
.015
.141
.011
.012
.010
.008
.024
.012
.009
.007
.017
.026
.024
.002
.023
KJEL
CONC
.689
1.508
.545
.165
.626
.166
.569
.200
.203
.519
.310
12.784
.384
.250
.523
28.270
.601
.215
.336
.283
1.032
.380
.576
.249
1.549
1.965
.488
.115
2.019
MV
3.53
3.62
4.82
5.01
4.25
2.32
5.18
5.04
4.98
4.37
3.87
3.23
3.37
4.20
6.04
3.84
6.75
5.13
2.46
5.37
5.50
2.16
7.32
3.60
4.36
3.82
5.25
3.27
4.67
Continued
-------
CO
I
TABLE B-i
TROPHIC VALUES OF SELECTED GENERA BASED UPON MEAN PARAMETER VALUES ASSOCIATED WITH THEIR OCCURRENCES
AS DOMINANTS (Continued)
GENUS
Cryptotnonae
Cyolotella
Dae ty loaoaaopa i a
Die tyoaphaerium
Dinobryon
Euglena
Eunotia
Fragilaria
Glenodinium
Gloeooyatia
Gloeothece
Golenkinia
Gomphonema
Gompho aphaevia
Gymnodinium
Kirohneriella
Lyngbya
Mallomonae
Meloaira
Meriamopedia
Meaoetigma
Micraatinium
Microcyatia
Mougeotia
Navicula
Nitaaahia
Oocyatia
Oacillatoria
Peridiniim
Phaaue
DOMINANT
OCCURRENCES
72
83
58
1
31
8
1
45
4
6
2
2
]
4
2
8
99
6
255
22
1
1
53
2
6
29
5
105
6
2
TOTALP
115
185
178
18
27
318
178
64
8
35
9
615
10
25
9
139
99
87
94
183
57
101
148
76
74
92
38
125
16
2523
CHLA
16.5
29.9
25.0
10.8
8.1
24.5
8.6
17.5
6.4
10.9
4.0
26.9
7.4
8.3
2.8
7.6
29.5
6.0
18.1
33.6
12.8
52.8
37.5
29.2
8.2
26.5
14.0
39.2
8.4
22.8
KJEL
798
1053
1041
949
594
1481
1199
843
403
639
412
1040
782
1270
256
755
1488
642
774
1387
571
1098
1457
990
490
883
1098
1356
595
4049
TOTALP
CONC
.102
.073
.026
.050
.043
.190
3.296
.019
.020
.057
.069
.195
.019
.123
.053
.123
.008
.798
.034
.059
.131
.041
.056
.058
.127
.042
.005
.014
.054
3.955
CHLA
CONC
.015
.012
.004
.030
.013
.015
.159
.005
.016
.018
.031
.009
.014
.041
.016
.007
.002
.055
.006
.011
.029
.021
.014
.022
.014
.012
.002
.004
.029
.036
KJEL
CONC
.711
.418
.153
2.658
.938
.884
22.204
.247
1.025
1.034
3.169
.330
1.507
6.225
1.506
.669
.115
5.890
.277
.444
1.310
.446
.547
.757
.838
.402
.157
.150
2.024
6.346
MV
4.53
4.10
5.05
3.45
3.16
5.70
4.88
4.13
2.34
3.50
2.23
5.60
—
3.65
1.68
4.15
4.98
3.62
4.49
5.34
4.04
5.22
5.27
5.09
3.93
4.78
3.97
5.27
3.01
7.59
Continued
-------
TABLE B-l
TROPHIC VALUES OF SELECTED GENERA BASED UPON MEAN PARAMETER VALUES ASSOCIATED WITH THEIR OCCURRENCES
AS DOMINANTS (Continued)
co
GENUS
Phormidium
Pinnu laria
Raphidiopaia
RhizoBolenia
Roya
Soenedeemue
Sohroederia
Selenaatrum
Sperma tozoopaia
Sphaere I lopeia
Sphaeroeyatia
Sphaerozoema
Spondyloaium
Stouraatrum
Stauroneie
S tepha nodi BCUB
Synedra
Synura
Tabellaria
Tetra&dron
Tetraetrum
Traohelomonae
GENERAL CATEGORIES
centric diatoms
pennate diatoms
flagellate
flagellates
chrysophytan
DOMINANT
OCCURRENCES
3
1
45
1
1
50
2
1
2
1
2
1
1
1
1
73
48
1
20
5
1
4
32
17
108
199
5
TOTALP
172
4
106
31
7
351
17
99
65
57
46
13
21
13
79
166
82
131
22
18
28
97
142
254
154
99
54
CHLA
113.2
0.5
30.5
15.9
2.4
60.4
4.1
9.3
8.8
6.4
11.3
16.6
6.4
16.6
1.9
37.0
19.0
8.9
7.7
5.2
6.9
6.0
24.9
46.8
13.7
14.6
10.5
KJEL
1955
264
1073
1161
332
1826
552
465
1631
532
1274
750
599
750
557
1112
797
1449
455
384
625
867
1000
1615
882
749
635
TOTALP
CONC
.102
.400
.010
.014
.030
.058
.063
.116
.085
.594
.032
.002
.058
.004
9.875
.045
.027
1.056
.015
.040
.043
.292
.033
.036
.075
.054
.010
CHLA
CONC
.067
.050
.003
.007
.010
.010
.015
.011
.012
.067
.008
.003
.018
.006
.238
.010
.006
.072
.005
.012
.011
.018
.006
.007
.007
.008
.002
KJEL
CONC
1.164
26.400
.097
.519
1.437
.303
2.060
.546
2.132
5.542
.897
.128
1.659
.251
69.625
.304
.261
11.685
.307
.859
.963
2.611
.234
.227
.427
.411
.118
MV
5.77
0.78
4.88
4.19
1.68
6.01
2.54
4.13
4.13
3.56
4.23
3.61
--
3.61
3.62
5.27
4.42
5.11
2.86
2.66
3.53
4.38
4.97
5.81
4.55
4.30
3.73
-------
TABLE B-2
PROCEDURE FOR CALCULATING THE TOTALP(PD) PHYTOPLANKTON TSI USING
FOX LAKE, ILLINOIS, AS AN EXAMPLE
Dominant Genera
In Fox Lake
(STORE! No. 1755)
Aphanizomenon
Me losira
Stephanodiacue
Percent
Occurrence
41.2
15.9
15.5
V
(TOTALP, from Table
147
94
166
8)
Sum Total = 406
406
TOTALP(PO) phytoplankton TSI = ~- = 135.6
B-6
-------
TABLE B-3
PROCEDURE FOR CALCULATING THE TOTAUVCONC(P) PHYTOPLANKTON TSI
USING FOX LAKE, ILLINOIS, AS AN EXAMPLE
Genera Counted in
Fox Lake, Illinois
(STORET No. 1755)
Anabaena
Avharri zomenon
Cloaterium
Cmeigenia
Cyalotella
Flagellates
Glenodi.ni.ian
Gcmphoaphaeria
Meloaira
Microcyatia
Oocy8ti,a
Oacillotoria
Fhounidiion
Seenedeamus
Sphaerocyatia
Stephonodiacua
Synedra
Percent of
Count
3.7
41.2
0.3
0.3
1.0
0.3
1.7
1.7
15.9
5.1
4.1
4.1
0.3
3.7
0.7
15.5
0.3
C
CAT gal Units
per ml ]
237
2631
22
22
65
22
108
108
1014
324
259
259
22
. 237
43
992
22
V
(TOTALP/CONC,
Table 8}
.098
.058
.007
.696
.073
.054
.020
.123
.034
.056
.005
.014
.102
.058
.032
.045
.027
V x C
23
153
0
15
5
1
2
13
. 34
18
1
4
2
14
1
45
1
SUM TOTAL = 332
TOTALP/CONCCP) phytoplankton TSI » 332
3-7
-------
TABLE B*4
PROCEDURE FOR CALCULATING THE TOTALP/CONC(PO) PHYTOPLANKTON TSI
USING FOX LAKE, ILLINOIS, AS AN EXAMPLE
Dominant Genera in
Fox Lake, Illinois
(STORE! No. 1755)
Aphaniaomencn
Me loeira
Stephanodiacus
Percent of
Count
41.2
15.9
15.5
C
(Algal Units
Per ml)
2631
1041
992
V
(TOTALP/CONC
Table 8)
.058
.034
.045
V x C
153
34
45
SUM TOTAL = 232
TOTALP/CONC(PD) phytoplankton TSI = 232
B-8
-------
APPENDIX C
CLASSIFICATION, BY VARIOUS AUTHORS, OF THE TOLERANCE
OF VARIOUS MACROINVERTEBRATE TAXA TO DECOMPOSABLE WASTES:
TOLERANT (T), FACULTATIVE (F), AND INTOLERANT (I)
Source: Weber, 1973
C-l
-------
CLASSIFICATION, BY VARIOUS AUTHORS, OF THE TOLERANCE OF
VARIOUS MACROINVERTEBRATE TAXA TO DECOMPOSABLE ORGANIC WASTES;
TOLERANT (T), FACULTATIVE (F), AND INTOLERANT (I)
Macromvertebrate T
Ponfera
Dcmosponpae
Monaxonida
Spongillidae
Spongilla fragilis
Bryozoa
Fctoprocta
Phylactolaemau
Plumatellidae
Plumatella repens
P. princeps var. mucosa . 1 1
P. p. var. mucosa tpongiosa
P. p. var. fruticosa 1 1
P. polymorpna var. repens
CmtatcUidae
Crutaiella mucedo
Lophopodidae
l.ophopodeila carteri
Pectinatella magnifies
Gndoprocu
U mate Ili dae
Urnatetla gracUa
Gymnolaemau
Ctcncntomata
Paludicellidae
Paludiceila ehrenbergi
Coclcntcrata
Hydro704
Hydroida
Hvdridae
Hydra
Clavidae
Cordylophora locusrris
Platyhclminthes
Turbcllaria
Tricladida
Planariidae
Planana
Ncmatoda -
Nematomorpha
Gordioida
Cordiidae
Annelida
Oligochaett 4,3
Plcsiopora
Naididae
/Van
Dero
Opnidonait 14
Stylaria
Tubiflcidae
Tubifex tubifex 11,9
Tubifex 11,6,14
Limnodfilus hoffmeisteri 11,2,9
L. clapjiredianus 1 1
Limnodrtlus 11,6,14
Brancfirura sower by t 9
F
1 1
13
1 1
13
1 1.9
1 1
9
9
9
11
9
1 1
11
11
9
11
9
I
' 9-
1 1
9
1 1.9
Macro in vertebrate T
Prosopora
Lumbriculidae 1 4
Hirudinca
Rhynchobdellida
Glossiphoniidae
Glosiiprtonia complenata \ 1
Helobdella aagnala 11.9
H. nepheloidea 1 1
Placobdella montifera 1 4
P. rugou
Placobdella
Piscicolidae
Piscicola punctata
Gnathobdellida
Hirudidae
Macrobdella 3
Pharyngobdellida
Erpobdellidae
Erpobdella punctata \ 1
Dtna parva \ 1
D. microstoma \ 1
Dina
Mooreobdella microstoma 9
Hydracarina
Arthropoda
Crustacea
Uopoda
Ascllidae
Asellui inter mediui
Asellus 14
Lirceus
Ampnipoda
Talitridae
Hyalleia aiteca
H. knickerbockeri 1 1
Gammaridae
Gammarus
Crangonyi pseudogracilis
Decapoda
Palacmomdae "
Palaemonetes paludosus
P. exilipes ] i
Astacidae
Cambarus stria rus 7
C. fodiera 1
C. bartoni bartoni
C. b. cavatus
C. conasaugaensis
C. asperimanus
C. latimanus
C. acumiratus
C. Hiwassensu
C. extraneus
C. diogenes diogenet 1
C. cryptodytert
F
1 1
9
14
9
1 1
9
9
3
4,2
3,9
9
9
4,2
3
vj
1
1
1
I
4
4,3
I
1
1
1
I
!
1
•Numbers refer to references enumerated in the "Literature"
section immediately following this table.
tAlbinistic
C-2
-------
(Continued)
Macromvertebrate tt T
C. flohdanus
C. carolinusi 1
C. longutus iongirosrris
Procamborus roneyi
- P. acutus acurus 1
P. paemnsulonus
P. spicuiifer
P. versunu
P. pubescent
P. litostemum
P. enoplostemum
P. angustanu
P. semtnolae
P. truculentus$ 1
P. advena$ 1
P. pygmaeust 1
P. pubischelae
P. barbatus
P. Howeilae «
P. troglodytes jf I
P. epicyrtus "*
P. fallax I
P. cfiacti
P. tunzi
Orconectes propinquus
O. rustiaa
O. juvenilis
0. erichsoniamu
Faxontila clypeata
Iruecta
Dipten
Chironomidae
Pentaneura inculta
P. ameosa
P. flavifrons 4
P. melanops * '0,5
P. americana I
Pentaneura ^*
A blabesmyia janta
A. americana
A. iilinoense 5
A. maUochi
A. ornate
A. aspen
A. peleensis
A. auriensis
A. rhampite
A blabesmyia
Procladius culiciforma 1 <*
P. denticulatus 9
Procladius 5
Labrundinia floridana
L piloseUa
L vireseent
Guthpeiopia
Conchapelopia T
Coelotanypus scapularisji
C. concuuna 9
F
1
1
1
1
1
1
I
1
1
I
i
I
1
9
9
t-
I
14
14. 10
2.3
9
1 1.14
10
9
3
9
10,5
3,10
5
9
9
9
1 1,14
10.5
I
I
1
I
I
1
2.3
14,5
,
10.5
9.10
4
3
3
3
3
9
3
9
3
10
Macroinvcrtebrate T i
Pniotanypui btllus -'
Tanypus stellatus '0,5
f. cannatus
T. punctipennis
Tanypus
Psectrotanypus dyari ! 0 . 5
Pucavtanypus
Ljrsia lurida
Clinotanypus catiginosus
Clinotanypus
Orthocladius abumbratus
OrrHocladius
.Vanocladius
Pxcrrociodius niger
P. julia
Psecirocladius
Metriocnemus lundbecki
Cricotopus bicinctus
C. bicincrus grouo -
C exilis
C. exilis group
C. erifasciana
C. trifasdatus #roup
C. poiirus
C. tricincrus
C. absurdus
Cricotopus
Corynoneura (arts
C. saitetlata
Corynoneura
Tftienemanniella xena
T.hienemanniella
fricttocladius robacki
Srillia par
Diamesa nivoriunda
Diamesa
Prodiamesa olivacea
Chironomus atrenuatus group 4 , 3
9,5
C. riparius 6,10
5
C. ripartus group '3
C. tentans
C. tentans-plumosus 1 4
C. ptumosus M,6
14
C plumosus group 9
C. carus 3
C. crassicaudatus 3
C. stigma terus 3
C. flovus
C. tquisitus
C. fulviptlus 3
C. anthracinu*
C. paganus
C. staegeri
F I
i
6. 1 4
-
10.5
10.5
1 \
10
3
i
T
4,10
o
•i
i 0
)
10
9
10
14
1 4
5
1
.*
10,5
' 4
14,9
1 VJ , 3
3.3
J, 10
.1
i . .5
1 u , 3
i
:".
in, 5
;>
6. 10
5
id
;^
•10, i>
~ 3
5
3,9
3, 10
2.3
3
6.9
10
14
5
10
5
11,5
5
5
|Not usually inhabitant of open water; are borrowers.
C-3
-------
(Continued)
Macroinvcrtebratc T
Chironomus 4
Kiefferullus dux 3
Cryptochironomus fulvus 2 . 3
C. fulvus group
C. digitatus
C. sp. B (Joh.)
C. blarina
C. p» tract nut
C. nais
Cryptochironomus. 4
Chaetolabis atroviridis
C. ochreatus
EndocMronomus nigricans
Stenochironomus macateti
S. hilaris
Stictochironomus devincrus
S. varius
Xenocftironomut xenolabis
X. rogersi
X. scapula
Pseudochironomus ricfvrdson
Pseudochironomus
Pancnironomus aborrivus group
P. pectinatellae
Crypcotendipes emorsus
Microtendipei pedellus
Microtendipes
Paratendipes albimanus
Tribelos fucundus
T. fuscicornis
Harnischia collator
H. tenuicaudata
Phaenopsectra
Dicrotendipes modesrus
D. neo modest us
D. nervosus
D. incurvus 9
D. fumidus
Gtyptotendipes senilis
C. paripes 3
G. meridionalis
G. lobiferus 11,3
9
G. barbipes 9
G. amptus
Gtyptotendipes 5
Polypedilum fulterale
P. fallax
P. scalaenum 3
P. illinoense
P. tritum
P. simulans
P. mibeculosum
P. vibex
Polypedilum
Tanytarsis neoflavellus
T. gracilenrus
T. daaiHilis
Rheotanytanus exiguus 4
Rfieounytanus
F
14
G
1 1
9
9
3,9
9
9
9
9
9
9
10
9
9
9
9
4, 10
5
9
2.3
9,10
9
9
1 1, 10
10.5
9
1
10.5
10.5
5
4
5
14
5
5
10,5
9,10
2.3
3.5
10
9
10,5
10,5
5
10,5
9
10,5
5
9
10
9
9,5
5
9.5
9
5
10.5
3,5
3
10.5
5
5
10
5
6
5
9
2.3
Macroinvertebrate T
Cladotanytarsus
Micropsecrra dives
M. deflects
M. nigripula
Caloptectra gregarius 4
Calopsecrra
Slempellina /ohannseni
Cuticidae 3
Culex ptpteru 6,10
A nopheles puncripennis
Chaobondae
Ovoborus puncripennis
Cera topogon idae 4,3
Palpomyu libiaSs
Palpomyia
Bezzie glabra 1 0
Stilobezzia antenaSs 1 0
Tipulidae 3
Tipub caloptera
T. abdominalis
PseudolimnopMla luteipennis
Hexatoma
Eriocen
Psychodidae 3
Prychoda attempts \ 0
P. sctuzwa . 1 0
Psycnoda 9
Telmatoscopus albipuncratus 1 4
Telmatoscopus
Simulidae 9
Simultum vittatum
' S. venustntm
Simultum
Pnsimulium johonnseni
Cnepnia pecuarum
Stratiomyiidae 3
Stntmmys discalis 1 0
S. meigeni 1 0
Odontomyia cincta
Tabanidae 3
Tabanus atratus 6
T. stygius
T. benedictus 1 0
f. gtganteus
T. lineola 1 0
T vaiegatus
Tabanus
Syrphidac 3
Syrpnus amehcanus 1 0
Enstatis bastardi 6,10
E. aenaus 1 0
E. broun 1 o
Ehstalis i o
Empididae
Ephydridae
Bncttydeutera argentata ' 0
Anthomyiidae
Lepidoptera
Pyralididae
Trichoptera
Hydropsychidae
Hydropsyche orris
t-
9
14
10
14.9
9
14
11,14
9
14
10
6.10
:o
10
10
9
9
4.3
9
i
i
5
j
10.5
10,5
5
10
10
10
10
10
10
*•
10
.;. j
10
2
10
10
10
10
10
C-4
-------
(Continued)
Macromvertebrate T
H. bifida group
H. simulant
H. fruoni
H. incommoda
Hydropsyche
Cheumatoptyche
Macronemum Carolina
Macronemum
Potamyia flava
Psychomyudae
ftyc/tomyw
.Veureclipsis crepuseularis
Polycentroput
Cyrnellus fraternus
Oxyethm
Rhyacophiiidae *
RHyacopntla .
Hyd/optiUdae
Hydroptila tuaubesiana
Hydroptila
Ockmtricnia
Agraylea
Leptocendae
Leptocella
A ttinpsodes
Oecttu
Philopotamidae
Chimarn perigua
Chimam
Brachycentridae
Bracnycentna
MoUnrudae
Ephemeroptera •'
Hepugenudae *
Stenonema integrum
S. rubnmaculacum
S. fuscum
5. putchcllum
S. ares
S. sdtulum
5. femorantm
S. terminatum
5. mterpunctaeum
S. L oHioenst
S. L canaderae
S. i. heterotarule
S. exiguum
S. smieiiae
S. proximum
5. aipuncatum
Stenonema
Hexageniidae
Hexagenia timbata
H. bilintau
Paitagtnia vingera
Baeudae *
Baetis vagam
Callibattis floridontu 3
dilibaecis
F
9
1 1
4, 14
2,3
9
. 9
9
9
4,3
4.3
15,9
15
ts
9
6.9
15
14
6
I
9
9
4.2.3
4,3
4,2.3
9
9
9
4, i 1
3
4,3
1 1
9
4,2.3
9
9
11
9
9
2,3
4.3
3
1 1
15
15
15
9
15.9
15
15
4,2,3
4,2.3
2
IS
15
9
1 1
9
9
Macromvertebrite T
Caenidae
Ciena diminuta 3
Caena
Tricoryttudae
Siphioniuidie
fsonycfiia
Plecoptera
Perlidae
Periesra placida
A croneuna abnarmis
A. arida
Nemouridae
Taeniopteryi nivalis
Altocapnia vniparvt
Perlodidae
Isoperia itttncara
Neuiopcera
Sisyrtdae
Climacia areolarif
Megaioptera
Cotydaiidae
Cory da lit comueus
Sialidae
Sialit infumata
Sialii
Odonau
Calopterygidae
Hetaenna tiria
Agnorudae
Argia apicaiis
A. translate
Argia
tichnura vernealis ' ''
Enallogma antennarum
E. signatum
Aeshnidae
A nax lunrui
Comphidae
Gomphus pailidus
G. piagiams
C. extemus
G. sptniceps
G. vastus
Gomphvs
Progomphut
Dromogompnut
Erpetogomphus
LibeUulidae
Libellula lydia
.Veurocordulia moesra
Plathemis
Macromia
Hemiptera 3
Cohxidae
Corixa 5
Hetptrocorixa 6
Cetridae
Gerris 6
Belostomaadae
Betonoma 6 , 2
Hydro me tridae
Hydromerra martini 2
F 1 1
i
9
9
6
3
\
6
i
9
9
9
a
9
3
4,2,3
o
9'
4,3
o
j
6
•a
9
4,9
9
! !
9
4,3
2
9
9
9
9
4 .^.
1 1
3
-t,3
1 i
1 1
1 1
1 i
4,3
3
-------
(Continued)
Coleopteri 3 §
Elnudae
Stetelmu crenatt
S. textineata
S. deconu ' 2
Dubirapfaa
PfOfnOfesa
Optiourvus
Macronycfna glabratus
Anacyronyx variegarus
Microcylloepus puailut
Gonieimis dietrichi
HydiophiMdae
Beroau 9
Tropiaenuu nautor 6
T. latentii 2
T. donate
Dydscidae
LaccopHUut macuioaa 6
Gyrinidae
Gyrinu floridamu 2
Dtneututamericama 6
Dineuaa
Molluscs
Gastropoda
Mesogutropoda
Valvatidae
Vahnta ficaruiata
V. pucinaUt
V. bieahnata
V. b. var. normalit
Viviparidae
Vtvaparus contectoules
V. subpurpvrea
Campeloma integnim
C.rufum
C conttcrus
Cfatdaau
C dedsum
C jubtoSdum
Camptloma
Lioplax subcainaaa
Pleurooeridae
Pleurocen acuta
P. etewtum
P. e. lewdi
Pleurocen
Goniooaas toetceiu
G. vJrji/ricB 3
Goniobasii
AncuioK
Buiinudae
Butinea tenmculatta
Amnicole emarginate
A. limom
Somatogyrus aibgloboaa
Basonunatophoca
Phyadae
Phyta Integra 6,3
P. heteroaropha Q
F
9, 12
9,12
12
12
9
3
3
3
8
3
8
11.8
14
11.8
3
8
8
11.8
3
3
3
3
3
i
6,12
6
12
12
12
12
1 1
11.8
1 1
1 1
1 1
1 1
8
1 1
4.3
1 1
1 1
11
Macroinvatebrate T
P. gyrina
P. acuta
P. fontinalit
P. anatitia 3
P. hatei 3
P. cubenas 3
P. pumitia 2
Phyja -1.3
Aplexa Hypnorum
Lymnaeidae
Lyrrwaea ovata 3
L peregn
L. capenta
L. humiOs
L. obruat
L pohutris
L. aioiculeria
L. itagTtatix
L. t. appreae
Lymnaea 3
Pseudoatcemea cohtmeUa
Gaiba catesaopium 8
Foaaria modtceOa 3
Planoibidae
Planorbit cainatut
P. trivolvu 3
P. paua 8
P. comeus
P. m&ginana
Planorbu
Segmentina armigera 3
Hetisoma anctpt
H. trwolvis
Kelisoma 2 . 3
Gyimitus arcticus
Gyreuhu
Ancybdae
Aneyba laaatris
A. fhtviefitis
Ferriaiafusos
F. larda
F. rivufera
Ferrisaa -».2.3
Bhratvia
Eulamellibranchia
Margaritifeiidae
MGrguritiffn tnti/gui irif&o
Unionidae
Unto complaneta 3
U. gibbosus a
U. batevui
U. piemmm
U. tvmidus
LampaOi luteola
L. aiets
L anodontoidet
Lgracm
L. porvut
Lampsilis
Quadrate pvstuioK
F
3
o
3
3
8
3
3
3
3
3
9
3
3
8
B
3
8
3
3
3
3
8
9
8
8
• 8
3
3
1 1
1 1,9
8.9
t
5
3
n
3
3
;1
8
r.
3
8
8
8
S
3
1 1
§ Except nflto beetles
C-6
-------
(Continued)
Macroinvertebraia T
2. unduleta
Q. rubigmoa
Q. lachrymoa
Q. piiaaa
Trunalla donaaformit
T. etegua
Tririgonia tubavulata
Symphynou cosata
Stroptotus edentuba
AnodoHU grandit
A. imbealHi
A. muubiUs
Alosmodoaa costs ta
Proptav alaa
Leptodea fngilis
A mbiema undulata
Lasmigoaa compianata
Obtujuaria reflexa *)
Heterodonu 4
Corbie uiidae
Corbicuia maniUnsu
Sphaerudae -^,3
SptuoTum notatum 3
S. corneum
S. rhomboideum
5. striatimtm
S. 1 var. oorjjutentum
F
3
3
3
3
a-
3
3
3,9
3,9
3
3
3
8
3
3
1 1
I
1 i
1 1
3
9
9
t a
9
Macroinvertebrace T
5. s. vai. lilycashense
S. sulcaium
S. sumineum
S. motnamtm
S. vivioolum
5. solidulum
Sphaenum
Musculium securis
M. rrantveratm 11,3
M. truncatum I 1
Musculium 1 ^
Pisidium abditum 3
P. fossartnum
P. paupercuhim crysulense
P. amnimm
P. casernnum
P. compmatm ' 1
P. fallax
P. Herutorvanum
P. idahoentis 3
P. complanatum ' ' , 3
P. subtnincatum
Pisidium
Dreussenndie
Mytilopas leucophoearus
Mactndae
Rangia cuneata
C.'
: i
•f
' ' . 3
3
^
o
9
3
3
3
1 1 .3
3
i
3
•;
! i . 3
3
• 1
3
3
!
3
3
3
ri
3
I
C-7
-------
REFERENCES
1. Allen, K.R. The Horokiwi Stream—a Study of a Trout Population. Mew
Zealand Marine Dept. Fish Bull. No. 10, 1951.
2. Beck, W.M., Jr. Biological Parameters in Streams. Florida State Board
of Health, Gainesville. (Unpublished).
3. Beck, W.M., Jr. Indicator Organism Classification. Florida State v
Board of Health, Gainesville. Himeo. Rept. (Unpublished).
4. Beck, W.M., Jr. Studies in Stream Pollution Biology: I. A Simplified
Ecological Classification of Organisms. J. Fla. Acad. Sciences,
17:211-227, 1954.
5. Curry, L.L. A Survey of Environmental Requirements for the Midge
(Diptera: Tendipedidae). In: Biological Problems in Water Pollution.
Transactions of Third Seminar, C.M. Tarzwell, ed., USDHEW, PHS, Robert
A. Taft Sanitary Engineering Center, Cincinnati, 1962.
6. Gaufin, A.R. and C.M. Tarzwell. Aquatic Macroinvertebrata Communities
as Indicators of Organic Pollution in Lytle Creek. Sewage and Ind.
Wastes. 28:906-924, 1956.
7. Hubbs, H.H., Jr. List of Georgia Crayfishes with their Probable
Reactions to Wastes (Lethal Chemicals not taken into Consideration).
Mimeo. Rept/ (Unpublished), 1965.
8. Ingram, W.M. Use and Value of Biological Indicators of Pollution:
Fresh Water Clams and Snails. In: Biological Problems in Water Pollu-
tion. C.M. Tarzwell, ed. USDHEW, PHS, R.A. Taft Sanitary Engineering
Center, Cincinnati, 1957.
9. Mason, W.T., Jr., P.A. Lewis, and J.8. Anderson. Macroinvertebrate
Collections and Water Quality Monitoring in the Ohio River Basin,
1963-1967. Cooperative Report, Office Tech. Programs. Ohio Basin
Region and Analytical Quality Control Laboratory, WQO, USEDA, NERC-
Cincinnati, 1971.
10. Paine, G.H., Jr. and A.R. Gaufin. Aquatic Diptera as Indicators of
Pollution in a Midwestern Stream. Ohio J. Sci. 56:291, 1956.
11. Richardson, R.E. The Bottom Fauna of the Middle Illinois River,
1913-1925: Its Distribution, Abundance, Valuation and Index Value in
the Study of Stream Pollution. Bull. 111. Nat. Hist. Surv. XVII
(XII):387-475, 1928.
12. Sinclair, R.M. Water Quality Requirements of the Familiy Elmidae
(Coleoptera). Tenn. Stream Poll. Cont. Bd., Dept. Public Health,
Nashville, 1964.
13. Tebo, L.B., Jr. Bottom Fauna of a Shallow Eutrophic Lake, Lizard Lake,
Pocahontas County, Iowa. Amer. Midi. Nat., 54:89-103, 1955.
C-8
-------
14. Wimmer, G.R. and E.W. Surber. Bottom Fauna Studies in Pollution
Surveys and Interpretation of the Data. Presented at: Fourteenth Mid.
Wild!. Conf., Des Moines, Iowa, 1952.
15. Lewis, P.A. Mayflies of the Genus Stenonema as Indicators of Water
Quality. Presented at: Seventeenth Annual Meeting of the Mid. Benthic
Soc., Kentucky Dam Village State Park, Gilbertsville, Kentucky, 1969.
C-9
-------
APPENDIX D
KEY TO CHIRONOMID ASSOCIATIONS OF THE PROFUNDAL ZONES OF
PALEARCTIC AND NEARARCTIC LAKES
Source: Seather, 1979
0-1
-------
APPENDIX D
Key to chironomid associations of the profundal zones of Palaearctic and
Nearctic lakes
In the key "absent" means less than 1% as accidental occurrence may take
place, "present" means more than 1%. The limit of 2% is regarded as the
level above which the species can be regarded as a persistant non-
accidental member of the community, while the 5% limit is a level above
which the species can be said to be a common member of the community.
These limits should of course not be regarded rigidly if the samples are
few.
1. Pseudodiamesa and/or Qliveria tricornis present a -oligotrophic
The above absent 2
2. Heterotrissocladiust Protanypus, Micropsectra or Paracladopelma
present and making up at least 2» of the profundal chironomids
oligo- mesotrophic lakes „ 3
The above absent or making up less than 2« of the profundal chirono-
mids eutrophic lakes 10
3. Heterotrissocladius subpilosus - group present, tribe Chironomini
absent from the true profundal zone- 0 -oligotrophic
H_. subpilosus group present or absent,
tribe Chironomini present 4
4. Heterotrissocladius subpilosus group, Protanypus caudatus group,
Micropsectra groe"nTand1ca or Paracladius spp. present and making up
more than 5% of the profundal chironomids 5
The above absent or making up less than 5«
of the profundal chironomids 7
5. Protanypus caudatus group or Paracladius usually present, Chironomus
absent, Phaenopsectra (including Sergentia) and Stictochironomus at
most present in very low numbers (<2%) y -oligotrophic
When Protanypus caudatus group or Paracladius present, Chironomus,
Phaenopsectra or Stictoc"hironomus present in low numbers
(>2%) 6
6. Heterotri ssocl adi us sufapilosus group plus H_. maeaeri group more common
than hL marcidus group; Chironomus
making up less than 2% <5 -oligotrophic
Heterotri ssocl adi us subpilosus group plus H_. maeaeri group absent or
less common than H_. marcidus group: Chironomus usually makes up more
than 2% e. -ol i gotrophi c
7. Heterotri ssocl adi us. Paracladopelma nigritula. P_. galaptera, Micro-
psectra notescens group, Monodiamesa tuberculata, Macropelopia
fehlmann'l and/or Tanytarsus bathophllus common (>5D f -oligotrophic
The above at most present in
very 1 ow numbers 8
D-2
-------
8. Mlcropsectra and/or Monodlamesa common, more or about as connnon as
Stlctochironomus and Phaenopsectra, or Chironomus except sallnarlus or
semireductus types T -mesotropni c
Mlcropsectra and/or Monodlamesa less common than Stictochironomus and
Phaenopsectra or spp. of Chironomus except salinarius or semireductus
types .*. ~3
#
9. Monodlamesa, Protanypus. Heterotrlssocladlus, Stictochlronomus.
Phaenopsectra or Chironomus salinan'us and semlreductus types more
common than other ChlrononJus spp 9 -mesotrophlc
The above less common than other Chironomus i -mesotrophic
10. Heterotrissocl adius. Protanypus, Mlcropsectra, Paracladopelma
nigritula or P. gaTaptera present in low numbers *-eutrophic
The above absent 11
11. No chironomlds present o-eutrophic
Chironomids present 12
12. Only Cnlronomus plumosus type and Tanypodinae present < -eutrophlc
Other Chironomids also present 13
1
13. Only Chironomus and subfam. Tanypodinae present ?-eutrophic
Other groups also present 14
14. Only tribe Chironomini, Tanytarsus spp. and subfam. Tanypodinae
present M -eutrophi c
Other groups also present X -eutrophic.
D-3
-------
B:
TABLE °-1 , X AMH
PROFUNDAL CHIRONIMIDS IN NEARARCTIC( ..... ) AND
} LAKES FULLY DRAWN LINES AND FILLED CIRCLES:
BB5rr% or. ssr
IN EUROPE, BOREAL.
SPECIES
OLIGOHUMIC
OLIGOTHOPWC
iMCSOTROPMd
a Ifllr I 8 I « 1C
: x-oi
•CPI
Pstudodiomtsa nivosa Goetgh.
Pstudodiamtsa arctica (Mall.)
Olittria tricornis (01.)
iouttrboritia sedna (01.)
Porocladius quodrinodosus Hirv.
Pf atony pus caudal us (Edw.)
Httfrotriisoclodius sutipilosus (Kieff.)
Hettrotfissoeladius olivtri S«ih.
Monodiam«sa tkmani 8'und.
flrg/aa/pus taelhtri Wied.
Tanytarsus palmtni Lind.
Laultrtiornia coraeina Kieff.
Paraeladius a I pie old (Zett.)
Monodiam0sa a/pico/a Bfund.
Protanypus forcipotus Egg.
Micropsectra graenlandica And.
Hef»rotrissoclodius matoeri S'und.
Protonypus hamiltotii Sastn.
Hicropsectro lindlbtrgi T^w.
ranytarsus lugtns Kieff.
Htttrotrissocladius so. ^ near sudpilosus
Htttrolrissoc/adius »p. 5 ««or maeatri
Protanypus ramosus Sczlh.
Micropsictra eon tract a Reiss
Uicropstetra insigniloous Kieff.
Poroc/odopt/ma galapttra (Town.)
Paractadoptlma nigritula (Gaetgh.)
Monodtamesa tuotrcuiata Sath.
Macroptlopia fehlmanni (Kieff.)
ranytarsus oattiapnilus Kieff.
Protanypus atorfa Zett.
Hettrgtrissocladius cfianqi Sdth
Hettrotrissoctadius scute/lotus Gosfqh
Htttrotrissocladius sp. '5 near cnangi
Pretonypus sp. ,4 near morio
PfOtanypus *p. 5 near morio
Tanytarsus decipitns Uind.
Monodiamtsa nitida (Kieff.)
Heterotrissocladius grimshawi Edw.
Manodiameso jp. pas*, prolilobata S«th.
Monodiamtsa oathyphila (Kieff.)
Stictochironomus rosenschoeldi (Zett)
Phoenopsectra coraeina (Zeft.)
Tanyiarsus n. sp, Igstagti • oqgl.
Monodiamtsa dspectinato S«th.
Chironomus atntioia Mall.
Chironomus anthracinus Zett.
Tanytarsus inaequglis Goetgfi.
fgnytarsus gregarius Kieff.
Chironomus plumosus f. semirgductus
Cryptotgndipgs casuarius (Town.)
Chironomus decorus Joh.
Cryptottndipts daroyi (Subl.)
Chironomus plumosus U.
laiutschia talutschicoia Lip.
Chironomus renuistytus Srund.
x|x
"
Lli'J
D-4
-------
TABLE D-2
CHARACTERISTIC SUBLITTORAL AND LITTORAL CHIRONOMID HABITATS IN
NEARACRTIC AND PALEARCTIC LAKES.
SPECIES
HtrtrofriStOdadiuS SuPpilOSuS (Kisff.)
HertrotriSSOC/adius oiirtri Satn.
Hydrooognus futisfy/ui (Goetqh.)
Zaiutschia inganaeigs Sain.
Abisnomyio virgo Edw.
Og*londid oorgalis ' Kieff""""" ~" • ~~
Ortnocladius (0.1 trigonolaois Edw
Ortnoclddius (PI consoormus Holmqr.
Olivgria trteornis (01.)
Hydrooaenus martini Sain.
Hydrooognus canformis conforms (Holmqr.)
Hydrooatnus canformis laoradortnsis Sain
Monodiomtsa sumani Brund
Pyroc/adius yuadrinodosui Hiry
Tanyiarsus luygns Ki«ff.
Pardtdayfanui hyptfoorgut Brund.
Porac/adius alpicola (Zstt)
Parddddopglma mgritulu (Goetgn.)
Sfictocfiironomus rosenscnotldi (Ztti.)
Micropsicfa gratnldndiea And.
Arcfopttopid OdfOilarsiS (Ze«t.)
Micropstctra lindebtrgi Sow
rhienentannifyig .fusactps (Edw.)
Mesocricoraa-js tfiiantmonni (Goetgn.)
Lduttroormo coracmo Kieff.
Hflsrgrrissaclsdius mdrcidus (Wolli.)
Zaiaiscnia oo septa I'.VeOO)
Hturotnssoclodi'js hirtgpti Satn.
Htrtrotanytarsus ptrennis Sath.
Hfttroranytgrsus nudatus Sath.
Zal utSChiQ t<3/'jt$CfiiCQiQ Lip.
Nwyctadivs (NJ ifiGomptus Sath
Na/iocta4ms (N J mt/it/rn/s Safh.
5/ff/TTptf//mj n 5o near <3/mt Brynd.
>Va/»^c/tfrf/tf5 (N.J at*X
\
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L
i
T
i v
3
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I^B^
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I/ VI,
tf«"
fc/ V.
••^
W1
....
Ou
/3
"""
PMMH
>* V '-*
'•-"-
tw
1 1
L..
. . j
'
."vw
....
U...J
^ ^
oo r
y
" "
....
90°
i
....
•^•^M
^••^
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«'C
<
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....
•••^•B
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.
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D-S
-------
TABLE D-2
CHARACTERISTIC SUBLITTORAL AND LITTORAL CHIRONOMIDS OF HABITATS IN
NEARACRTIC AND PALEARCTIC LAKES (Continued)
SPECIES
Ai8
OLiGOHUMlC
:=S
cur
"
a I 0 ! y
Paracladopelmo. »inn»lli Jaefcs.
Paraciadopeimo undine (Town.)
Slemptllintlla minor (Edw.)
Stempellineila ortvis Edw.
Pogastie/la orophila (Edw.)
Pogastiella oslansa (W«bb)
Nanocladius IN.) distinctus (Mall.)
Cladopelma edwordsi (Kru»)
Zalutscnia Ungulata S-ath.
Saetheria tylus (Town.)
Heterotrissododius latilaminus S«t(v
Pseudocnironomus rtx Haub.
Phaenopsectra albescens (Town.)
Psectrocladius fPJ psilopttrus Ki«H.
Uonopelopia lenuicalcar (Kieff.)
Cladoptlmo viridula (Fabf.)
Nanocladius (N.) dicotor (Zstt)
ffoaacliia demaijerei (Krus.)
Cryptottndipes casuarius (Town.)
Uonodiamtsa deptctinata SoMh.
Pseudochironomus futvi«»ntris (Joh)
Pstctrocladius (P.) simulans (Jon)
Chironomus plumosus t. semireductus
Pseudochironomus pseudoviridis (Moll.)
Nanocladius (N.) ba/Hcus (Palm.)
Chironomus antftracinus Zett.
Harnischia curtilamellatd (Mall.)
Slictochironomus histrio (Fabr.)
Oemicryptoctiironomus talneratus (Z«lf.)
Cryptotendipes darbyi (Subl.)
Chironomus deeorus (Joh.)
Oicrotendipes narvosus (Sloeq.)
Endochironomus subiendens (Town.)
Eadochironomus nigr icons (Joh.)
Endochironomus alDipennis (Meiq.i
Cricolopus (I.J syl\f«stris (Fobr)
Chironomus plumosus L.
Glyptotendipts (P.) paripts Edw.
Polypedilum (Po.i nuOacu/osum (Meiq.)
Cladotanytarsus mexionensis 8 fund.
Cladotanytarsus near *exionensis
Tynytorsus usmatnsis Paq.
Einfeldia synchrony (01.)
Ein/eldia dissidens (Walk.)
Chironomus fCaJ tentans Fabr.
Tanypus punctipennis (Meiq.i
Laorundinia longipolpis (Goetqh.)
Psectrocladius (A.) pldtypts Edw.
Zalulschia mucronata (Srund.)
I
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D-6
------- |