vvEPA
United State*.
,-rvironmenfp; Protection
Agency
EPA-600'3 80-078
July 1980
Research and Development
Ecological Studies of
Fish Near a Coal-
Fired Generating
Station and Related
Laboratory Studies
Wisconsin Power
Plant Impact Study
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RESEARCH REPORTING SERIES
Research reports of the Office of Research and Development, U S Environmental
Protection Agency, have been grouped into nine series These nine broad cate-
gories were established to facilitate further development and application of en-
vironmental technology Elimination of traditional grouping was consciously
planned to fosler technology transfer and a maximum interface in related fields
The nine series are
1 Environmental Health Effects Research
2 Environmental Protection Technology
3 Ecological Research
4 Environmental Monitoring
5 Socioeconomic Environmental Studies
6 Scientific and Technical Assessment Reports (STAR)
7 Interagency Energy-Environment Research and Development
8 "Special" Reports
9 Miscellaneous Reports
This report has been assigned to the ECOLOGICAL RESEARCH series This series
describes research on the effects of pollution on humans, plant and animal spe-
cies, and materials Problems are assessed for their long- and short-term influ-
ences Investigations include formation, transport, and pathway studies to deter-
mine the fate of pollutants and their effects This work provides the technical basis
for setting standards to minimize undesirable changes in living organisms in the
aquatic, terrestrial, and atmospheric environments
This document is available to the public through the National Technical Informa-
tion Service, Springfield, Virginia 22161.
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EPA-600/3-80-078
July 1980
ECOLOGICAL STUDIES OF FISH NEAR A COAL-FIRED
GENERATING STATION AND RELATED LABORATORY STUDIES
Wisconsin Power Plant Impact Study
by
John J. Magnuson
Frank J. Rahel
Michael J. Talbot
Anne M. Forbes
Patricia A. Medvick
Institute for Environmental Studies
University of Wisconsin-Madison
Madison, Wisconsin 53706
Grant No. R803971
Project Officer
Gary E. Glass
Environmental Research Laboratory-Duluth
Duluth, Minnesota
This study was conducted in cooperation with
Wisconsin Power and Light Company,
Madison Gas and Electric Company,
Wisconsin Public Service Corporation,
Wisconsin Public Service Commission,
and Wisconsin Department of Natural Resources
ENVIRONMENTAL RESEARCH LABORATORY-DULUTH
OFFICE OF RESEARCH AND DEVELOPMENT
U.S. ENVIRONMENTAL PROTECTION AGENCY
DULUTH, MINNESOTA 55804
. V* '
• H*
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DISCLAIMER
This report has been reviewed by the Environmental Research Laboratory-
Duluth, U.S. Environmental Protection Agency, and approved for
publication. Approval does not signify that the contents necessarily
reflect the views and policies of the U.S. Environmental Protection Agency,
nor does mention of trade names on commerical products constitute
endorsement or recommendation for use.
ii
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FOREWORD
The U.S. Environmental Protection Agency (EPA) was established to
coordinate our country's efforts toward improving and defending the quality
of the environment. These efforts depend greatly on research to monitor
environmental change and to determine health standards.
One project the EPA is supporting through its Environmental Research
Laboratory in Duluth, Minnesota, is the study "The Impacts of Coal-Fired
Power Plants on the Environment." The Columbia Generating Station, a coal-
fired power plant near Portage, Wisconsin, has been the focus of all field
observations. This interdisciplinary study is conducted by the
Environmental Monitoring and Data Acquisition Group of the Institute for
Environmental Studies at the University of Wisconsin-Madison and involves
investigators from many departments at that same university. Several
utilities and state agencies also are cooperating in the study: Wisconsin
Power and Light Company, Madison Gas and Electric Company, Wisconsin Public
Service Corporation, Wisconsin Public Service Commission, and Wisconsin
Department of Natural Resources.
Reports from the study will appear as a series within the EPA
Ecological Research Series. The topics will treat chemical constituents,
chemical transport mechanisms, biological effects, social and economic
effects, and integration and synthesis.
The Columbia Generating Station has caused changes in nearby
wetlands. Since the area has a diverse fish community, the fish-monitoring
group of the Columbia Generating Station impact study has been studying the
effects of habitat loss and habitat degradation on fish. This report
assesses the station's effects on fish reproduction and documents research
on the use of temperature preference to detect sublethal concentrations of
zinc in bluegills (Lepomis macrochirus) and zinc selection for tolerance
over four generations of flagfish (Jordanella floridae).
Norbert A. Jaworski
Director
Environmental Research Laboratory
Duluth, Minnesota
iii
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ABSTRACT
Construction of a coal-fired electric generating station on wetlands
adjacent to the Wisconsin River has permanently altered about one-half of
the original 1,104-ha site. Change in the remaining wetlands continues as a
result of waste heat and ashpit effluent produced by the station. Leakage
of warm water from the 203-ha cooling lake is causing a shift in the
wetlands from shallow to deep-water marsh. Coal-combustion byproducts enter
the wetlands from the station's ashpit drain. Since this area was known to
have a diverse fish community and to be a spawning ground for Wisconsin
River game fish, we studied the effects of this habitat loss and degradation
on fish populations. In laboratory experiments we investigated the use of
temperature preference and activity as a sublethal bioassay. In selection
experiments we examined the potential of fish to evolve metal-tolerant
populations in chronically contaminated environments.
Three years of netting documented the continued use of this area by
spawning fish despite extensive habitat alterations. An inventory of
potential spawning marshes along the Wisconsin River between the dams at
Wisconsin Dells and Prairie du Sac showed that the station site still
contained 22.0% of the deep-water sedge meadow and 0.8% of the shallow-water
sedge meadow likely to be used by spawning game fish. Construction of the
power plant resulted in a loss of 18% of the shallow water sedge meadow
formerly available in this section of the Wisconsin River. Loss of deep
water sedge meadow was negligible. In situ and laboratory experiments
showed that the ashpit effluent was not acutely toxic to eggs or larvae of
northern pike (Esox lucius), although some reduction in hatchability was
attributed to the flocculent precipitate found in the ashpit drain.
Analysis of population structures of northern pike showed a weak year-class
for fish hatched in the first post-operational year. Further monitoring
will be needed to determine if the reduction was due to the generating
station or to natural factors.
A bioassay utilizing temperature preference and activity proved no more
sensitive than bioassay methods used by previous investigators. Bluegills
(Lepomis macrochirus), themoregulating in a temporal gradient, tended to
increase activity and decrease preferred temperature after exposure to 2.5
mg/liter zinc. Neither change was statistically significant, however, and
both factors returned to normal levels within 2 days.
A population of flagfish (Jordanella floridae) selected for zinc
tolerance was more resistant than, the control population for the first two
generations but not after three generations. The failure of continued
selection to produce ixicreasing zinc tolerance may have been caused by
inbreeding depression or by cumulative carry-over effects of zinc passed
iv
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from mother to offspring through the egg cytoplasm. The discrepancy between
laboratory-selection experiments and field observations of fish living in
chronically metal-contaminated environments is discussed.
This report was prepared with the cooperation of faculty and graduate
students at the Limnology Laboratory at the University of Wisconsin-Madison.
Most of the funding for the research reported here was provided by the
U.S. Environmental Protection Agency (EPA). Funds were also granted by the
University of Wisconsin-Madison, Wisconsin Power and Light Company, Madison
Gas and Electric Company, Wisconsin Public Service Corporation, and
Wisconsin Public Service Commission. This report was submitted in
fullfillment of Grant No. R803971 by the Environmental Monitoring and Data
Acquisition Group, Institute for Environmental Studies, University of
Wisconsin-Madison, under the partial sponsorship of the EPA. The report
covers the period of July 1975-78 and work was completed as of April 1979.
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CONTENTS
Foreword
Abstract iv
Figures viii
Tables x
Acknowledgments • xii
1. Introduction 1
Design of the fish study 5
2. Conclusions •• 7
3. Effects of the Columbia Generating Station on Fish Spawning . . 9
Introduction • 9
Methods 10
Results 19
Discussion 45
4. Zinc Tolerance in Four Generations of Flagfish 52
Introduction 52
Literature review of metal tolerance 52
Objective of this study 53
Methods 53
Results—exposures and calculations 58
Discussion 64
5. Use of Temperature Preference and Activity as a
Sublethal Bioassay for the Toxic Effects of Zinc to the
Bluegill 71
Introduction ..... 71
Design of the study 71
Materials and methods 72
Results 73
Discussion 76
References 78
Appendices
A. Number of Fish Caught at Each Sampling Station, 1976-78 ... 85
B. Marshes Near the Columbia Generating Station ... 88
C. Review of Literature on Entrainment From Cooling
Lake Intake Structures 90
D. Review of Literature on Acid Precipitation 95
E. Review of Literature on Alternative Disposal of Fly Ash . . . 101
vii
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FIGURES
Number Page
1 Map of study site at Columbia Generating Station. ... 2
2 Ground-water flows before and after construction of the
Columbia cooling lake 4
3 Location of fyke nets in spawning marshes on the Columbia site. . 14
4 Fyke-net catches averaged over the 3-yr period, 1976-78 20
5 Spring water levels in the spawning grounds at the
Columbia site during 1976-78 23
6 Catch of northern pike per unit effort on the Columbia site
during 1976-78 24
7 Survival of northern pike eggs hatched -in si.tu in the
wetlands at the Columbia site during April 1977 ........ 27
8 Survival of northern pike larvae placed in the wetland
at the Columbia site for 11 days in April 1977 29
9 Survival of northern pike eggs hatched in the laboratory
during April 1978 in unfiltered water from the ashpit
drain, Rocky Run Creek, and a downstream natural mixture. ... 30
10 Survival of northern pike eggs hatched in the laboratory
in April 1978 using filtered water from the asphit drain,
Rocky Run Creek, and a downstream natural mixture 32
11 ft>pulation-age structure of northern pike caught on
the Columbia site in 1976, 1977, and 1978 35
12 Ibpulation-age structure of walleye caught on the Columbia
site in spring 1977 36
13 Major areas of potential northern pike spawning habitat in the
Wisconsin River and tributaries near the
Columbia Generating Station ... 37
14 Detailed maps (a-g) of potential northern pike spawning
habitats in the Wisconsin River and tributaries near
the Columbia Generating Station . 38-44
viii
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15 Cumulative mortality (probit scale) as a function of exposure
time for flagfish exposed to three zinc concentrations .... 55
16 Procedure used in selecting for zinc resistance in
laboratory populations of flagfish 56
17 Survival rates of selected and unselected flagfish populations
over four generations 59
18 Summary of three generations of selection for zinc resistance
in flagfish 65
19 ffedian selected temperatures of bluegill in control
aquaria and in aquaria treated with zinc in a
7-day experiment . ........... 74
20 Median number of tunnel passes per hour by bluegill in
control aquaria and in aquaria treated with
zinc in a 7-day experiment 75
ix
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TABLES
Number Page
1 Fish Species at the Columbia Generating Station Site 11
2 Water Temperature (°C) at Various Sites in the Spawning Marsh . . 21
3 Size (in Millimeters) of Individual Northern Pike Fry
Captured in Rocky Run Slough, Spring 1976 25
4 Water Quality Data for Various Stations in the Ilarsh—
March 21-May 12, 1977 28
5 Concentrations of Trace Elements in Northern Pike Eggs and Fry
Used in the In Situ Bioassay, Spring 1977 31
6 Water Chemistry for 1978 Laboratory Experiment 33
7 Relationship of Construction Activities to Pike Year-Classes. . . 45
8 Vegetation Types for Various Wetlands Located Along the
Wisconsin River ......... 46
9 Chemical Characteristics of Madison, Wis., Tap
Water in which Flagfish Were Raised, and Dilution Water
in which Zinc Exposures Were Conducted 57
10 Nominal Zinc Concentration (ppm), Length of Exposure (Days),
and Recovery Time Before Breeding (Weeks), for the Zinc
Exposures of the Parental and Three Generations
of Flagfish 57
11 Zinc Concentrations for the Parental, First, Second,
and Third Generation Exposures 58
12 Standard Lengths of Survivors and Nonsurvivors for
Parental-Generation Flagfish Exposed to 0.8 mg/Liter Zinc
for 17 Days 60
13/ Standard Lengths and Wet Weights of Parental-Generation
Flagfish Used to Produce the First Generation 60
14 Spawning Data for Parental-Generation Flagfish 61
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15 Lengths and Wet Weights of First-Generation Fish Used
for Spawning ......... 62
16 Summary of Spawning Data for F^-F^ Generations of Flagfish ... 63
17 Number of Parents Contributing Larvae for Each Generation .... 68
18 Routinely Determined Characteristics of Water Used in the
Temperature-Preference Bioassay ... 72
19 Zinc Concentrations (ppm) of Water in Treatment and Control
Aquaria ..... 73
20 Zinc Tissue Concentrations (ppm) at the End of the
Experiment for Randomly Selected Fish from Treatment
and Control Tanks ....... 76
A-l Number of Fish Caught at Each Sampling Station, 1976 85
A-2 Number of Fish Caught at Each Sampling Station, 1977 86
A-3 Number of Fish Caught at Each Sampling Station, 1978 87
B-l Marshes Near the Columbia Generating Station 88
D-l Summary of pH Effects on Aquatic Organisms 97
XI
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ACKNOWLEDGMENTS
We are grateful to numerous students and workers at the Laboratory of
Limnology of the University of Wisconsin-lladison for assistance in the field
and the laboratory. In particular, we recognize Jeffery Boxrucker, Douglas
Stamra, Steven Voss, Walter Gauthier, Samuel Sharr, Jane Hillstrora, and
Kathryn Webster. Warren Buchanan interpreted the infrared aerial wetland
photographs. Dr. Philip Helmke and his staff at the University of
Wisconsin-lladison Soil Science Department analyzed the trace-metal content
of northern pike fry. Steve Horn prepared many of the figures, including
the wetland vegetation maps. Stephanie Brouwer and Kari Sherman deserve
special mention, for their efforts in editing this report.
Researchers responsible for the three major sections of this report,
under the direction of Dr. John J. Magnuson, were Frank J. Rahel, Michael J.
Talbot, and Anne 11. Forbes (Effects of the Columbia Generating Station on
Fish Spawning); Frank J. Rahel (Selection for Zinc Tolerance Over Four
Generations of Flagfish); and Patricia A. Medvick (Use of Temperature
Preference and Activity as a Sublethal Bioassay for the Toxic Effects of
Zinc to the Bluegill). The literature reviewed in the appendices were
prepared by Anne 11. Forbes, Walter A. Gauthier, Dorothy 11. Harrell, and
Frank J. Rahel.
xii
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SECTION 1
INTRODUCTION
Coal-fired electric generating stations play an important role in
energy production in the United States and are likely to increase in
importance as other fossil fuels become scarce (Gordon 1978). Increased
coal use has serious environmental consequences, and the changes resulting
from power-plant construction and operation are now receiving considerable
attention nationally (Glass 1978). Impact studies have been conducted to
predict and, we hope, to mitigate the negative impacts of coal-fired
generating stations on both terrestrial and aquatic ecosystems. One of
these stations, the Columbia Generating Station near Portage, Wis., has been
the focus of an extensive 3-yr study funded by the U.S. Environmental
Protection Agency (EPA). As part of this multidisciplinary effort, our
research has sought to assess the effects of the station on the local fish
populations. The study has three components: (1) a field study to
determine the importance of the generating station site as a spawning ground
for fish and to assess the station's effects on fish reproduction; (2) a
laboratory study to determine if fish populations can rapidly evolve
tolerance to trace-element contaminants released by coal-fired generating
stations; and (3) a laboratory study to determine whether a bioassay
utilizing temperature preference and activity could detect changes in fish
behavior after exposure to sublethal levels of trace elements.
The Columbia Generating Station is located on wetlands near the
Wisconsin River in south-central Wisconsin (Figure 1). The 1,104-ha site
contains a wide range of plant and animal communities and includes aquatic,
wetland, and forested areas. The site is bordered by Duck Creek on the
north, Rocky Run Creek on the south, and the Wisconsin River on the west.
Of the original acreage, one-half has been altered permanently by the
installation of a 203-ha cooling lake, 28-ha ash basin, coal-handling
facilities, and various other structures. The station has two power-
generating units, Columbia I and Columbia II. Construction of Columbia I, a
527-MW unit with a 152-m boiler chimney equipped with two hot-side
electrostatic precipitators, began in 1971; operation began in April 1975.
Cooling water for the unit is recycled continuously through the cooling
lake. Columbia II, a unit of similar size but with a 198-m stack and sulfur-
removal scrubbers, began operation in March 1978. Cooling towers were built
to remove excess heat from Columbia II to minimize further thermal loading
to the cooling lake. Fly ash and bottom ash produced during operation of
Columbia I are pumped as a slurry into the ashpit where the ash particles
settle out in a series of lagoons. The water is then pumped to the ashpit
drain and eventually reaches Rocky Run Creek. Columbia II adds bottom ash
to the ashpit; all fly ash from Columbia II is disposed of dry.
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•Vi
Figure 1. Map of the study site at Columbia Generating Station.
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A diverse fish community lives in the remaining wetlands, in creeks
bordering the site on the north and south, and in the Wisconsin River west
of the site. Wetlands on the station site are spawning and nursery grounds
for important game fish, including northern pike (Esox lucius), muskellunge
(Esox masquinongy), and walleye (Stizostedion vitreum vitreum) (Ives and
Besadny 1973). Four factors related to construction and operation of the
generating station could potentially affect resident fish populations:
habitat loss, habitat alteration, cooling lake intake, and acid
precipitation.
About 446 ha of the original 1,104-ha site have been altered by
construction of the facility. Much of the habitat lost, including land used
for the cooling lake, was formerly sedge meadow, hence ideal spawning ground
for game fish (Priegel and Krohn 1975). Since fish that use these spawning
grounds are part of the Wisconsin River population, negative effects on the
station site could affect the Wisconsin River fishery. The magnitude of
these effects would depend on the availability of alternate spawning sites
and the possible involvement of unique homing stocks.
Habitat alterations in the remaining wetlands continue because of
increased ground-water discharge and waste heat from the cooling lake
(Bedford 1977). Before construction of the facility, the upland sloped
gradually to the flood plain sedge meadow and the Wisconsin River. Ground-
water flow of 1 ft /s from adjacent uplands maintained the sedge meadow
water levels. The flow varied seasonally, being high during spring and lower
during the summer. The cooling lake established a 9-ft hydrostatic head
above water levels in the adjacent wetlands and drastically altered the
natural ground-water pattern (Figure 2). Now warm water from the cooling
lake at a flow of 4 ft /s seeps west into the sedge neadow. The seepage of
warm ground water into the rooting zone of plants and the associated rise in
water level resulted in dramatic vegetation changes. The area west of the
cooling lake, formerly dominated by perennial sedges, is being transformed
into a community dominated by annuals and hydrophytic perennials such as
cattails. Sedges offer dense mats of vegetation for spring spawning fish,
but the annuals and hydrophytic perennials generally die down each winter
and provide little suitable spawning substrate.
The warm-water seepage into the meadow west of the cooling lake is
masked by volumes of Wisconsin River flood water in spring and does not
directly affect spawning fish or eggs. However, the ashpit drain (east and
south of the cooing lake) also receives cooling lake seepage and remains
several degrees warmer even in the spring.
Habitat alterations in the wetlands south of the site along Rocky Run
Creek are also caused by effluent from the ashpit. Metal oxides that
compose the major reactive portions of the ash cause the pH of the ashpit
water to rise 10 to 11 (Andren et al. 1977). Since Wisconsin water quality
standards prohibit the release of water at a pH above 8, sulfuric acid is
added before the ash effluent is discharged. The addition of acid causes
the precipitation of elements such as barium and aluminum into a floe that
coats the bottom of the ashpit drain and flows into Rocky Run Creek. Thus,
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River
UJ
River *-Studv Area
Cooling Lake
CO
UI
500
1000
METERS
1500
2000
Figure 2. Ground-water flows before and after construction of the
Columbia cooling lake. Arrows represent integrated flows
1 m3/min, normal to the east-west cross section along the
length of the cooling lake (Andrews and Anderson 1980).
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dissolved and particulate trace elements, fly-ash particles, and perhaps
organic contaminants in the ash effluent continually enter these streams,
which then flow into the extensive Rocky Run wetland area. Many fish
species, including the sensitive early life-history stages of important game
fish, live in these waters and are exposed to the effluent.
Although the generating station recycles cooling water through the
cooling lake, evaporative losses are made up by using Wisconsin River
water. Since water and any organisms withdrawn from the river are not
returned, the plant is analogous to a predator. Fish loss due to
impingement and entrainment depends on the volume of water removed, the
patchiness of fish distribution, and the ability of fish to avoid
entrainment. A 1-yr study of egg and larval fish entrainment and juvenile
and adult fish impingement at the Columbia site (Swanson Environmental, Inc.
1977) reported insignificant numbers of fish losses. The total river flow
removed by the intake water at Columbia presently averages 0.3%, with a
maximum of 1.08%. As long as the Columbia intake continues to remove a
small percentage of the river flow, we expect no measurable effects of
entrainment on the river system. An exception might occur when organism
distribution is patchy near the intake, and a significant portion of one
year-class (e.g., walleye larvae) is entrained. Aside from acting as a
predator by removing organisms from the Wisconsin River, the usual types of
entrainment effects (mechanical, toxic, and thermal) do not apply to the
Columbia station, since organisms and water are not returned to the river.
Loss of fish populations due to lake acidification is related directly
to acid precipitation from fossil-fuel combustion (Gorham 1976). The
likelihood of waters undergoing acidification depends on the edaphic
characteristics of their drainage basins and the intensity of acid input.
Waters near the Columbia station site have a high buffering capacity because
the drainage basin is calcareous. Although the problem has not been studied
directly at the Columbia site, predictions based on the extensive literature
indicate little potential for acid precipitation damage to the aquatic
systems near the Columbia station.
DESIGN OF THE FISH STUDY
The impacts of the four above factors on fish populations were
investigated through both the field and laboratory components of this study,
as well as by reference to existing literature. Some of the concerns are
site-specific; others are more general and therefore applicable to other
locations. Overall conclusions are presented in Section 2.
Section 3 documents the site-specific effort to determine the
importance of the generating station as a spawning ground for northern pike,
muskellunge, and walleye. Spring sampling of these populations yielded
information on the use of the marsh for spawning as well as on year-class
strengths related to both natural and artifical changes. The effect of the
ashpit effluent on fish reproduction is discussed in Section 3, as is the
relative importance of the spawning marsh to the total Wisconsin River
fishery.
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Sections 4 and 5 concern questions that may arise at other generating
sites. Section 4 describes the methods and results of the metal-tolerance
study, in which flagfish were bred in the laboratory for resistance to
zinc. Section 5 deals with the use of temperature preference and activity
as a bioassay to detect subtle changes in fish behavior after exposure to
sublethal metal levels.
Appendices include reviews of the literature on entrainment, acid rain,
and fly-ash disposal.
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SECTION 2
CONCLUSIONS
The major factors affecting fish at the Columbia site are habitat loss
and habitat modification. Construction of the Columbia Generating Station
eliminated approximately 18% of the shallow-water sedge meadow located
between the Prairie du Sac and Wisconsin Dells dams. This habitat, when
inundated by spring floods, is utilized by spawning northern pike and
muskellunge. The station site currently contains about 22% of the deep-
water sedge meadow and 0.8% of the shallow-water sedge meadow likely to be
used by spawning northern pike in this section of the Wisconsin River.
Tagging efforts indicated that northern pike from as far away as Lake
Wisconsin (17 km downstream) migrate to the Rocky Run wetlands to spawn.
However, these wetlands continue to be affected by effluent from the ashpit
drain and by underground seepage of warm water from the cooling lake.
Northern pike, muskellunge, and walleye spawned in areas affected by
ashpit effluent; in fact, northern pike were apparently attracted to the
ashpit drain because of warmer spring water temperatures and higher current
speeds.
The presence of ripe and spent northern pike adult spawners in 1976-78
indicated that spawning occurred in areas affected by the ashpit effluent.
The presence of newly hatched fry in 1976 indicates that reproduction was
successful.
Analysis of northern pike year-class strengths suggested that the 1976
year-class (the first year-class affected by the plant's operation) may be
reduced. Further monitoring of population structure is warranted since
year-classes hatched after the station began operation are now reaching
maturity and should be returning to spawn.
In a laboratory bioassay, hatching success of pike eggs incubated in
unfiltered ashpit drain water was lower than for eggs raised in filtered
ashpit drain water; therefore, the flocculent precipitate found in ashpit
drain water appears to hinder pike egg development. When the flocculent
precipitate was removed by filtering, eggs hatched equally well in ashpit
drain water and Rocky Run Creek water.
Fry hatched in the ashpit drain contained elevated levels of only one
element, sodium, compared to fry hatched at other locations in the marsh.
Acute toxicity due to trace-element bioaccumulation is unlikely to be a
problem for fish eggs or newly hatched fry in the ashpit drain.
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In the metal-tolerance study the selected population had a higher
resistance to zinc after the first two generations, but did not differ from
the unselected population in the third generation. Possible explanations
for the failure of selection to continually increase zinc tolerance include
inbreeding depression and carry-over effects passed from mother to offspring
through the egg cytoplasm. After 2 to 8 weeks recovery time, flagfish which
showed a zinc exposure lethal to the majority of the population, reproduced
as successfully as unexposed fish.
The temperature preference and activity apparatus tested in the
sublethal zinc bioassay is no more sensitive than bioassay methods used by
previous investigators.
Potential damage to Wisconsin River fish and invertebrate populations
from entrainment or impingement on water intake structures appears minimal.
Acid precipitation is not considered a potential problem for aquatic
ecosystems at the Columbia site because of the high hydrogen-ion buffering
capacity resulting from the calcareous nature of the drainage basin.
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SECTION 3
EFFECTS OF THE COLUMBIA GENERATING STATION ON FISH SPAWNING
INTRODUCTION
Construction of a coal-fired electric generating station on a flood
plain of the Wisconsin River (Columbia County, Wis.) has resulted in
alteration of an important fish-spawning habitat. The station site formerly
contained extensive wetland areas, particularly during spring floods when
many fish species migrate to such areas to spawn on inundated vegetation.
Among those species known to have used the Columbia Generating Station site
for spawning are northern pike (Esox lucius), muskellunge (Esox
masquinongy), and walleye (Stizostedion vitreum vitreum). Construction of
the station permanently altered about one-half of the original 1,104 ha;
much of the wetland affected was sedge meadow, an ideal spawning habitat for
these fish. Portions of the remaining wetlands and Rocky Run Creek have
undergone physical and chemical changes that may also influence fish
reproduction. Given these considerations we undertook a site-specific study
to determine both the importance of the station site as a spawning ground
for Wisconsin River game fish and the effect of the power plant on the
reproductive success of these fish. Our study involved the following: (1)
netting on the site to discover which areas were important spawning grounds;
(2) both -in situ and laboratory bioassays to assess the effects of ashpit
effluent on hatching and larval survival; (3) aging of scale samples to
relate year-class strengths to both natural and artificial changes in the
environment; and (4) use of infrared aerial photography to compare the
spawning habitat at the station site to the total spawning habitat available
in this section of the Wisconsin River.
One of the most obvious effects of the generating station is the
introduction of ashpit effluent into potential spawning areas. This
effluent is pumped from the ashpit settling basins into the ashpit drain.
The ashpit drain joins a creek that flows through portions of the wetlands
adjoining the site and then enters Rocky Run Creek 1.5 km above its mouth at
the Wisconsin River (Figure 1). The effluent waters contain elevated levels
of some trace elements and other coal-combustion by products (Andren et al.
1980, Helmke et al., Unpublished). Beginning in January 1977 sodium
bicarbonate was routinely added to the pulverized coal to increase the
efficiency of the electrostatic precipitators. This treatment resulted in
increased conductivity in the ashpit drain and Rocky Run Creek and, in fact,
served as a useful tool for measuring ash-effluent concentration downstream
from the generating station. The water upstream of the ash effluent in the
mint drain and in Rocky Run Creek is usually higher in alkalinity, hardness,
and pH and lower in turbidity (Magnuson et al. 1980).
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Invertebrate populations in the ashpit drain decreased in abundance and
species composition after generating station operation began (Magnuson et
al. 1980). Schoenfield (1978) demonstrated a tenfold increase in chromium
and barium concentrations in several taxa of these ashpit-drain
invertebrates (especially Asellus and Hyalella) since the start of station
operation. The ashpit drain itself has a very sparse resident fish
population (ictalurids and some cyprinids) probably as a result of
inhospitable chemical and physical conditions as well as a lack of food
organisms. Studies of another ashpit-drain system also documented a
depauperate invertebrate fauna and the presence of only one fish species
(Gambusia affinis) (Cherry et al. 1976).
The Rocky Run Creek area has a diverse fish community and is an
important site for northern pike, muskellunge, and walleye. These species
spawn on flooded wetlands where the newly hatched larvae remain for up to
several months before emigrating to nearby rivers or lakes. A
preoperational impact statement completed by the Wisconsin Department of
Natural Resources (Ives and Besadny 1973) documented the occurrence of 47
fish species on or near the station site (Table 1). Included was one
species listed as threatend in Wisconsin, the mud darter (Etheostoma
aspirgene). Eggs and fry of both walleye and northern pike were also
collected on the station site during the preoperational study.
METHODS
We conducted extensive fyke netting on flooded wetlands accessible from
Rocky Run Creek from 1976 to 1978 to determine which areas were important
spawning grounds, to see if fish bypassed the station site to spawn, and to
assess year-class strengths of northern pike, the predominant species. By
tagging fish we hoped to document the areas from which fish traveled to
reach the spawning grounds and to determine whether homing to a certain site
was recurring in successive years.
After discovering that pike were spawning in areas receiving ashpit
effluent, we compared reproductive success in those areas with reproductive
success in nearby unaffected control areas. Collection of pike fry in the
field to study the natural hatch was unsuccessful; hence, we conducted both
an in situ and a laboratory bioassay to assess the effects of ashpit
effluent on egg hatching and larval survival. Water from both affected and
unaffected areas was used, and survival was monitored daily. Northern pike
fry hatched in various locations in the marsh were analyzed for trace
elements to assess uptake in areas affected by the ashpit effluent.
After spawning, fish return to a 63-km section of the Wisconsin River
bordered upstream by a dam at Wisconsin Dells and downstream by a dam at
Prairie du Sac which forms Lake Wisconsin. During spring floods in this
stretch of the river, many other backwater areas are formed that could serve
as pike spawning grounds. The relative importance of spawning habitat on
the plant site to the total spawning habitat available in this section of
the river was determined by infrared aerial photography. Photo-
interpretation allowed us to classify wetlands into various types according
to their value as pike-spawning grounds and to calculate the acreage of each
-10-
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TABLE 1. FISH SPECIES AT THE COLUMBIA GENERATING STATION SITE
Common name
Scientific name
Collection
sites3
Reference
Bigmouth buffalo
Black bullhead
Black crappie
Blacknose shiner
Blackstripe topminnow
Bluegill
Bluntnose minnow
Bowfin
Brook silversides
Brown bullhead
Brown trout
Bullhead riinnow
Carp
Central mudminnow
Channel catfish
Chestnut lamprey
Emerald shiner
Fathead minnow
Flathead catfish
Freshwater drum
Golden redhorse
Golden shiner
Grass pickerel
Green sunfish
Johnny darter
Lake sturgeon
Largemouth bass
Mooneye
Mud darter
Muskellunge
Ictiobus cyprinellus
Ictalurus melas
Pomoxis nigromaculatus
Notropis heterolepis
Fundulus notatus
Lepomis macrochirus
Pimephales notatus
Amia calva
Labidesthes sicculus
Ictalurus nebulosus
Salmos trutta
Pimephales vigilax
Cyprinus carpio
Umbra limi
Ictalurus punctatus
Ichthyorayzon castaneus
Notropis atherinoides
Pimephales promelas
Pylocictis olivaris
Aplodinotus grunniens
Moxostoma erythrurum
Notemigonus chrysoleucas
Exox americanus vermiculatus
Leponis cyanellus
Etheostoma nigrum
Acipenser fulvescens
Micropterus salmoides
Hiodon tergisus
Etheostoma aspirgene
Esox masquinongy
C 2
B,C,D 1,2
A,B,C,D 1,2
B 1
B 1
A,B,C,D 1,2
B 1
A,B,D 1,2
B 1
B,D 1
C 2
B 1
A,B,C,D 1,2
B 1
A,B 1
C 2
B 1
B 1
A 1,2
A,B,C,D 1,2
B,D 1
B,D 1
B 1
B 1
B 1
A 1
B 1,2
A 1
B 1
A,B,C 1,2
(continued)
-11-
-------
TABLE 1 (continued)
Common name
Northern pike
Pirate perch
Pumpkinseed
Quillback
Rainbow trout
Redhorse
Red shiner
Rock bass
Sand shiner
Sauger
Shovelnose sturgeon
Smallmouth bass
Smallmouth buffalo
Spotfin shiner
Spotted sucker
Tadpole madtom
Walleye
White bass
White crappie
White sucker
Yellow bullhead
Yellow perch
i
Scientific name
Esox lucius
Aphredoderus sayanus
Lepomis gibbosus
Carpiodes cyprinus
Salmo gairdneri
Moxostoma sp.
Notropis lutrensis
Ambloplites rupestris
Notropis stramineus
Stizostedion canadense
Scaphirhynchus platorynchus
Micropterus doloraieui
Ictiobus bubalus
Notropis spilopterus
llinytrema melanops
Noturus gyrinus
Stizostedion vitreum vitreum
Morone chrysops
Poraoxis annularis
Catostomus coranersoni
Ictalurus natalis
Perca flavescens
Collection
sites3
A,B,D
B,C
A,B,D
B
C
C
B
A,B,D
B
A,B,C,D
A
A,B
C
B
B
B
A,B,C,D
A,B
A,B,D
B,D
A,B,D
A.B.C.D
Reference
1,2
1,2
1,2
1
2
2
1
1,2
1
1,2
1
1
1
1
1,2
1
1,2
1,2
1,2
1,2
1,2
1,2
aCollection sites include: A—Wisconsin River, B—Duck Creek, C—Rocky Run
Creek, and D—On-site flooded areas.
List compiled by investigators from (1) the Wisconsin Department of
Natural Resources and (2) the University of Wisconsin-Iladison.
-12-
-------
type. By comparing various watersheds we could determine what percentage of
the total available pike spawning habitat in that section of the Wisconsin
River was being influenced by the generating station.
Scale samples from pike, muskellunge, and walleye were aged in an
effort to relate year-class strengths to both natural and man-caused changes
in the environment. The introduction into spawning areas of the ashpit
effluent, which had a complex and largely unknown chemical make-up, was one
such change. Construction activities, which began in January 1971,
destroyed a large portion of the sedge meadow and undoubtedly influenced
spring flood patterns. Studies have shown that flooded vegetation,
preferably dense mats of sedges (Carex sp.), is necessary for successful
pike and muskellunge reproduction (McCarraher and Thomas 1972, Priegel and
Krohn 1975). Since the generating station began operation, seepage under
the cooling lake dikes has modified water temperatures and vegetation types
in the adjacent wetlands (Bedford 1977). The general trend has been toward
replacement of shallow-water marsh dominated by the sedge (Carex lacustris)
and other perennials to deep water dominated by annuals and hydrophytic
perennials such as arrowhead (Sagittaria sp.) and cattail (Typha sp.).
These changes reduced the amount of densely matted vegetation available for
spring spawning fish. Finally, one natural factor that strongly influences
reproductive success is the timing and extent of spring flood levels.
Johnson (1956) found a direct correlation between high spring water levels
followed by a small decline in levels during egg incubation and production
of a strong northern pike year-class.
With the above considerations in mind, our overall goal was to
determine the importance of the generating station site as a spawning ground
for spring-spawning game fish and to assess the plant's effects on the
reproductive success of these fish. More specific information on methods is
included in the following sections.
Survey of Spawning Grounds
To sample fish species that spawn on the site, fyke-net sites were
established at five strategic locations (Figure 3) and checked regularly
from ice-out in late February through the end of the spawning season in late
April. All fish captured during upstream movements were released on the
upstream side of each net. A description of each site and the years that
the nets were worked follows.
Net 1, worked between 1976 and 1978, was located on Rocky Run Creek at
the bridge on County Trunk Highway JV. This site, upstream from the plant,
was unaffected by construction or operation of the generating station. This
net completely crossed the river channel, except for peak floods, and caught
fish that passed by the station site. Water depths ranged from 1.5 to 3 m,
depending on flood stage, and the current was generally strong. Spawning
habitat was available immediately above the net when the creek flooded
surrounding marshland.
Net 2, worked between 1976 and 1978, was located in Rocky Run Creek
downstream of its confluence with the ashpit drain. The net was placed in a
-13-
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1.5 km
Figure 3. Location of fyke nets in spawning marshes on the Columbia site.
(* on insert shows the location of the Columbia Generating
Station in south-central Wisconsin.)
-14-
-------
channel between two old bridge abutments and, except for peak flood periods,
captured most fish entering the marsh system. Water depth ranged from 1.5
to 3 m depending on flood stage, and the current was moderate. Fish passing
this point might spawn in the habitat immediately above the net, or they
might proceed to spawning areas near nets 1, 3, or 4. Ashpit effluent was
well mixed with Rocky Run water by the time it reached net 2.
Net 3, worked between 1976 and 1978, was located at the northern end of
the Rocky Run Creek backwaters and in the main current channel leading into
the sedge meadow near the cooling lake. This net caught fish that migrated
into wetlands beyond those affected by ashpit effluent. Passage around the
net was often possible when the entire area was flooded, and generally water
levels dropped sharply at this site after the spring floods. Although
spawning habitat was available near the net, it was mainly located upstream
in flooded sedge meadows.
Net 4, worked between 1976 and 1978, was located in the ashpit drain
above its junction with Rocky Run Creek in a diked channel that caught most
of the fish migrating upstream. This net caught fish entering spawning
areas most affected by ashpit effluent. The current was moderate and water
levels were generally 0.5 to 2 m. Unlike net site 3, this site retained a
flow of water throughout the year. Abundant spawning habitat was available
below and above this net. Clumps of the precipitated floe described earlier
were often found floating downstream, and finer floe particles could be
observed in the water column.
Net 5, worked in 1978, was located above the confluence of the mint and
ashpit drains, and was used to determine if fish would migrate all the way
up the ashpit drain to reach spawning ground unaffected by ashpit
effluent. It completely sealed off a small channel that was usually less
than 1 m deep but contained water throughout the year.
The Rocky Run slough areas were electroshocked in 1976 and 1977 to
determine if many fish were avoiding the nets, but still utilizing the site
for spawning. Areas downstream from net 2 were shocked also to determine
their importance as spawning grounds. Because of low water levels and
technical problems, we found that our fyke nets were a more effective method
of collecting fish.
Northern pike, muskellunge, and walleyes were measured, sexed, and
fitted with a monel tag, from the National Band and Tag Co., inserted into
their preopercular bone. The tag carried an identification number as well
as the label "University of Wisconsin-Madison." To determine the
distribution of these fish during the nonspawning season, we posted notices
at various boat landings and fishing areas offering a reward to anglers who
returned tags from fish they caught.
To estimate spawning success on the plant site, we set fry traps in
1976 and 1977 in areas corresponding to net sites 2, 3, and 4. The traps
were generally ineffective, as were light traps tried in 1978. After study
of the natural hatch proved unfeasible, we conducted field and laboratory
-15-
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bioassays to assess the effect of the ashpit effluent on northern pike egg
and fry survival.
Effects of Ashpit Effluent on Reproductive Success
After documenting the use of areas receiving ashpit effluent as
spawning grounds in 1976, we conducted an in s~itu bioassay in 1977 to
determine if the effluent was affecting survival of eggs and larvae of
northern pike, the predominant game species in the area. Fertilized eggs
were obtained from the Wild Rose Fish Hatchery of the Wisconsin Department
of Natural Resources. Eyed eggs were incubated in the main stream channel
at three locations: (1) in the ashpit drain; (2) above the plant site in
Rocky Run Creek; and (3) downstream Rocky Run Creek after ashpit effluent
water was mixed well with creek water. These locations corresponded to net
sites 4, 1, and 2, respectively.
Ten incubation bottles, each containing approximately 250 eggs, were
anchored at each location. Incubation bottles were 1-gal plastic containers
with plastic screen on the sides and bottom, current deflectors to maintain
water exchange, and floats to keep them at the surface. Each day we emptied
accumulated silt, counted live eggs and larvae, and observed the stage of
egg development. Dead eggs were removed to minimize disease. Water
temperatures at each station were recorded continuously with a Ryan
Instruments temperature recorder, and relative siltation loads were
estimated from silt deposition in graduated cylinders over 24-h periods.
Dissolved oxygen, conductivity, water current speeds, pH, alkalinity, and
EDTA hardness were measured regularly with standard techniques (American
Riblic Health Association et al. 1975). Survival was calculated daily as
the percentage of stocked eggs still as eggs or fry. The experiment was
repeated with newly hatched pike larvae, also obtained from the Wisconsin
Department of Natural Resources.
Northern pike fry that hatched in our incubation bottles were analyzed
for 20 trace-element concentrations by neutron activation. The analyses
were performed by the Trace Elements Subproject under the direction of Dr.
Riillip Helmke (Helmke, unpublished). Fry hatched in areas affected by the
ashpit effluent were compared with fry from unaffected waters. Since the
ashpit effluent was known to contain elevated levels of certain trace
elements, we felt it important to document any biological accumulation by
this sensitive life-history stage.
At all three locations, the incubation bottles accumulated silt, but
the problem was greatest at the Rocky Run stations where the daily
accumulation of fine dark silt often completely covered the eggs. Although
sediment also accumulated in the ashpit drain bottles, most of it consisted
of the semibuoyant chemical floe. Ashpit drain eggs remained amber in
color, whereas Rocky Run eggs were coated with brown silt. In nature, eggs
are spawned over a mat of sedge grasses and do not become buried by
sediments. Therefore, we repeated the experiments in 1978 under controlled
laboratory conditions where siltation effects could be eliminated.
-16-
-------
Egg cups made from polyinylchloride (PVC) piping covered at the bottom
with plastic screen were used to hatch eggs in a laboratory chamber
maintained at 12°C. Ten cups of approximately 100 eggs each were incubated
in individual 500-ml beakers containing ashpit drain water, Rocky Run water,
or the naturally occurring mixture. A similar set of egg cups was exposed
to water from these same sources that had been passed through a 3-ym filter
to eliminate sediment or floe, or both. Each lot of eggs was given a fresh
exchange of their respective test water daily, and dead eggs and fry were
removed. The experiment proceeded until all eggs either hatched or died.
Year-Class Strengths
Northern pike, muskellunge, and walleye were aged by counting annuli on
their scales (Williams 1955). Scales were collected from above the lateral
line below the dorsal fin. After cleansing in a solution of 0.1 N NaOH,
impressions of the scales were made on cellulose-acetate slides. These
impressions were then projected onto a screen, and the annuli were
counted. Year-class strengths were then related to factors such as water
levels or site construction activities that might have influenced
reproductive success.
Importance of the Station Site to the Wisconsin River Fishery
An inventory of potential pike spawning areas in the section of the
Wisconsin River bordered by the Wisconsin Dells and Prairie du Sac dams was
prepared by manual photo-interpretation of 70-mm color infrared
transparencies taken from an altitude of 11,000 ft. A series of 180 images
along 12 flight lines was taken during low water levels on 9 September 1977
at wetland at a .scale of 1:62,500. Flights took place in the fall when
maximum vegetation was exposed.
The first step in the analytical process was preparation of a base map
on which to trace the photo-interpretations. Because the photoscale
approximated that of the standard U.S. Geological Survey (USGS) (15-min)
topographic quadrangle maps, transparent mylar copies of the USGS maps were
used as base maps. Each photograph was placed under the transparent base
map, and vegetation areas were transferred to the base map.
An experienced photo-interpreter identified the resources on the
imagery. Ground checking of mapped wetland classes showed an excellent
correlation between photo-interpretation and actual vegetation types. From
the infrared photographs the following wetland classifications were
delineated:
1) Deep-water sedge meadow. The deep-water sedge meadows consist of
lake sedge (Carex lacustris), bullrush (Scirpus spp.), and bluejoint grass
(Calamagrostis canadensis and C^. inexpansa); some colonies of cattail (Typha
sp.) are mixed in. Other floating and emergent aquatic macrophytes that
often occur in deep-water sedge meadows include arrowhead (Sagittaria spp.)
and burreed (Sparganium eurycarpum). A deep-water sedge meadow is
characterized by inundation throughout the year.
-17-
-------
2) Shallow-water sedge meadow. This category is dominated largely by
either tussock sedge (Carex stricta) or bluejoint grass. Forbs such as
aster (Eupatorium spp.) are occasionally prominent. Cattail is fairly
common, as is reed canary grass (Phalaris arundinacea). Shallow-water sedge
meadow, as the name implies, has a drier moisture regime than deep-water
sedge meadow. For long periods during the growing season the water table is
somewhat below the surface.
3) Emergents and deep-water floating macrophytes. This category
includes deep-water species floating or barely emerging above the surface.
Floating mats of duckweed (Lemna spp.) frequently cover stagnant open water
between other vegetation. Large mats of arrowhead (Sagittaria spp.), water
lily (Nymphaea spp.) and (Nuphar spp.), or pondweed (Potamogeton spp.) often
compose this vegetation type, which is often associated with the borders of
open bodies of water.
4) Cattail. Monospecific stands of cattail are separated on the maps
if large enough to outline.
5) Reed canary grass. This densely grown monospecific vegetation type
often occurs along water courses or in regularly shaped plantations.
6) Bluejoint grass. This species seldom occurs in large pure stands
and is an indication of a fairly undisturbed wetland community.
7) Drained wetland. A growth of weedy forbs and shrubs often
indicates ditching and desiccated conditions. Typically, remnants of the
original vegetation type are found under the weeds.
8) Shrub carr and swamp. Dogwood (Cornus stolonifera), spiraea
(Spiraea alba), or willow (Salix Interior) are the most common constituents
of shrub carr. Lowland tree species of willow, silver maple (Acer
saccharinum), aspen (Populus tremuloides), or green ash (Fraxinus
pennsylvanica) are commonly interspersed or occur as solid stands of lowland
forest.
9) Mixed vegetation. Vegetation complexes too detailed to separate by
individual boundaries were combined and labeled by predominant types. For
example, complexes of shrub-cattail-sedge meadow were fairly common.
Areas for each vegetation class were quantified bu using a Hewlett
Packard Model 9107A digitizer and calculator. Comparison of pike spawning
areas (viz., deep- and shallow-water sedge meadows) on the plant site with
other potential spawning areas allowed us to draw some conclusions about the
importance of the plant site to the Wisconsin River fishery.
-18-
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RESULTS
Survey of Spawning Grounds
Our fyke nets caught a total of 21 species of fish over a 3-yr period
(Appendix A). Northern pike was the dominant game fish, but walleye,
muskellunge, and largeraouth bass were caught along with many carp and
various catostomid species. Trie average yearly catch (Figure 4) was largest
at the entrance to the Rocky Run slough area (net 2) and smallest in upper
Rocky Run (net 1). No fish were caught in the mint drain (net 5). The low
number of fish caught at nets 1 and 5 indicates that most fish entering the
system remain on the station site to spawn. Furthermore, our catches in the
ashpit drain (net 4) were larger than those at the access point to the sedge
meadow (net 3), indicating that many northern pike move into the ashpit
drain during spawning migrations.
Water temperatures at each net location indicated that ashpit drain
water was consistently warmer by several degrees centigrade than water
draining the sedge meadow adjacent to the cooling lake (Table 2). The
greater number of northern pike spawning near the ashpit drain may be
attributed to the pike's affinity for warmer water currents during migration
(Johnson 1956, Franklin and Smith 1963). Fish entering the Rocky Run slough
through main current channels encounter an intersection where cooler water
flowing through net 3 mixes with warmer water originating in the ashpit
drain and flowing through net 4. Preference for the warmer water currents
may lead fish close to the ashpit drain.
In addition to being warmer, the ashpit drain had a markedly higher
conductivity in 1977 and 1978 than the other stations. This condition was
due to the use of sodium bicarbonate in the coal to increase the efficiency
of the electrostatic precipitators. For example, February-April 1978
averages for conductivity at 25 C were 468 ymhos/cm at net 1, 658 at net 2,
805 at net 3, and 1,218 at net 4. The conductivity decreases downstream,
but is still elevated at the mouth of the Rocky Run slough (net 2).
Effect of Spring Water Levels on Spawning—
Spring water levels also appear to influence the distribution of
spawning fish. In 1976 and 1978 spring flood periods were normal, but in
1977 flooding was greatly reduced (Figure 5). Thus, although large numbers
of spawning northern pike enter the slough area each year, they proceed to
spawning sites near the cooling lake (net 3) or in the ashpit drain (net 4)
only if water levels are high enough to provide suitably flooded vegetation.
In 1977 such vegetation was unavailable, and pike remained in marshy areas
just above the entrance to the Rocky Run slough area (net 2) or moved
further upstream in Rocky Run. This situation is indicated by a reduced
catch per unit effort (= total number of pike caught per number of days net
was set) for nets in the sedge meadow and ashpit in 1977 compared to other
years, but an increased catch per unit effort for the net in upstream Rocky
Run (net 1) (Figure 6).
-19-
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AVERAGE FISH CATCH
SEDGE
MEADOW
Esocidae (2 spp.)
Catostomidae (3 spp)
Other (14 spp.)
ASH PIT
DRAIN
LOWER
ROCKY RUN
UPPER
ROCKY RUN
Figure 4. Fyke-net catches averaged over the 3-yr period, 1976-78. Upper
Rocky Run corresponds to net 1 in Figure 3, lower Rocky Run to
net 2, sedge meadow to net 3, and ashpit drain to net 4. Circles
are scaled to indicate relative abundance at each net site.
-20-
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TABLE 2. WATER TEMPERATURE (°C) AT VARIOUS SITES IN THE SPAWNING HARSH
1976
March
April
May
1977
March
5
13
15
16
17
18
19
20
21
22
23
24
25
26
27
28
8
10
13
15
20
22
23
28
30
3
5
7
12
2
4
7
9
11
14
Upper
Rocky Run
(Net 1)
2.8
3.4
1.7
2.0
4.7
4.5
4.8
3.4
3.1
6.0
7.0
9.2
6.3
10.4
10.2
8.8
12.5
11.5
9.4
10.1
5.6
12.7
7.7
0.5
4.0
6.0
Lower
Rocky Run
(Net 2)
2.5
3.6
1.3
2.2
5.0
6.6
5.4
3.6
3.6
3.5
7.0
7.2
8.5
6.3
6.8
11.5
11.8
14.2
10.2
12.4
11.0
13.2
6.5
13.5
10.3
0.4
2.0
4.0
6.1
9.4
6.5
Access to
sedge meadow
adjacent to
cooling lake
(Net 3)
1.9
1.0
1.8
3.2
3.6
7.7
3.4
5.0
3.1
5.8
6.1
6.5
4.5
5.1
10.5
9.9
10.6
15.1
11.6
10.4
12.5
10.5
12.1
5.8
13.7
8.4
13.5
Lower
ashpit
drain Hint drain
(Net 4) (Net 5)
8.0
3.7
4.2
7.5
8.6
10.4
3.6
7.5
4.5
7.2
6.7
10.6
6.5
7.7
15.5
11.0
11.4
13.4
12.8
10.1
12.7
10.5
10.4
4.9
14.2
10.2
O • U f — 4. 4 — — . J\
-21-
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TABLE 2 (continued)
Upper
Rocky Run
(Net 1)
April
1978
March
April
16
18
21
23
24
28
30
4
5
7
2
3
6
8
10
13
15
17
20
22
24
27
29
30
31
3
6
8
10
12
14
16
6.
4.
5.
6.
5.
8.
8.
3.
4.
10.
5.
1.
8.
8.
13.
7.
6.
7.
8.
8.
11.
11.
5
0
0
0
0
2
0
0
5
0
3
5
0
2
7
9
8
7
8
9
0
0
Lower
Rocky Run
(Net 2)
7
4
6
6
7
9
9
2
3
9
2
1
0
1
0
1
1
1
3
5
2
5
10
6
12
9
7
8
9
8
11
12
•
•
*
•
•
•
•
•
•
•
•
•
*
•
•
•
•
•
•
•
•
•
•
•
•
•
•
•
•
•
•
•
0
0
0
5
7
2
7
2
9
0
8
0
2
0
9
8
0
8
0
0
0
8
1
9
5
9
5
1
7
3
1
0
Access to
sedge meadow Lower
adjacent to ashpit
cooling lake drain Ilint drain
(Net 3) (Net 4) (Net 5)
5.7 10
2.5 7
4.0 7
9.0 10
0.5 3
3.8 6
16.0 14
8
9
3
9.5 11
7.9 16
9.8 11
9.0 8
8
8.9 9
8
12
13
.0
.5
.0
.0
.5
.5
.0
.0
.0
.5
.0
.1
.5
.5
.0
.0
.3
.5
.0
14.1
8.2
6.5
6.9
8.1
-22-
-------
Figure 5. Spring water levels in the spawning grounds at the Columbia site during 1976-78.
-------
7.0
5.0
3.0
z
ID
•x.
I
u
u
z
o
o
liO
0.5
0.1
0.05
1976 77 78
LOWER
ROC K Y
RUN
76 77 78
UPPER
ROC K Y
RUN
76 77 78
SEDGE
ME A DOW
76 77 78
A S H PIT
DRAIN
LOCATION AND YEAR
Figure 6. Catch of northern pike per unit effort on the Columbia site during
1976-78. Upper Rocky Run corresponds to net 1 on Figure 3, lower
Rocky Run to net 2, sedge meadow to net 3, and ashpit drain to net
4. Catch per unit effort was calculated by dividing the total
number of pike caught by the total number of days a given net was
worked.
-24-
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Capture of Pike Fry—
A few northern pike fry were captured in the Rocky Run slough area in
1976. This indicates that eggs were hatching and larvae were surviving in
these wetlands (Table 3). About one-third of the fry were caught in the
ashpit drain where waters are most affected by ashpit effluent. Small
sample sizes precluded any comparisons of growth rates or estimates of fry
abundance between sampled areas.
TABLE 3. SIZE (IN MILLIMETERS) OF INDIVIDUAL NORTHERN PIKE FRY
CAPTURED IN ROCKY RUN SLOUGH, SPRING 1976
Date
captured
5 May
7 May
10 May
12 May
14 May
17 May
19 May
21 May
24 May
Total
Location captured3
Net 2 Net 3
32,30
34 25,27,28
48 35,35
50,60,70,72 35
60,70
31 43,43
11 8
Net 4
27
42
30,47
30,33,35,35,44,46,47
38
12
aNet 2 was located near the mouth of Rocky Run slough; net 4 was located in
the ashpit drain; net 3 was located in the main current channel draining
the sedge meadow adjacent to the cooling lake (see Figure 3).
No fry were captured in 1977, however. Following the spawning period
that year, 12 fry traps were placed in the marshes and were fished from 1
May to 10 June 1977. Seven traps duplicated the 1976 effort, and five more
were set at other locations in the marsh. No fry of northern pike, walleye,
or muskellunge were caught. Light trapping on two nights in June 1978 also
failed to produce fry.
Because we caught no pike fry in 1977, we suspected a poor year-
class. However, the 1977 year-class was represented in our 1978 fyke-
netting efforts (refer to Figure 11). Comparisons between year-classes
based on the low numbers of fry caught.
Tag Returns—
During the 3 years of fyke netting, 208 northern pike, 9 muskellunge,
and 39 walleye were tagged on the station site. To date we have received
three tag returns. Two northern pike tagged in early Spring 1978 were
-25-
-------
caught in May 1978 approximately 17 km downstream by fishermen in Lake
Wisconsin near Okee, Wis. A third northern pike also tagged in spring 1978
was caught in October 1978, 5 km downstream in the Wisconsin River.
Effects of Ashpit Effluent on Reproductive Success
Egg mortalities for the field incubation tests were high at all net
locations (Figure 7). Eggs incubated in the ashpit drain, however, had a
higher survival rate (4.6%) after 10 days than eggs incubated in upstream
Rocky Run (3.5%) or downstream Rocky Run (0.2%). Water quality data in the
ashpit drain showed higher conductivity and warmer temperatures, while total
alkalinity and hardness were reduced (Table 4). Although more material
precipitated in sediment traps in the ashpit drain qualitatively it was very
different from that collected in sediment traps set in Rocky Run. The
ashpit drain material was almost entirely the white flocculent material from
the ashpit, whereas the Rocky Run material consisted of heavier, brown,
organic sediments that settled on the eggs.
Newly hatched larvae were held at the same sites as the egg incubation
tests (Figure 8). Again, survival was better in the ashpit drain (54.8%)
compared to downstream Rocky Run (39.7%) or upstream Rocky Run (6.6%).
Although yolk-sac fry survival was better than egg survival, we still felt
siltation confounded the results and hence turned to controlled laboratory
bioassays to assess the effect of the ashpit effluent on fish reproduction.
Trace-Element Analysis of Fry—
Northern pike fry hatched at various locations in the marsh during the
above field incubation tests were analyzed for 20 trace elements (Table
5). The 10 days that these fish spent in the marsh covered the
embryonically active eyed-egg period through the early larval period. For
all incubation sites a general increase in trace-element levels was found in
fry as compared to levels in eggs. This observation reflects both embryonic
development and the greater ionic constitution of the water compared to
waters where the parents were caught (Lake Butte des Morts, Winnebago Co.,
Wis.). Fry hatched in the ashpit drain contained elevated levels of only
one element, sodium, compared to fry hatched at marsh locations not
influenced by the ashpit effluent. High iron levels were found in fry
hatched in upper Rocky Run but the reason for the elevated iron levels is
not known.
Egg Survival in Laboratory Tests—
In the laboratory, pike egg survival was significantly lower in water
from the ashpit drain than in water from either a control area in upper
Rocky Run or in a natural mixture of ashpit and Rocky Run water (Figure
9). Pike eggs were most susceptible to the toxicity of ashpit drain water
during the developmental period following gastrulation (day 3 in Figure
9). Daily mortality rates were significantly greater for eggs hatched in
ashpit water only on days 2 through 6 (p < 0.05, analysis of variance,
Snedecor and Cochran 1967).
-26-
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100
u
8 10
1.0
z
LLJ
U
EGG SURVIVAL 1977
.ASH PIT
.ROCKY RUN
..MIXTURE
95% HATCH
1 234567
DAYS
APRIL 16,1977
8 9 10 11
Figure 7. Survival of northern pike eggs hatched -in situ in the wetlands at
the Columbia site during April 1977. The ashpit site corresponds
to net 4 and the Rocky Run site to net 1 in Figure 3. The
mixture site corresponds to a natural mixture of Rocky Run and
ashpit water at net 2 of Figure 3. Ten bottles with approximately
50 eggs in each were used at each site.
-27-
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TABLE 4. WATER QUALITY DATA FOR VARIOUS STATIONS
IN THE MARSH—MARCH 21-MAY 12, 1977
Temperature
(°C)
Current
Speed (cm/s)
Dissolved oxygen
(mg/liter)
Conductivity
(ymhos/cm)
@ 25°C
Total alkalinity
(ppm)
Hardness
(ppm)
PH
Sediment
(ml/day)
Turbidity
(JTU)
Upper
Rocky Run
12. la
±5.8
(4.5-20.5)
9
25.6
±5.5
(20.6-36.8)
7
0.6
±6.0
(9.6-12.2)
8
485
135
(381-831)
9
220
35
(155.1-253.0)
8
301
±57
(234-368)
8
7.9
±0.1
(7.7-8.0)
/ 9
1.0
±0.6
(0.5-1.8)
4
36.4
±12.8
(13.7-54)
8
Lower
Rocky Run
13.5
±6.8
(3.9-23.0)
9
19.9
±7.3
(10.1-30.3)
6
9.8
±1.1
(8.2-11.6)
8
870
460
(448-2,028)
9
203
33
149.6-239.8)
8
247
±61
(150-240)
8
8.0
±0.5
(7.7-9.4)
9
0.8
±1.1
(0.1-2.0)
3
31.4
±15.9
(15-57.7)
7
Entrance to
sedge meadow
7.95
±3.8
(3.8-13.0)
4
11.2
±8.0
(2.6-18.3)
3
7.2
±1.6
(5.8-8.9)
3
670
720
(282-1,747)
4
132
9
(126.5-143)
3
167
±16
(150-182)
3
7.8
±0.9
(7.2-9.1)
4
0.8
1
13.9
±12.1
(0.7-30)
4
Ashpit
drain
14.2
±5.6
(6.2-21.8)
9
19.6
±10.5
(7.4-44.9)
9
9.7
±0.7
(8.3-10.9)
9
1,425
490
(443-1,940)
9
173
33
(133.1-226.6)
8
187
±78
(130-352)
8
7.9
±0.3
(7.4-8.4)
9
2.9
±2.5
(0.5-6.3)
4
19.3
±6.9
(11-34.7)
8
The first entry in each cell is the mean; the second is the standard
deviation; the third is the range; and the fourth is the number of
measurements.
-28-
-------
100
~ 50
u
CO
O
O
10
Of.
Z>
CO
o
of.
1.0
FRY SURVIVAL
1977
,ASH PIT
ROCKY RUN
•MIXTURE
9596 HATCH
1 2
APRIL 16,1977
6
DAYS
8
9
10 11
Figure 8. Survival of northern pike larvae placed in wetland at the
Columbia site for 11 days in April 1977. The ashpit site
corresponds to net 4 and the Rocky Run site to net 1 in Figure
3. The mixture site corresponds to a natural mixture of Rocky
Run and ashpit water at net 2 in Figure 3. Ten bottles with
50 larvae each were used at each site.
-29-
-------
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00
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fD
PERCENTAGE SURVIVAL
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p4 rt H fD 3
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X O (-"• i-l
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TABLE 5. CONCENTRATIONS OF TRACE ELEMENTS (ug/g=ppm ON
FREEZE DRIED WEIGHT BASIS) IN NORTHERN PIKE EGGS
AND FRY USED IN THE IN SITU BIDASSAY, SPRING 1977a
Pike fry hatched at various locations
in the marsh
Element
Br
Sm
La
As
Sb
Na
K
Cr
Se
HE
Ba
Sc
Rb
Eu
Fe
Zn
Co
Ca
Cs
Eggs (before
placement in the marsh)
13+1
0.04+0.01
0.32+0.04
0.49+0.08
0.09+0.01
2,750+27
7,133+277
1.3+0,1
1.6+0.2
0.05+0.01
6.0+1.8
0.07+0.001
6.8+0.7
0.01+0.001
366+16
65+2.0
0.15+0.01
2,533+500
0.04+0.01
Ashpit drain
(n-11)
25+2
0.13+0.01
0.58+0.06
b.d.
0.36+0.06
4,230+50
5,100+1,000
20.0+1.4
9.3+1.1
0.34+0.10
69+20
0.19+0.01
b.d.
0.15+0.03
1,910+100
222+11
0.34+0.07
58,000+8,000
b.d.
Lower
Rocky Run
(n=8)
44+2
0.13+0.02
b.d.
b.d.
b.d.
3,820+70
6,300+1,200
12.8+1.9
5.8+1.4
b.d.
105+34
0.12+0.01
b.d.
b.d.
1,530+160
239+12
0.51+0.10
b
b.d.
Upper
Rocky Run
(n=4)
36+9
b.d.
b.d.
b.d.
b.d.
1,760+120
b.d.
27.0+5.0
7.0+3.0
b.d.
170+50
0.91+0.04
b.d.
b.d.
5,100+400
142+15
1.5+0.3
66,000+3,600
b.d.
aMean = 1 S.D.; b.d. = below detection limits: b = analysis not done.
None of the filtered-water treatments differed in mortality (Figure
10); the percentage survival at the end of the experiment was highest in the
Rocky Run filtered water, but this was not statistically significant.
Whatever factor caused the increased mortality in the ashpit drain was
removed by filtering. Water quality characteristics differed little in
filtered and unfiltered samples except that turbidity was reduced in
filtered water (Table 6). Ashpit drain water contained a suspended floe
-31-
-------
100
90
80
D
(A
Ul
S
UJ
o
oc
Ul
Q.
70
EGG SURVIVAL 1978
FILTERED TREATMENTS
ASH PIT
ROCKY RUN
,MIXTURE
GASTRULATION
95% HATCH
t
1 2 3 4 5 6 7 8 910111213
DAYS
APRIL 6,1978
Figure 10. Survival of northern pike eggs hatched in the laboratory in
April 1978 using filtered water from the ashpit drain, Rocky
Run Creek, and a downstream natural mixture. Ten lots of
approximately 100 eggs each were used for each treatment.
-32-
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TABLE 6. WATER CHEMISTRY FOR 1978 LABORATORY EXPERIMENT
Dissolved Total
oxygen Conductivity alkalinity Hardness Turbidity
(mg/liter) (ymhos/cm (mg/liter) (mg/liter) pH (JTU)
at 25°C)
Unfiltered Treatments
Upstream 8.9a 390 185 208 8.0 1.0
Rocky Run (8.3-9.7) (360-440) (118-230) (142-247) (7.8-8.2) (0.9-1.8)
7 7 7776
Ashpit 9.0 1,155 104 158 7.6 .9
(7.6-9.8) (310-1,700) ( 99-122) (141-187) (7.6-7.7) (0.6-3.5)
7 7 7776
Downstream 9.1 525 157 189 8.0 1.2
Rocky Run (8.5-9.9) (290-630) ( 89-189) (184-218) (7.6-8.2) (0.6-2.4)
7 7 7776
Filtered Treatments (3 pm Millepore filter)
Upstream 9.1 435 184 210 8.0 0.9
Rocky Run (8.6-9.6) (375-560) (115-229) (144-250) (7.8-8.2) (0.7-1.5)
7 7 7776
Ashpit 9.2 1,200 101 162 7.7 1.3
(8.1-9.7) (300-1,540) ( 73-118) (94-189) (7.6-7.8) (0.6-2.3)
7 7 7776
Downstream 9.0 490 160 189 8.0 1.1
Rocky Run (7.5-9.7) (275-595) ( 79-186) ( 99-220) (7.7-8.2) (0.7-2.0)
7 7 7776
aThe first entry in each cell is the mean; the second is the range; and the
third is the number of samples.
that was filtered out and we suspect that this material caused the
difference in toxicity between filtered and unfiltered ashpit drain water.
Concentrations of trace contaminants in solution were not thought to be
altered significantly by filtering because pH, temperature, and conductivity
were similar in both filtered and unfiltered water. Actual trace-element
concentrations were not determined, however. Routine water quality
measurements showed ashpit drain water to have a higher conductivity and
lower pH than water from the other sources (Table 6).
-33-
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Year-Class Strengths
Our objective was to determine if harmful effects from construction and
operation of a coal-fired power plant on a pike spawning ground could be
detected as weak or missing year-classes. Having only one preoperational
sampling and no preconstruction sampling proved a limitation in attempting
to relate habitat changes caused by the plant to spawning success of
northern pike. Our data show typical age-dependent mortality for northern
pike (Figure 11) with most spawners between 2 and 4 years old.
The 1971 year-class was the first affected by major construction on the
generating station site (Table 7). Fish from that year-class were already 5
years old at the time of our first sampling in 1976. If destroying a
significant portion of the spawning marsh resulted in a loss of some fixed
percentage of pike fry, fewer fish in all subsequent years would reflect
this loss. Hence, the northern pike spawning population may be reduced
because of habitat loss, but without historical data on the population
before 1971, this loss would be undetectable. Since the 1973 year-class was
well represented in our samples (Figure 11), construction of the ashpit
dikes in 1973 did not result in a detectable reduction of year class
strength.
Because the plant began operation in mid-spring 1975, any effects from
the ashpit effluent would not be evident until the 1976 spring spawning
season. Our 1978 catch data indicate a reduced 1976 year-class. Two-year-
old fish constituted a large portion of the 1976 (17%) and 1977 (37%)
catches, but represented only 7% of the 1978 catch. Water levels for the
1976 spawning season were similar to those of 1975 and 1974; therefore,
adequate spawning habitat was available for all three year classes.
Only in 1977 were enough walleyes captured to allow assessment of
population age structure (Figure 12). The 1976 year-class was only 1 yr old
at that time and hence not very susceptible to our sampling techniques. As
this would be the first year-class to be affected by the ashpit effluent, we
can not determine what effects the effluent might have on walleye
reproduction. Furthermore, this species naturally undergoes wide
fluctuations in year-class strengths (Kelso and Ward 1977).
Importance of this Site to the Wisconsin River Fishery
Wisconsin River wetlands were grouped into 13 major areas of potential
northern pike spawning habitat (Figures 13 and 14, Table 8, Appendix B).
Because no field data exist to show if northern pike actually utilized these
areas for spawning, the comparison of spawning areas was accomplished solely
by whether suitable vegetation types existed. Vegetation types were
determined by infrared aerial photography. This comparison shows the
wetland area on the plant site (bordered by Duck Creek on the north, Rocky
Run on the south, and the Chicago, Milwaukee, St. Raul, and Pacific Railroad
tracks on the east) to be only a small percentage of the total remaining
wetland between the Wisconsin Dells and Prairie du Sac dans (Table 8). Only
13.1% of the deep-water sedge meadow and 0.5% of the shallow-water sedge
meadow areas are on the station site. Other areas with substantial deep and
-34-
-------
X
o
6
o
60
40
20
1978
n=i23
69
70
71
72 73 74
75
60
40
20
O
z
o
DO
E
z
o
o
UJ
H 60
UJ
o
oc
S 40
20
76 77
1977
n=59
70
72 73 74 75
76 77
1976
n=35
75
69 70 71 72 73 74 75 76 77 78
YEAR CLASS
Figure 11. Population-age structure of northern pike caught on the Columbia
site in 1976, 1977, and 1978. For each year fish caught in all
nets were pooled to form one population. Sample size is given
by n; the cooling lake dikes were built in 1971; the ashpit
drain was built in 1973, and plant operation began in 1975.
-35-
-------
u
s
u
40
o
z
° 30
GO
O
u
LJJ
O
u
LJU
Q-
20
10
WALLEYE 1977
n=34
1968 1969 1970 1971 1972 1973 1974 1975 1976 1977
YEAR OF HATCH (YEAR CLASS)
Figure 12. Population-age structure of walleye caught on the Columbia site
in spring 1977. Total sample size was 34 fish.
-36-
-------
PRAIRIE da
SAC DAM
Kilometers
Figure 13. Major areas of potential northern pike spawning habitat in the
Wisconsin River and tributaries near the Columbia Generating
Station. Figure 14 shows details.
-37-
-------
I
u>
00
KEY:
WATER
DEEP WATER SEDGE
I"7"] SHALLOW WATER SEDGE
|[||jjjj|| CATTAIL
*£:Z SWAMP-SHRUB
prp OTHER
Figure 14a. Detailed map of area A (Figure 13).
-------
o
KILOMETERS
Figure 14b. Detailed map of area B (Figure 13). (See Figure 14a. for key)
-------
-P-
o
I
Figure 14c. Detailed map of area C (Figure 13). (See Figure 14a. for key.)
-------
KILOMETERS
Figure 14d. Detailed map of area D (Figure 13). (See Figure 14a. for key.)
-------
N3
KILOMETERS
\
Figure 14e. Detailed map of area E (Figure 13). (See Figure 14a. for key.)
-------
r . —4
KILOMETERS
Figure 14f. Detailed map of area F (Figure 13). (See Figure 14a. for key.)
-------
I
.p-
V,
MILES
0
KILOMETERS
Figure 14g. Detailed map of area G (Figure 13). (See Figure 14a. for key.)
-------
TABLE 7. RELATIONSHIP OF CONSTRUCTION ACTIVITIES TO PIKE YEAR-CLASSES
Date
Event
Effect on year-class
January 1971
1973
May 1975
to present
First bulldozer, cooling
lake, dikes completed by April
Ashpit dikes built
Generating station in
operation
Unknown
None detected
Possibly reduced year-class
for 1976; too early to
detect changes in
subsequent year-classes
shallow sedge-meadow areas are upper Duck Creek, upper Rocky Run Creek,
Powers Creek and its tributaries, and Lodi Marsh.
Although we know from our fyke netting that the Rocky Run A and B areas
(described in Appendix B) are not important spawning grounds, they were
included in the analysis to avoid biases. On the basis of aerial
photography (the criteria used for judging sites where no fish survey data
existed) they appeared suitable; therefore, it would have been unfair to
exclude them. Information provided by the Wisconsin Department of Natural
Resources (J. Chizek, personal communication) indicates that the Lodi marsh
is also not used as a pike spawning ground. If these areas are eliminated
from the comparison, the station site contains 22.1% of the deep-water sedge
meadow and 0.8% of the shallow-water sedge meadow. The relative importance
of the generating station wetlands would probably increase even further if
fish-survey data for other marshes were available and additional areas were
eliminated as spawning grounds.
The largest habitat loss due to construction of the generating station
involves replacement of 203 ha of shallow-^water sedge meadow with the
cooling lake. This represents a loss of 18% of this wetland type formerly
occurring in this section of the Wisconsin River. Before destruction of
this wetland, the station site contained 28% of the shallow-water sedge
meadow likely to be used by spawning northern pike as opposed to the current
figure of 0.8%.
DISCUSSION
Our work has documented the use of wetlands on the site of the Columbia
Generating Station as spring spawning grounds for several important game
fish. The extent of use varies with annual flood levels, but under normal
and high water levels fish enter both the sedge meadow adjacent to the
-45-
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Floating
macrophytes
deep water
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cooling lake and the ashpit drain. Both of these areas have changed because
of the construction and operation of the Columbia Generating Station.
Wetland Water-Level and Vegetation Change
Much of the original sedge meadow at the site was replaced with a
cooling lake, and the remaining meadow is changing from shallow to deep-
water marsh (Bedford 1977). This transition is the result of a fourfold
increase in ground-water discharge rates since the cooling lake was filled
in January 1975. Prior to that time, water levels rose during spring
floods, decreased frequently in summer to below the soil surface, and rose
again in autumn. Water now stands consistently above the soil surface at
depths of from 1 to 60 cm. The resulting vegetation change from a community
of perennial sedges to one dominated by annual forbs and emergent aquatic
species has serious implications for fish reproduction. The fall dieback of
sedges leaves much densely matted intact vegetation that provides excellent
spawning substrate during spring floods. By contrast, the autumn die-off of
annuals and emergent aquatics provides no such vegetation because stalks
decompose and are dissipated before spring floods. Because these
vegetational changes began in 1975 and accelerated through 1976-77, negative
effects on fish reproduction would not have been evident until the 1976 and
later year-classes. Since these year-classes are just reaching spawning
age, monitoring of adult spawners should be continued.
Effect of_ Ashpit Effluent on_ Reproductive Success
Northern pike utilize, and may even be attracted to, the ashpit drain
area because of higher water temperatures. This is of interest because the
drain contains elevated levels of various trace elements and possibly trace
organics. McKim (1977) reviewed 56 life-cycle toxLctty tests involving a
variety of organic and inorganic chemicals and concluded that the embryo-
larval and early juvenile life stages of fish are the most sensitive to
toxicants. These are the stages during which the northern pike studied at
Columbia have the greatest exposure to the ashpit effluent. The embryo
period generally lasts about 2 weeks (Franklin and Smith 1963) and, since
juvenile pike do not begin emigrating from nursery areas until 10-24 days
after hatching, these young fish are exposed to any chemicals in the ashpit
for a minimum of 4-6 weeks. We have observed that some pike may even spend
their entire first year in the Rocky Run area. In addition to exposure
through the water, young pike may also accumulate toxicants via the food
chain. Preferred food items follow a sequence of microcrustacea, insects,
and vertebrates (chiefly other pike and tadpoles) with increases in fish
size (Hunt and Carbine 1951). These food items accumulate certain trace
elements, notably barium and chromuim (Schoenfield 1978) from the ashpit.
They may also contain harmful trace organic compounds.
Considering the elevated trace-element levels in ashpit drain water and
invertebrates (Magnuson et al. 1980), it is surprising that pike fry hatched
there did not show higher element concentrations than pike from unaffected
areas. The only elevated element, sodium, is of little toxicological
importance. The reasons for the higher iron concentrations in upstream Rocky
Run fry are unknown. The small sample blomass and inherent analytical
-47-
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errors cause differences between incubation sites for the other elements to
be insignificant.
The lack of apparent biological accumulation of toxic trace elements in
pike fry hatched in the ashpit drain does not imply that the ashpit effluent
had no negative long-term effects. Because few pike eggs survived to
hatching, our sample sizes were small. Also, the 10-day exposure period in
the marsh constitutes only a small portion of the 4-6 weeks that the fry
minimally spend in nursery areas before emigration. Furthermore, our fry
were still in the yolk sac stage when frozen for analysis; hence, they had
not yet begun feeding and no uptake via the food had occurred.
Crayfish caged in this same ashpit-drain system accumulated chromium,
zinc, selenium, and iron over a 2-month period (Harrell 1978, llagnuson et
al. 1980). Trace-element accumulation was also shown for a variety of
invertebrates living in the drain (Helmke et al. Unpublished). These
studies indicate that long-term exposure probably does result in biological
accumulation of trace substances from the ashpit effluent. Yet the survival
of organisms and the 1976 catch of pike fry in the ashpit drain indicate
that ashpit water is not acutely toxic. Any long-term negative effects from
the trace-element contaminants or flocculent precipitate entering the Rocky
Run slough will only be evident by monitoring the northern pike population
structure during the next several years. Weak or missing year-classes that
cannot be attributed to climatic factors will be evidence for such negative
effects.
Results for 1976—
Despite the changes in these wetlands, pike continued to use them for
reproduction. Successful reproduction in the affected areas was documented
in 1976 by the capture of pike fry in both the ashpit drain and the outflow
channel from the sedge meadow adjacent to the cooling lake. However, the
chemical and physical changes in wetlands were, in many cases, just
beginning by 1976, 1 year after the plant began operation. Therefore,
successful reproduction in 1976 is no guarantee that reproduction will
remain unaffected in future years.
Results for 1977 and 1978—
Despite intensive sampling efforts in 1977, no fry were caught anywhere
in the Rocky Run slough area. In 1977, spring water levels were very low
and resulted in limited amounts of good spawning habitat. Although some
1977 fry were caught as yearlings in 1978 (Figure 11), evidence of a weak
1977 year-class may appear in future fyke-netting efforts. Sampling with
light traps, highly efficient at collecting fish larvae in other areas,
failed to catch any pike fry in 1978, despite high spring water levels.
Whether these results represent poor hatching success or simply inadequate
sampling gear is unknown. Since documentation of the natural hatch proved
unfeasible in the extensive Rocky Run slough area, we utilized two other
methods for assessing reproductive success: Egg and fry survival bioassays
and an analysis of year-class strengths.
-48-
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Egg and Survival Bioassays
Our attempts to study hatching success by in situ bioassays were
complicated by factors not directly related to the ashpit effluent. Our
hatching rate was less than 5% at all sites; similar work has shown success
rates to average 60-70% for Lake Malar, Sweden, and 74% for Lake George,
Minn. (Franklin and Smith 1963). We feel that our poor hatch was the
result of excessive siltation in the incubation bottles and probably was
anomalous. Our results for larval survival were better, but again excessive
siltation was a problem. Since accumulation of natural sediments was least
in the ashpit drain, the greater survival of eggs and fry there may be
attributed to this factor.
Results from the laboratory bioassay indicated that unfiltered ashpit
water was toxic to pike eggs, but only during the developmental period
following gastrulation. Survivorship was not monitored during the early
juvenile period, but based on literature reports, we expect this to also be
a sensitive life stage. When ashpit water was filtered, the milky white
suspended floe was removed. Since filtering eliminated the toxicity of
ashpit water to pike eggs, the implication is that this floe is harmful to
egg survival. We aerated the eggs to minimize settling out of any suspended
materials. Since such settling out might be greater under some field
conditions and since we have observed extensive areas on and near the site
where the marsh bottom is covered by the floe, detrimental effects on egg
survival are likely.
Analysis of Year-Class Strength
Analysis of year-class strengths revealed no negative effects due to
the construction and early operation of the generating station. ttowever, it
is too early to assess the eventual effects of the station on pike year-
class strengths. Although the station began operating in the spring of
1975, physical and chemical changes in the area's wetlands were just
beginning to become obvious in 1976 and 1977. With the operation of an
additional power generating unit in 1978, further effects may be expected.
Our fyke-netting efforts have provided some background data from which
further changes in pike populations might be detected. Since it will be
several years at least until the chemical and physical alterations of the
marsh result in the establishment of a new equilibrium, monitoring of adult
year-class strengths will be an important method of assessing reproductive
success of the game fishes.
Success of Tagging
The failure to recapture any fish tagged in previous spawning runs
could indicate that fyke nets are inefficient during extreme flood periods
or that northern pike do not necessarily return to the same spawning grounds
each year. During peak flood periods, water levels were occasionally 1 m
over the nets. Hence, fish could move upstream over the net or could use
alternate, temporary channels. This hypothesis is supported by the capture
of several pike in waters upstream of net 2 (entrance to Rocky Run slough)
that had not been captured in that net.
-49-
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To improve the capture of adult spawners, we suggest that after the
upstream spawning run is completed, nets be turned around to catch fish
leaving the marsh after spawning. Since water levels are usually lower
then, net efficiency should be enhanced. This procedure would also allow an
estimate of the adult spawning population by the Petersen tag and recapture
method (Priegel and Krohn 1975).
The other explanation for the lack of recapture involves the issue of
whether or not northern pike home to certain grounds. Such behavior is well
documented in salmon and trout (Hasler et al. 1978) and has been suggested
for walleye. Studies on homing in northern pike are scarce, but they do not
suggest such a tendency (Franklin and Smith 1963). If homing does occur,
however, then loss of spawning habitat on the station site could endanger a
genetically distinct pike population and contribute to the genetic
impoverishment of the species in this portion of its geographic range.
Sampling of alternate spawning marshes for fish previously tagged on the
Rocky Run slough area should indicate if pike use alternate spawning areas
or return to the same areas yearly.
The three tags returned by fishermen supported our belief that pike
spend the majority of the year in the Wisconsin River and particularly in
Lake Wisconsin. Therefore, effects from the power plant are likely to be
important for only the embryo and early juvenile life stages, which are also
the most sensitive to chemical toxicants. This finding demonstrates one of
the many roles flood plains play in the functioning of river ecosystems and
how far-reaching the effects of flood-plain disturbance can be.
Inventory of Spawning Areas
The inventory of potential spawning areas by infrared aerial
photography shows that many alternate sites exist in this portion of the
Wisconsin River. Strict inventory by photo-interpretation can be
misleading, however, since our data show that fish do not utilize the upper
stretches of Rocky Run Creek, even though considerable sedge meadow is
available there. Inclusion of such areas in the overall estimate of
spawning habitat would underestimate the importance of smaller, but more
productive, wetlands. Nevertheless, areas currently affected by the
Columbia Generating Station do not constitute the major portion of the
wetland areas remaining in this section of the Wisconsin River. Wetlands on
the plant site are known to attract large numbers of spawning fish, however,
and therefore their importance to the Wisconsin River fishery should not be
overlooked. Future degradation of the area wetlands should be minimized or
else the loss of further documented spawning marsh is likely.
Final Considerations
We feel that our study has successfully documented the use of areas
affected by the power plant as pike spawning grounds. In situ egg hatching
bioassays were inconclusive, but our laboratory bioassays indicate that
ashpit-drain water is toxic to certain developmental stages of pike
embryos. One of the major considerations in determining the importance of
this area to the Wisconsin River fishery is whether unique homing fish
-50-
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stocks are involved. In addition to northern pike, muskellunge, and
walleye, other spring spawning fish, such as catostomids, are affected. If
homing is important, then loss of habitat involves loss of these stocks. If
fish simply search for the best suitable habitat, then loss of wetlands on
Rocky Run Creek might be compensated for by a switch-over to alternate
spawning sites, but this could also represent a reduced area for young-of-
the-year production.
Finally, the year-classes of 1976 and later should be studied carefully
because they will probably reflect the long-term effect of the plant. A
series of weak year-classes that cannot be correlated with low water levels
or extreme water-temperature changes, and which are not evident in spawning
populations at other sites, would be a strong indication of reproductive
failure due to operation of the Columbia Generating Station.
-51-
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SECTION 4
ZINC TOLERANCE IN FOUR GENERATIONS OF FLAGFISH
INTRODUCTION
In Section 2 we observed that, despite the changes in wetland water
levels and substrate, northern pike and other game fish continued to use the
affected marsh area at the Columbia site for spawning. In this section we
will examine a question that may arise at any wetland site affected by the
construction and operation of a generating station: Can fish populations
adapt genetically to increased trace-element levels in the environment by
evolving tolerance? Tolerance is defined as the relative capacity of an
organism to grow or thrive when subjected to a normally unfavorable
environmental factor. Chronic exposure to toxicants may favor those
individuals with a genetic make-up that confers resistance. Through the
process of natural selection, the tolerant members of the population survive
and transmit the trait of tolerance to their offspring.
LITERATURE REVIEW OF METAL TOLERANCE
Contaminants can cause selective pressures that result in tolerant
populations as shown by the many documented cases of pesticide-resistant
insects (Crow 1957). The number of insect species and the types of
chemicals involved are numerous, but two general observations are
noteworthy: (1) the evolution of tolerance is rapid and (2) the mechanisms
of resistance are diverse. Among the mechanisms identified are the
development of behavior patterns that lessen exposure to the poison,
decreased uptake through reduced permeability of the cuticle, and enzymatic
detoxification.
The widespread and often indiscriminate use of chemical poisons has
subjected many other types of organisms to similar selective pressures.
Studies have found pesticide-resistant populations of fish (Vinson et al.
1963), crayfish (Albaugh 1972), frogs (Ferguson and Gilbert 1967), and mice
(Webb and Horsthall 1967). In many cases resistance may be an acquired
trait through inducible detoxifying enzymes (Webb and Horsthall 1967, Brown
1976), but it can also be inherent in certain organisms regardless of prior
pesticide exposure (Crow 1957, Ferguson 1967).
In addition to synthetic organic chemicals, tolerance to trace-element
contaminants also has been documented. Studies of both terrestrial plants
(Antonovics et al. 1971) and aquatic plants (Stokes et al. 1973, McLean and
Jones 1975) show again that the evolution of tolerance is rapid, involves
-52-
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diverse mechanisms, and can occur in response to a variety of trace
elements.
Tolerance to metals has been demonstrated in recent studies of
invertebrates from contaminated aquatic habitats. Bryan (1976) found the
marine polychaete Nereis diversicolor to have zinc, copper, silver, and
possibly lead-tolerant populations. Tolerance was limited generally to the
metals in a particular habitat. Laboratory experiments indicated the
mechanism was based on reduced permeability to metal ions. Whether
tolerance was an acquired (inducible) or an inherited trait was not
determined. B. Brown (1976) found copper and lead-resistant populations of
Asellus meridianus Rac. in rivers with a history of metal pollution from
abandoned mines. Both acute lethal bioassays and chronic growth studies
demonstrated tolerance. The persistence of tolerance in second generation
organisms from a laboratory culture indicated a genetic basis for this
trait.
Despite such demonstrations of the evolution of metal tolerance in
aquatic plants and invertebrates, no examples of naturally metal-tolerant
fish populations have been reported in the literature. Although waters with
high metal contamination frequently contain plants and invertebrates, fish
are conspicuously absent (Carpenter 1924, Jones 1958, Weatherly et al.
1967). In waters with slightly elevated metal levels, fish populations
exist but reproduction is severely depressed and their long-term survival is
uncertain (Van Loon and Beamish 1977, McFarlane and Franzin 1978).
OBJECTIVE OF THIS STUDY
This study tested the hypothesis that if genetic factors partially
determine the variation in susceptibility to metal toxicants, the resistance
of a fish population can be increased through selection. The ability of
fish to develop tolerance is of interest since trace elements are often
released by coal-fired power plants at levels that may have severe long-term
effects on fish populations even though they are not acutely lethal (Cherry
and Guthrie 1977). The approach used was to breed the flagfish (Jordanella
floridae), a species well suited for extended laboratory studies (Smith
1973), for resistance to the long-term effects of the element zinc.
METHODS
Zinc as an Experimental Toxicant
Zinc was chosen as the toxicant for this experiment because it is a
common pollutant whose biological effects are well documented (Skidmore
1964, European Inland Fisheries Advisory Commission 1974). Use of zinc and
flagfish provided a convenient laboratory model for the selection processes
that would affect other fish species exposed to other trace-element
contaminants. Fish exposed to zinc may die from either acute or long-term
effects. Acute mortality, the result of extensive gill damage, occurs
within the first 4 days of exposure to high zinc concentrations. Death from
long-term effects occurs at lower concentrations and involves damage to
internal organs. To determine the level of zinc resulting in chronic
-53-
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mortality, groups of flagfish were exposed to various concentrations of the
element (Figure 15). A concentration of 0.8 mg/liter was used for the
parental generation, but this was increased to 1.0 mg/liter for successive
generations to accelerate mortality and increase the selective pressures
affecting the population.
To begin the experiment, a stock supply of flagfish was randomly
divided into two groups. One group (the selected population) underwent
three generations of selection for zinc resistance while the other group
(the unselected population) was maintained under the same laboratory
conditions but received no zinc exposure (Figure 16). Selection was
accomplished by exposing adult fish to a chronically lethal zinc
concentration until 50 to 60% of the population had died. To determine if
selection was increasing zinc resistance, a subsample of each generation of
the unselected population was exposed to zinc along with the selected
population and survival times were compared between the two groups.
Biological Procedures
The original stock of flagfish was purchased from a commercial supplier
(Ross Socoloff Farms, Bradenton, Fla.) and maintained in tap water in
Madison, Wis.
A constant photoperiod of 16:8 h (light:dark) and constant temperature
of 25°C were used throughout the experiment. Fish ate frozen brine shrimp
supplemented with commercial fish food (Tetra Conditioning Food, Tetra Werke
Co.). Because of the hardness of Madison tap water (Table 9), zinc
exposures were at a dilution ratio of 1:3 (tap to distilled water). Fish
were acclimated to the dilution water for 1 week before exposure to zinc.
During the bioassay deaths were recorded at 12-h intervals; death was
defined as the failure to respond to a mechanical stimulus.
Breeding proceeded according to methods outlined by Spehar (1976).
Breeding tanks were 55-liter aquaria divided into thirds by plexiglass
partitions. Eggs, deposited on orlon yarn spawning substrates, were
collected daily. Eggs hatched in egg cups made of PVC piping cut to 6-cm
lengths and covered with plastic screen at one end. Larvae ate newly
hatched brine shrimp nauplii until old enough to take frozen adult brine
shrimp. Generation time was approximately 6 months.
Bioassay Procedure
The flow-through bioassays were conducted along guidelines issued by
the U.S. Environmental Protection Agency (1975). The characteristics of the
tap and dilution water are given in Table 9. Zinc concentrations,
conductivity, and temperature were monitored daily; hardness, total
alkalinity, pH, and dissolved oxygen were measured weekly. All measurements
were taken according to methods outlined by American Public Health
Association et al. (1975). Zinc samples, some unfiltered and some filtered
through 0.4 um nucleopore filters, were measured by atomic absorption
spectrophotometry. Filtering allows an estimate of how much zinc is
actually in solution and hence readily available for uptake by fish (U.S.
Environmental Protection Agency 1975). Nominal (added) zinc concentrations,
-54-
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I
Ui
Ul
I
90
t 50
QC
O
HI
O
<
z
HI
O
DC
UI
Q.
10
0.5
TIME [DAYS]
10
20
Figure 15.
Cumulative mortality (probit scale) as a function of exposure time for flagfish exposed
to three zinc concentrations.
D 5 mg/liter
• 2.5 mg/liter
00.8 mg/liter
-------
STOCK POPULATION
GENERATION
UNSELECTED POPULATION
SELECTED POPULATIOK
b
<=>
ZN EXPOSURE
OFFSPRING
ZN EXPOSURE
OFFSPRING
ZN EXPOSURE
OFFSPRING
ZN EXPOSURE
PARENTAL
Figure 16. Procedure used in selecting for zinc resistance in laboratory
populations of flagfish. An initial stock population was
randomly divided into two groups; one underwent three genera-
tions of selection for zinc resistance while the other remained
as a control population.
aSurvivors bred to obtain data on spawning success but larvae
were then discarded.
^Comparisons of survival times between these groups were used
to assess effects of selection.
-56-
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length of exposure, and recovery time before breeding for each generation
are shown in Table 10.
TABLE 9. CHEMICAL CHARACTERISTICS OF MADISON, WIS., TAP WATER
IN WHICH FLAGFISH WERE RAISED, AND DILUTION WATER
IN WHICH ZINC EXPOSURES WERE CONDUCTED
Item
Tap water
Dilution water
(weekly samples,
all generations)
Hardness
(mg/liter CaC03)
Total alkalinity
(mg/liter CaC03)
Conductivity
(umhos/cm at 25°C)
PH
Dissolved
oxygen (mg/liter)
Zinc (ug/liter)
28 Oa
(235-296)
590
(510-640)
7.4
(7.2-7.8)
20
(10-50)
73
(61-88)
59
(44-67)
140
(120-180)
7.3
(6.7-7.9)
7.6
(6.7-8.1)
23
(2-45)
lFirst entry in each cell is the mean and the second is the range.
'Not measured.
TABLE 10. NOMINAL ZINC CONCENTRATION (ppm), LENGTH OF
EXPOSURE (DAYS), AND RECOVERY TIME BEFORE BREEDING (WEEKS), FOR
THE ZINC EXPOSURES OF THE PARENTAL AND THREE GENERATIONS OF FLAGFISH
Nominal zinc
Length of
concentration exposure
Generation
Parental
First
Second
Third
(ppm)
0.8
1.0
1.0
1.0
(days)
17
17
6.5
13
Selection intensity
(% dead for selected
population)
60.0
53.7
49.3
55.4
Recovery time
before breeding
(weeks )
8
3
2
1
-57-
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RESULTS—EXPOSURES AND CALCULATIONS
Parental Generation
Parental generation flagfish of the selected population were exposed to
0.8 mg/liter zinc for a 17-day period until 60% had died (Figure 17A).
Unselected population fish were not exposed to zinc. Approximately 75% of
the zinc (Table 11) was in a soluble form considered most toxic to fish
(European Inland Fisheries Advisory Commission 1974).
TABLE 11. ZINC CONCENTRATIONS FOR THE PARENTAL,
FIRST, SECOND, AND THIRD GENERATION EXPOSURES
Generation
Zn-exposed fish
(yg/liter)
Control fish
(jig/liter)
Filtered
Unfiltered
Filtered
Unfiltered
Parental
First
Second
Third
575+100a
(18)
750+65
(40)
1,122+97
18)
850+76
(18)
811+48
(7)
1,000+140
(11)
1,348
(1)
915+9
(2)
25+1
(2)
22+12
(12)
30+20
(2)
< 50
(3)
40+16
(2)
28+16
(67
40+10
(27
< 50
(1)
aFirst entry in each cell is the mean + standard deviation for water samples
taken during the course of the zinc exposure; the second is the number of
samples.
Survival times of fish exposed to zinc may be correlated positively
with fish size (Bengtsson 1974). Therefore, standard lengths of survivors
and nonsurvivors of the initial exposure were compared to determine if large
size and not an inherent resistance to zinc could explain survival (Table
12). The comparison was based on a two-sample t-test with unequal
variances. Males that died and those that survived did not differ in length
(t = 1.12, d.f. = 85, p = 0.27) (Table 12). Females, however, were longer
than nonsurviving females (t = 2.38, d.f. = 1.34, p = 0.02), but the length
difference was small. The 95% confidence limits for the length difference
shows that surviving females were only 0.3-3.4 mm (0.8-9.9%) longer than
nonsurviving females. For fish that died during exposure to zinc, there was
no correlation between body length and time to death (r = -0.06 for males
and r = -0.11 for females). Thus, for the size range of fish used,
-58-
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o
80
60
40
20
1
80
60
40
20
• A
i
*
"~ •
• m
- •
"\
'•
_
i i i i i in
1 5 10 15 20 25
* * * 2 o B
. ft 8.
•
o 0*
^ • —
o •
O A
o
•» O ™"
OQ
o
_ —
IVA.
80
LLJ
< 60
LU
O
5 40
LU
O_
20
"
• inn
IUU
80
II i
< 60
LU
3
|
Z
,m, 40
^
LLJ
CL.
20
'
o • C
O ^
O
- 2 0 —
o •
* —
0 •
o ~
o
— —
T , , , ff
1 5 10 15 20 25
• • Q D
* 8 o o
- FS °
• "~
• °
0
- *o —
• 0
9
« ^
M
, , , 2
I 5 10 15 20 25 , 5 10 15 20 25
TIME (DAYS)
•Population selected for zinc tolerance.
oPopulation not selected for zinc tolerance.
TIME (DAYS)
Figure 17. Survival rates of exposed (selected) and control (unselected) flagfish populations over
four generations. Survivorship is plotted against elapsed time in days (log scale). A,
parental generation; B, first generation of offspring; C, second generation; D, third
generation.
-------
selection was not for large size but for some other factor that allowed
certain fish to be more resistant to zinc than others.
TABLE 12. STANDARD LENGTHS (MILLIMETERS) OF SURVIVORS AND
NONSURVIVORS FOR PARENTAL-GENERATION FLAGFISH EXPOSED TO
0.8 mg/LITER ZINC FOR 17 DAYS
Sex
Nonsurvivors
Survivors
Male
Female
Mean
(S.D.)
n
Mean
(S.D.)
n
36.9
(5.3)
106
34.2
(4.2)
83
38.1
(6.7)
53
36.0
(5.4)
72
Residual effects on reproduction of a single sustained zinc exposure
were determined by comparing spawning data for both experimental and control
fish. Since fecundity is affected by fish size, wet weight and standard
lengths of fish used for spawning from both experimental and control
parental populations were compared with a two-sample t-test^with unequal
variances (Table 13).
TABLE 13. STANDARD LENGTHS AND WET WEIGHTS OF PARENTAL GENERATION
FLAGFISH USED TO PRODUCE THE FIRST GENERATION
Male
Item length(mm)
Experimental 44a
population (6)
(zinc -exposed) 30
Control ^ 47
population (6)
(not zinc-exposed) 27
Female
length(mm)
42
(4)
30
43
(5)
27
Male wet
weight (g)
3.06
(1.13)
30
3.48
(0.99)
27
Female wet
weight (g)
2.43
(0.58)
30
2.53
(0.98)
27
aThe first entry of each cell is the mean; the second is standard deviation;
and third is number of fish.
No difference in weight or length was found for either males or females
(p > 0.5). Size differences therefore were not responsible for any
fecundity differences. The spawning data, not normally distributed, were
analyzed by the Mann-Whitney U test (Siegel 1956). The two populations did
not differ in the number of days until the first eggs were laid (p = 1.00)
or in the percentage hatch of eggs (p = 0.80) (Table 14). Experimental fish
-60-
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averaged more eggs per spawn than control fish (p = 0.03) with a 95%
confidence limit of 0.4 to 13.1 more eggs per spawn. Thus, fish surviving a
concentration of zinc lethal to the majority of the population showed no
harmful residual effects on reproduction after 8 weeks recovery time.
TABLE 14. SPAWNING DATA FOR PARENTAL GENERATION FLAGFISH3
Item
Experimental
population
(zinc-exposed)
Control
population
(not zinc-exposed)
Days until
first spawn
9b
(5-20)
20
8
(15-17)
13
Number of
eggs/spawn
8.2
(0-86.7)
21
7.6
(0-53.6)
15
Percentage
hatch
74.7
(0-100)
19
76.6
(0.4-96.4)
13
Experimental fish survived a 17-day exposure to 0.8 mg/liter zinc. Control
fish were not exposed to zinc.
First entry is median; second is range; and third is number of fish pairs.
First Generation
The effect of one generation of selection for zinc resistance was
determined by comparing the survival of first-generation fish from both the
selected and unselected populations exposed to 1 mg/liter zinc. As in the
initial exposure, about 75% of the zinc was in a soluble form (Table 11).
The first deaths from zinc exposure occurred in the first-generation
(Fj) selected population (Figure 17B). Once fish began to die in the zinc-
exposed unselected population, however, mortality was higher and surpassed
the experimental population after 8 days. After 17 days, 68.9% of the
unselected population but only 53.7% of the selected population had died.
Median survival times were 12 days for the zinc-exposed unselected
population and 16.5 days for the selected population. A Mantel-Haenszel
test was used to determine the significance of this difference (Snedecor and
Cochran 1967). This test computes x values for the observed deaths during
each time period, given the number of fish alive at the start of the period
and the null hypothesis that there is no difference in susceptibility to
zinc poisoning between the groups. Each time period is treated
independently, and the dependency on previous events (as with a parameter
such as cumulative mortality) is avoided. Survival times were significantly
different (x2 = 9.9, p = 0.001), indicating that selected population fish
were less suceptible to zinc poisoning than unselected population fish.
A comparison of standard lengths of F, selected and zinc-exposed
control unselected indicated that size differences were not a factor in the
increased resistance of the experimental group (t-test, p = 0.85).
-61-
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No significant differences in length or weight for either sex were
found that might affect reproduction (Table 15) (t-test, p > 0.05). Neither
did the groups differ significantly in time to first spawn, average number
of eggs per spawn, or the percentage of hatched eggs (Table 16) (Kruskal-
Wallis Test, p > 0.05, Siegel 1956).
TABLE 15. LENGTHS AND WET WEIGHTS OF
FIRST-GENERATION FISH USED FOR SPAWNING
Item
Selected
population
(zinc-exposed)
Unselected
population
(zinc -exposed)
Unselected
population
(control)
Male
length (mm)
34a
(6)
12
37
(6)
8
33
(7)
14
Female
length (mm)
30
(3)
12
33
(2)
8
32
(4)
14
Male
wet
weight (g)
1.24
(0.62)
12
1.53
(0.57)
8
1.24
(0.81)
14
Female
wet
weight(g)
0.89
(0.28)
12
1.12
(0.18)
8
1.05
(0.37)
14
aThe first entry is the mean; second is standard deviation; and the third is
number of spawning pairs.
Second Generation
The effect of breeding two generations for resistance to zinc was
determined by comparing fish survival from both selected and unselected
populations when exposed to a lethal zinc level (Figure 17C). Zinc analysis
(Table 11) revealed a slightly higher zinc concentration (1.1 mg/liter) than
the intended 1.0-mg/liter level.
As in the first generation, second-generation fish from the selected
population proved more resistant to zinc poisoning than fish from the
unselected population. After zinc exposure of 6.5 days, mortality was 68.5%
for the zinc-exposed unselected population, but only 49.3% for the selected
population. Median survival times were 5.25 and 6.4 days for the
zinc-exposed control and experimental populations, respectively. The
difference in survivorship was significant at the p = 0.004 level (x1 =8.55,
Mantel-Haenszel test).
1
Selected-population fish were significantly longer (p = 0.05) than
unselected-population fish with average standard lengths of 23.8 and 25.5
mm, respectively. Considering results from the parental generation, the
size difference probably was not a significant source of variability in
zinc-tolerance levels between these two populations. The two groups did not
-62-
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TABLE 16. SUMMARY OF SPAWNING DATA FOR Fj-F3 GENERATIONS OF FLAGFISH
Generation
F,
1
Unselected pop-control
Unselected pop — Zn exposed
Selected pop — Zn exposed
F2
Unselected pop-control
Selected pop — Zn exposed
F.,
3
Unselected pop-control
Unselected pop — Zn exposed
Selected pop — Zn exposed
Days until
first spawn3
11
(6-18)c
4
7
(5-20)
8
6
(3-21)
7
4
(4-6)
7
7
(4-14)
6
33
(23-37)
10
25
(18-31)
8
22
(18-25)
4
No. of
eggs/spawn
0
(0-6.1)
10
2.2
(0-3.5)
10
3.7
(0-13)
9
2.4
(0-27.6)
11
0
(0-6.7)
14
0
(0-12.7)
10
2.0
(0-18.3)
11
0
(0-8.9)
12
Hatch (%)a
89
(2-100)
4
90
(50-100)
7
81
(4-100)
7
84.8
(73.7-100)
7
73.7
(0-86.4)
8.7
(83-94)
4
99
(89-100)
7
92
(71-100)
4
Only spawners producing eggs are included.
Includes all spawners with pairs producing no eggs counted as zero.
GThe first entry is the median, the second is the range, and the third is
number of fish pairs.
differ in the number of days to first spawn (p = 0.15), the average number
of eggs per spawn (p = 0.22), or the percentage of hatched eggs (p = 0.1)
(Mann-Whitney U Test) (Table 16).
-63-
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Third Generation
The effect of breeding three generations for resistance to zinc was
determined by comparing survival of fish from the experimental and zinc-
exposed control populations when exposed to a lethal zinc level (Figure
17D). Zinc water analyses are given in Table 11.
In contrast with the first two generations, mortality was higher in
third-generation selected-population fish than in third-generation
unselected-population fish. After an exposure period of 13 days, 46.8% of
the zinc-exposed unselected population and 55.4% of the selected-population
fish had died. Median survival times were 13.5 days for the zinc-exposed
unselected population and 12.5 days for the selected population. Despite
this decline in tolerance for third-generation experimental ftsh, the
difference in survival between the two groups was not statistically
significant (xf = 1.63, p = 0.20, Mantel-Haenszel test). The two
populations did not differ in standard lengths (t-test, p = 0.59); hence,
size was not a factor in the failure of the experimental population to
display superior tolerance.
The only difference in spawning success between selected nad unselected
population fish exposed to zinc and unselected population not exposed to
zinc was that the latter group took longer to produce eggs than the other
groups (Kruskal-Wallis Test, xf = H.l. P = 0.05) (Table 16).
The results of breeding three generations of flagfish for resistance to
zinc are summarized in Figure 18. The absence of a trend toward increased
tolerance with continued selection suggests that two confounding factors may
be involved, inbreeding depression and carry-over effects. Both phenomena
are discussed below.
DISCUSSION
Variability ^Ln Zinc Tolerance
These experiments demonstrate considerable variation in the tolerance
of fish to zinc. In the parental generation, for instance, some fish died
after only several days of exposure to zinc whereas others lived for the
entire 17-day period. In a similar study (Bengtsson 1974) some fish
survived a 100-day exposure to concentrations of zinc that were lethal to
the majority of the population.
As with lethal tests, sublethal studies also reveal considerable
variation in the response of organisms to toxicants. The sublethal effects
of contaminants are being examined in many bioassays ranging from the
subcellular to the community level (McKim et al. 1976, Sprague 1971, G.W.
Brown 1976, Maki and Johnson 1976). In a study of the sublethal effects of
copper on the coujh frequency, locomotor activity, and feeding behavior of
brook trout, response variation was evident (Drummond et al. 1973). The
cough frequency of some fish increased markedly, whereas that of others did
not change.
-64-
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Ui
I
-~
40
u
LJU ±-
U- >
u.
20
LU
O Z 0
£ <
z a
LLJ LU
Lti -20
Q_
1234
GENERATIONS OF SELECTION
Figure 18. Summary of three generations of selection for zinc resistance in flagfish.
-------
The two sources of this variation, environmental and genetic factors,
must be understood if we are to predict the long-term effects of
contaminants. Environmental factors alone can have a dramatic effect on an
organism's response. In a study by Spehar (1976) flagfish larvae exposed as
embryos to zinc and cadmium were more tolerant than unexposed larvae,
although both sets were produced by unexposed parents.
Differences in the response of organisms to a pollutant are termed
"phenotypic variability" (observable variability due to both an organism's
heredity and environment). Even in bioassays with strains of laboratory
fish raised under uniform conditions, much phenotypic variability
persists. Such variability might be caused by differences in the general
health or vigor of individual fish and not by genetic factors conferring
resistance to a particular toxicant. A study by Sparks et al. (1972)
demonstrates the importance of stress on fish tolerance. When pairs of
bluegills were exposed to lethal zinc levels in bare aquaria, dominant fish
had a significantly longer survival time. When the experiment was repeated
with shelters provided, survival times did not differ. Presumably,
subordinate fish were no longer subject to the additional stress of
harassment by dominant fish.
If the response variation to toxicants is caused by nongenetic
differences in health and vigor of fish, then toxicants should act simply as
a general stress on the population, culling out the weak individuals. An
increase in the gene combinations conferring resistance should not
continually increase the tolerance of the population. Simply stated,
selection should not continually increase the resistance of future
generations. However, an increase in resistance through successive
generations has been demonstrated for many contaminants and many species
(see Introduction), indicating that genetic factors must be partly
responsible for phenotypic variability.
Differences in the reaction of fish to toxicants also might be caused
by factors such as size, sex, age, diet, and degree of acclimation. In this
study, as in many others, diet and degree of acclimation were carefully
controlled. We found that size had little effect on survival time for the
parental generation.
The literature provides conflicting results about the effect of age on
resistance to a toxicant. Age was not considered in this study except that
only adult fish were exposed to zinc. Jones (1938) found no difference in
survival times between juvenile and sexually mature sticklebacks exposed to
a range of zinc concentrations. Bengtsson (1974) reported, however, that
the resistance of the minnow Phoxinus phoxinus to zinc increased with age.
Adelman et al. (1976) found that even for fish of a constant age or size
class raised under identical laboratory conditions, variability in response
to a toxicant still exists.
The presence of this variation in bioassays, despite uniform laboratory
conditions, suggests that genetic differences between individuals may be a
significant source of variability. Although genetic studies on zinc and
fish have not been done, work on lead and mercury suggests that considerable
-66-
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variability in response to metal pollutants may be genetically based.
Burger (1973) studied the inheritance of lead resistance in guppies and
obtained a heritability estimate of 0.26 to 0.57 for survival time of fish
exposed to acutely lethal levels of lead. [Heritability measures that
fraction of the total phenotypic variability due to the additive effects of
genes; this part of the phenotypic variability is the most responsive to
selection (Crow and Kimura 1970)]. In a study of mercury resistance in
steelhead trout, Blanc (1973) estimated the heritability to be 0.5 for
survival time in chronically lethal levels of methylmercury.
Despite these demonstrations of genetically based differences in the
tolerance of fish to toxic metals, we should be cautious of ascribing the
variability found in bioassays to genetic sources. In a study by Rachlin
and Perlmutter (1968) guppies still varied in their response to zinc after
31 generations of inbreeding. Since such prolonged inbreeding should have
greatly reduced genetic variability, environmental factors appear as the
chief source of the variation, although the fish were raised and exposed
under uniform laboratory conditions.
Effects of Zinc on Reproduction
This study found that after only 2 to 8 weeks recovery time, flagfish
surviving a zinc exposure level lethal to the majority of the population
were able to reproduce as successfully as the control population. Similar
work with zebrafish indicated that a 9-day period of zinc exposure reduced
egg production and egg fertility, but these processes returned to normal
levels within several weeks of a return to uncontaminated water (Speranza et
al. 1977). Results of these relatively short-term exposures to zinc contrast
with results of exposures encompassing the entire life cycle of fish. In
the long-term studies, reproduction was reduced significantly although
survival was unaffected (Brungs 1969, Bengtsson 1974). Since contaminants
in an aquatic environment may be present only intermittently or temporarily
(Cairns et al. 1971, Leland et al. 1976), it is important to know if fish
have the ability to recover from temporary exposures. Our work suggests
that such recovery is possible for fish exposed to zinc.
Whether selection for resistance to lethal effects also causes
resistance to sublethal effects could not be determined in this study
because the zinc-exposed control and experimental populations for both the
first and second generations did not differ in spawning success and no other
sublethal aspects were investigated. Such a phenomenon might be expected in
fish, however, since it has been demonstrated in metal-tolerant
invertebrates (B.E. Brown 1976).
Selection for Resistance
After three generations of selection, the selected line was more
tolerant for the first two generations, but showed no difference from the
unselected line in the third generation. The absence of any trend toward
increased resistance with continued selection (Figure 18) does not
necessarily indicate an inability to evolve metal tolerance. The actual
decline in the relative tolerance of the experimental population suggests
-67-
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that confounding factors are responsible. Even if none of the variability
in metal tolerance was genetic, the experimental population should have
remained equal to the control line.
The continued decline in performance of the experimental population
suggests the phenomenon of inbreeding depression. Although both lines had a
similar number of parents for each generation (Table 17), it has been shown
that in populations undergoing selection, selected parents are related more
closely than randomly chosen parents (Robertson 1961). Because close
relatives are more likely to produce offspring with harmful homozygous
recessive gene combinations, the result of such inbreeding is a general
depression of survival and vigor in the selected population (Kincaid
1976). In our experiments, parents of the selected population were chosen
on the basis of having survived an exposure to zinc, whereas unselected
population parents were chosen randomly. Fish surviving the zinc exposure
were more likely to be closely related, if survival is genetically based,
having inherited favorable genes from a common ancestor. Therefore, the
decline in overall health and vigor because of inbreeding would be greater
for selected-population fish than for zinc-exposed unselected-population
fish. Hence performance of the experimental population would decline with
increasing generations of selection.
TABLE 17. NUMBER OF PARENTS CONTRIBUTING LARVAE FOR EACH GENERATION
Unselected line Selected line
Generation Male Female Male Female
Parental
First
Second
Third
14
16
10
7
11
16
10
7
20
17
11
8
20
17
9
8
Another possible explanation for the decline of the experimental
population is the carry-over of harmful effects from mother to offspring
through the egg cytoplasm. Exposed fish are expected gradually to lose zinc
retained in their body tissue upon returning to clean water. It is possible
that females incorporate some of the zinc into the cytoplasm of their eggs
and that offspring of females exposed to zinc begin life with elevated zinc
levels. According to this hypothesis, exposure of these offspring to zinc
later in life results in increased susceptibility (McKim 1977). The
magnitude of this carry-over effect would increase with a decrease in the
recovery time allotted to females before breeding. Because recovery time
was shortened with each generation of selection (Table 10), any carry-over
effects work against the direction of selection. By the third generation
the parents had had only 2 weeks' recovery time before breeding, and the
carry-over effect may have finally negated the effects of selection.
-68-
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The relative importance of inbreeding depression and carry-over effects
in producing the observed decline in zinc tolerance will be tested in future
work. Because the selected population showed an initial increase in zinc
tolerance and similiar laboratory studies have produced an increase in metal
tolerance (Blanc 1973, Burger 1973), it was concluded that fish possess the
genetic potential to evolve metal-tolerant populations. Laboratory studies
are usually limited to observing a few generations and are often subject to
confounding influences (e.g., inbreeding). Hence, ultimate resolution of
whether fish can genetically adapt to increased metal levels will come from
studies of populations chronically exposed to metal contaminants.
No studies have documented whether natural fish populations in
chronically contaminated waters ever realize this genetic potential. Fish
are reported to live in a series of Canadian lakes receiving metal inputs
from nearby smelters (Van Loon and Beamish 1977), but a more recent study
indicates that long-term survival of these fish is uncertain. McFarlane and
Franzin (1978) reported that a population of white suckers, Catostomus
commersoni, suffered severe reproductive impairment in one of these lakes
with high zinc levels (141 to 341 mg/liter) although adult survival was
relatively unaffected. Similar effects were noted on other fish species
from the same lake. The smelter has operated since 1930, and metal
concentrations presumably have been increasing since then, although
historical data on metal levels in the water are unavailable. The onset of
harmful effects on local fish populations indicates that the fish failed to
adjust genetically to stressful metal levels during this relatively brief
period.
A similar situation exists in the soft-water lakes of the Adirondacks,
where aluminum leached from surrounding soils by acid precipitation has
killed many fish (Cronan and Schofield 1979). The development of harmful
metal concentrations in these lakes is a recent event, and again fish are
not adjusting rapidly to this sudden change in their environment.
Mclntosh and Bishop (1976) used bluegills from a metal-contaminated and
an uncontaminated lake to compare relative survival in an acutely lethal
exposure of cadmium. They found no difference in the 96-h LCcQ value for
the populations. In a sublethal exposure of cadmium, they reported a
significantly lower cough rate for fish from the contaminated lake than for
control fish; no difference was found in breathing rates, however. Whether
fish from the contaminated lake actually represented a population exposed to
selective pressures for metal tolerance is questionable, however, because
metals in the lake were not distributed evenly and it was not known how long
fish had been in the contaminated areas.
This study focused on the potential for the rapid evolution of metal
tolerance in fish populations. The results suggest that fish possess this
potential, but the limited studies of natural fish populations living under
chronic metal stress do not support our findings. Further work on fish
populations inhabiting waters with high trace-element levels is needed
before we can determine if and how fish can adapt ultimately to these
conditions. Man's pollution of aquatic systems is extremely rapid on an
-69-
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evolutionary time scale. If this pollution continues, fish populations must
adapt at an equally rapid pace to avoid decimation or local extinctions.
-70-
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SECTION 5
USE OF TEMPERATURE PREFERENCE AND ACTIVITY AS A SUBLETHAL
BIOASSAY FOR THE TOXIC EFFECTS OF ZINC TO BLUEGILL
INTRODUCTION
The fly ash emitted from coal-fired generating stations contains
microcontaminant.s, including metal ions, that dissolve and may enter natural
water systems. To monitor biological reactions to these toxicants, we need
methods that are quick, sensitive, and accurate. Such methods aid in
setting standards, monitoring spills, discovering synergistic effects (the
cooperative effect of several factors working independently), and monitoring
microcontaminant levels in mining runoff and industrial waste waters.
This section documents our effort to test one of these methods: A
temperature-preference apparatus. Preferred temperature is a stable
behavioral trait for fish (Magnuson and Beitinger 1978). Change in
temperature preference indicates a response to some other factor such as
stress from starvation (Javaid and Anderson 1967, Stuntz 1975, Stuntz and
Magnuson 1976) or pollution (Ogilvie and Anderson 1965, Peterson 1973). In
fact, knowledge of the concentration at which contaminants affect fish
temperature preference might serve as a useful indicator of sublethal
toxicity. Behavioral tests are a more sensitive indicator of these harmful
effects than lethality experiments and require less time and space than
tests on sublethal chronic effects (Schere 1977, Henry and Atchison 1979).
DESIGN OF THE STUDY
We tested an electronically controlled, temperature-selection apparatus
(Neill et al. 1972) for possible use as a tool for the detection of
sublethal concentrations of zinc by the bluegill (Lepomis macrochirus). This
system uses a temporal gradient and provides a record of activity and
temperature. Zinc was chosen as the contaminant because its acute and
sublethal toxicity are well documented (Cairns and Scheier 1957, Sprague
1968, Brungs 1969, Burton et al. 1972, Waller and Cairns 1972, Cairns et al.
1973). A sublethal zinc concentration (2.5 ppm) was selected on the basis
of preliminary experiments and previous studies (Burton et al. 1972, Waller
and Cairns 1972, Cairns et al. 1973).
-71-
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MATERIALS AND METHODS
Selection Apparatus
Nine aquaria, each divided into two compartments with an
interconnecting tunnel, were equipped with a system that permits fish to
control the temperature of the aquaria. With a constant 2°C temperature
difference between the two compartments, the fish controls the direction of
temperature change (3°C/h) in the tank. If the fish is in the warmer
compartment, the temperature of the whole tank is increasing; if the fish is
in the cooler compartment, tank temperature decreases. Temperature is
selected oa a temporal rather than spatial basis (Neill and Magnuson 1974,
Beitinger et al. 1975). A computer continuously monitors and records tank
temperatares and movement through the tunnel (activity). Changes due to
zinc may occur in the activity rate or selected temperature (e.g., increase,
decrease, or diurnal pattern modification).
General Conditions
The water used for the experiments was Madison, Wis., city water
diluted at a ratio of 1:7 with distilled water. Madison city water is
unusually hard (300 ppm CaCOo), and zinc is less toxic to the bluegill in
hard water (Cairns and Scheier 1957). The 1:7 dilution results in hardness
of 40 to 50 ppm (Table 18), a level within the range of much previous v;ork
with zinc (Cairns et al. 1973). Experiments were conducted during November
and December 1976. The 12:12 light-dark cycle included periods of
intermediate light levels at dawn and dusk. We analyzed data from eight
control fish and 10 zinc-exposed fish. Data were not utilized if a given
fish did not pass through the tunnel at least three tlaes 'luring a day or
night period. Temperature data from one fish were lost because of
thermistor malfunction.
TABLE 18. ROUTINELY DETERMINED CHARACTERISTICS OF WATER USE IN
THE TEMPERATURE-PREFERENCE BIOASSAY
Water
characteristic
pH3
Total hardnessb
Alkalinity0
Dissolved oxygen
Conductivity
Unit
__
ppm
ppm
% saturation
pmhos/cm at 25°C
No. of
analyses
250
114
114
68
250
Median
7.6
44
66
101
102
Range
7.1-8.0
24-58
30.90
96-105
73-121
aFisher pH meter Model 150.
bEDTA Titrametric method (American Public Health Association et al. 1975).
GMethyl Orange indicator nethod (American Public Health Association et al.
1975).
dYSI Model 54A.
eYSI Model 33, S-C-T meter.
-72-
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Fish
Young bluegill (7 to 10 cm) were captured with beach seines and fyke
nets in Lakes Wingra and Mendota in Madison, Wis., and placed in holding
tanks for at least 1 month before an experiment. The fish were fed trout
pellets once daily and were acclimated to experimental light and water
conditions. Fish were not fed during the selection experiments.
Procedure
The experiment was started by placing one fish in each aquarium. One
side of each tank was set at 21°C, the other at 23°C. After 1 day of fixed
temperatures, each fish was allowed to thermoregulate for 2 days before zinc
was added to half of the tanks, selected at random. Enough zinc sulfate (Zn
SO,) dissolved in distilled water was added to both compartments to achieve
a concentration of 2.5 ppm zinc in half of the aquaria, selected randomly.
Zinc was added only once at the start of the experiment. Distilled water
was added to the control aquaria. Water in both compartments of each
aquarium was analyzed for zinc daily (Table 19). Fish were allowed to
thermoregulate for 4 more days. At the end of the experiment, the bluegill
were weighed, measured, and frozen for analysis of zinc concentrations in
the gills, liver, and muscle (Table 20).
TABLE 19. ZINC CONCENTRATIONS (ppm) OF
WATER IN TREATMENT AND CONTROL AQUARIA3
Zinc-exposed
(n-97)
Control
(n=86)
Day
Median
Quartiles
Median
Quartiles
4
5
6
7
8
1.49
1.24
1.13
1.16
1.00
1.33-1.55
1.13-1.31
1.02-1.28
0.99-1.36
0.91-1.21
0.03
0.04
0.04
0.03
0.03
0.02-0.11
0.02-0.06
0.01-0.13
0.06-0.15
0.00-0.11
Analyses were done by atomic absorption photospectrometry.
RESULTS
After the zinc was added, its concentration in the water continuously
decreased probably because of absorption on particulate matter and uptake by
the fish (Table 19). Neither selected temperatures (Figure 19) nor rates of
activity (number of tunnel crossings/h) (Figure 20) of control and zinc-
exposed fish differed significantly (Mann-Whitney U Test, p < 0.05, Siegel
1956). Preferred temperatures on the day before zinc was added and of the
day zinc was added were not significantly different in either the control or
the zinc-exposed fish (Wilcoxon signed ranks test). There was a trend for
-73-
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30
Ul
cc
28
HI
Q
HI
26
(2.5 ppm Zn
added)
+4-
23.8
1
N-8
Control O O
Zn
N=9
4 5
DAY
Figure 19. Median selected temperatures of bluegill in control aquaria and in aquaria treated with
zinc in a 7-day experiment. Vertical bars represent the 25th to 75th percentiles.
-------
cc
040
DC
LU
CDC/)
IS
30
20
HI
LU
I-
LJ_
O
10
Control O O
N=8
Zn
N=10
( 2.5 ppm Zn added)
I TI
DAY
Figure 20. Median number of tunnel passes per hour by bluegill in control aquaria and in aquaria
treated with zinc in a 7-day experiment. Vertical bars represent the 25th to 75th
percentiles.
-------
TABLE 20. ZINC TISSUE CONCENTRATIONS (ppm) AT THE END OF THE EXPERIMENT
FOR RANDOMLY SELECTED FISH FROM TREATMENT AND CONTROL TANKS3
Zinc-treated
(n-4)
Control
(n=4)
Median
Range
Median
Range
Gill
Liver
Muscle
90.4
107.5
28.6
75.3-107.2
90.7-129.8
23.0-33.8
86.8
91.2
30.8
76.9-89.81
83.6-125.2
26.7-34.9
aAnalyses were done by neutron activation and by each median n=4.
the zinc-exposed fish to prefer a lower temperature, but this lasted for
only 1 day. The rate of activity of the control fish, however, was
significantly lower (P <^ 0.5) on the day distilled water was added than on
the previous day. This followed a trend of decreasing activity rates
(Figure 20). Activity rates of the zinc-^exposed fish were not significantly
lower after the addition of zinc; in fact, the median was greater than that
of the previous day.
DISCUSSION
Our selection system does not detect sublethal effects of zinc at lower
concentrations than other methods tested in water of similar quality.
Cairns et al. (1973) detected sublethal zinc concentrations of 2 to 3 ppm by
continuously monitoring bluegill movement patterns perceived by light-beam
interruptions, and by measuring bluegill ventilation rates. Sprague (1968)
found no change in selected temperature of Atlantic salmon (Salmo salar) in
a horizontal temperature gradient after 24 h of exposure to 0.16 ppm zinc.
The effect of temperature on zinc lethality varies with the type of
lethality test and species of fish, but seems greater at higher
temperatures. Survival time of rainbow trout exposed to zinc decreases at
higher temperature (Lloyd 1960). Temperature stress induced by increasing
the temperature at a rate of 1.5°C every 10 min reduced survival time at a
concentration of 32 ppm zinc (Burton et al. 1972). Also, at 5.6 ppm zinc
deaths occurred in 96 h at 30°C but not at 20°C. Pickering and Henderson
(1966) found no significant difference in toxicity to fish at 15° and 25°C,
but the trend was for higher toxicity at higher temperatures. Cairns and
Scheier (1957) found 100% survival at 18° and at 30°C for overlapping zinc
concentrations.
Apparently, a bluegill that behaviorally reduced its temperature while
exposed to lethal zinc concentrations would increase its probablity of
survival. A nonsignificant trend toward lower temperatures in the presence
of zinc was observed in both Iteterson's (1976) and the current work, perhaps
indicating that further experiments could determine which zinc concentration
results in a change in temperature preference.
-76-
-------
In conclusion, our method does not appear to be any more suitable as a
sensitive indicator of sublethal effects of metal ions than other methods.
-77-
-------
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APPENDIX A
NUMBER OF FISH CAUGHT AT EACH SAMPLING STATION, 1976-78
TABLE A-l. NUMBER OF FISH CAUGHT AT EACH SAMPLING STATION,1976
Station
Species
Northern pike (Esox lucius)
Walleye (Stizostedion vitreum vitreum)
Muskellunge (Esox masquinongy)
Largemouth bass (Micropterus salmoides)
Rainbow trout (Salmo gairdneri)
Yellow perch (Perca flavescens)
Spotted sucker (Minytrema melanops)
Pirate perch (Aphredoderus sayanus)
White sucker (Catostomus commersoni)
Black crappie (Pomoxis nigromaculatus)
White crappie (Pomoxis annular is)
Rock bass (Ambloplites rupestris)
Pumpkinseed (Lepomis gibbosus)
Bluegill (Lepomis macrochirus)
Yellow bullhead ( Ictalurus natalis)
Black bullhead (Ictalurus melas)
Redhorse (Moxostoma sp.)
Bowfin (Amia calva)
Carp (Cyprinus carpio)
Buffalo (Ictiobus cyprinellus)
1
1
0
0
0
2
0
0
0
0
0
0
1
0
0
0
0
0
0
0
0
2
21
2
0
0
1
0
2
0
12
20
1
5
0
0
2
3
1
8
many
many
3
8
3
6
1
0
0
9
3
3
0
0
0
22
4
0
12
0
2
many
many
4
18
1
3
0
0
1
2+a
0
3
1
0
0
2
0
0
11
0
0
many
many
+ signifies more than two, but exact number not recorded.
-85-
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TABLE A-2. NUMBER OF FISH CAUGHT AT EACH SAMPLING STATION, 1977
Station
Species
Northern pike (Esox lucius)
Walleye (Stizostedion vitreum vitreum)
Muskellunge (Esox masquinongy)
Largemouth bass (Micropterus salmoides)
Rainbow trout (Salmo gairdneri)
Yellow perch (Perca flavescens)
Spotted sucker (Minytrema melanops)
Pirate perch (Aphredoderus sayanus)
White sucker (Catostomus commersoni)
Black crappie (Pomoxis nigromaculatus)
White crappie (Pomoxis annularis)
Rock bass (Ambloplites rupestris)
Pumpkinseed (Lepomis gibbosus)
Bluegill (Lepomis macrochirus)
Yellow bullhead (Ictalurus natalis)
Black bullhead (Ictalurus melas)
Redhorse (Moxostoma sp.)
Bowfin (Amia calva)
Carp (Cyprinus carpio)
Buffalo (Ictiobus cyprinellus)
White bass (Morone chrysops)
Golden shiner (Notemigonus crysoleucas)
Chestnut lamprey ( Ichthyomyzon castaneus)
Freshwater drum (Aplodinotus grunniens)
1
10
0
0
0
2
0
0
0
11
0
0
2
0
0
0
0
1
1
0
0
0
0
0
0
2
45
4
30
7
7
3
54
0
69
25
0
20
40
4
4
3
0
16
5
0
1
1
3
1
3
2
0
0
0
0
0
0
0
0
0
0
0
0
1
0
1
0
0
0
0
0
0
0
0
4
7
0
0
0
0
0
0
0
0
0
0
0
0
0
0
0
1
2
0
0
0
0
0
0
-86-
-------
TABLE A-3. NUMBER OF FISH CAUGHT AT EACH SAMPLING STATION, 1978
Station
Species
Northern pike (Esox lucius)
Walleye (Stizostedion vitreum vitreum)
Muskellunge (Esox masquinongy)
Largemouth bass (Micropterus salmoides)
Rainbow trout (Salmo gairdneri)
Yellow perch (Barca flavescens)
Spotted sucker (Minytrema melanops)
White sucker (Catostomus commersoni)
Pirate perch (Aphredoderus sayanus)
Black crappie (Pomoxis nigromaculatus)
White crappie (Pomoxis annularis)
Rock bass (Ambloplites rupestris)
Pumpkinseed (Lepomis gibbosus)
Bluegill (Lepomis macrochirus)
Yellow bullhead (Ictalurus natalis)
Black bullhead (Ictalurus melas)
Redhorse (Moxostoma sp.)
Bowfin (Amia calva)
Carp (Cyprinus carpio)
Buffalo (Ictiobus cyprinellus)
Brown trout (Salmo trutta)
White bass (Morone chrysops)
Freshwater drum (Aplodinotus grunniens)
1
0
0
0
0
0
0
0
1
0
0
0
0
0
0
0
0
0
1
0
0
0
0
0
2
101
1
0
0
2
0
42
40
0
20
0
2
0
3
11
2
0
14
21
0
1
1
6
3
7
0
0
0
0
0
0
0
2
0
0
0
0
0
0
0
0
0
0
0
0
0
0
4
12
0
0
0
0
0
0
0
0
0
0
0
0
1
0
0
0
0
0
0
0
0
0
-87-
-------
APPENDIX B
MARSHES NEAR THE COLUMBIA GENERATING STATION
TABLE B-l. MARSHES NEAR THE COLUMBIA GENERATING STATION
Marsh area Description
Station site
(after construction)
Duck Creek
Rocky Run A
Rocky Run B
Corning—Weeting Lakes
The area north of County J, west of the Chicago,
Milwaukee, St. Paul and Pacific Railroad tracks,
south of Duck Creek. Includes the mouths of
Rocky Run and Duck Creek.
Includes all wetlands along Duck Creek east of
the Chicago, Milwaukee, St. Paul and Pacific
Railroad tracks, up to the dam located along State
Hwy 22-24 at Wyocena, Wis.
Wetlands south of County Hwy J, east of the
Chicago, Milwaukee, St. Paul and Pacific Railroad
tracks and west of State Hwy 51. Includes
wetlands drained by the mint drain and by
Rocky Run Creek. Fyke nets set in both the mint
drain and Rocky Run showed few northern pike moved
this far upstream to spawn.
Wetlands along Rocky Run Creek east of State Hwy 51
including Mud Lake. Unlikely to be pike spawning
habitat for reasons given for Rocky Run A.
Wetlands associated with Corning and Weeting Lakes
located north of the Wisconsin River and
west of Portage, Wis. Accessible by a small
creek flowing 8 km (about 5 miles) south of the
Wisconsin River. Ground survey in August 1978
showed no obstructions to fish movement, but very
shallow stream flow (4 to 5 cm deep, 1 m across) in
upper reaches of the creek.
(continued)
-88-
-------
TABLE B-l (continued)
Marsh area
Description
Powers Creek
Whelen Bay
Hinkson-Rowan Creek
Lodi Marsh
O'Kee Bay
South Dekorra
Inlet
Merrimac Inlet
Prentice Creek
Baraboo River Mouth
The mouth of Powers (Rowen) Creek east to
Interstate 90-94. Includes a portion of
Lake Wisconsin known as Whelen Bay.
Upstream tributaries of Powers Creek starting
at Interstate 90-94 eastward to stream headwaters.
Includes wetlands along Spring Creek from
its mouth at the Wisconsin River to marsh
upstream of the town of Lodi, Wis.
Ground survey in August 1978 indicated two small
spillways of about 0.5 m in the town of
Lodi which would prevent upstream migration except
during spring floods. Fish would have to migrate
about 8 km (5 miles) upstream through the town
of Lodi to reach suitable spawning habitat.
Wetlands associated with a bay of Lake Wisconsin
east of O'Kee, Wis., and a bay
east of Pine Bluff at Harmony Grove, Wis.
Wetlands associated with a small stream south of
Dekorra, Wis., and just south of where Interstate
90-94 crosses the Wisconsin River.
A bay of Lake Wisconsin and associated wetlands
south of Merrimac, Wis. State Hwy 113-78 and
the Chicago-Northwestern Railroad tracks
cross the bay, but do not prevent access
by spawning fish.
Wetlands associated with Prentice Creek which
joins the Wisconsin River north of Merrimac, Wis.
Wetlands located at the mouth of the Baraboo
River upstream to where Interstate 90-94
crosses the river.
-89-
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APPENDIX C
REVIEW OF LITERATURE ON ENTRAINMENT FROM COOLING LAKE INTAKE STRUCTURES
In the appendix, the possible entrainment damage to fish and
invertebrate populations at the Columbia site is discussed; the possible
damage appears minimal. In addition, a 1978 study of fish entrainraent at
the site by Swanson Environmental, Inc., has revealed that fish loss due to
the present water-intake systems is minor.
The effects of cooling-water intake on aquatic systems have been
studied at many power plants over the last 20 years. Although the studies
differed in their approach, detail, and conclusions, four general areas of
concern have emerged: (1) Removal of animals suspended or swimming in the
water column; (2) mechanical injury because of impingement upon intake
screens or abrasion in pumps, pipes, and condensers; (3) the toxic effects
of biocides used in reducing the fouling of pipe systems by microorganisms;
and (4) the various effects of thermal shock during condenser passage.
The removal of animals from the water column, including the impingement
of adult and juvenile fish, has become the focus of a federally mandated
monitoring program, pursuant to the requirements of Public Law 92-500.
Freeman and Sharma (1977) have conducted a survey of these programs, but a
summary volume is not complete. The removal aspect of cooling-water intake
is relevant to the Columbia site; mechanical, toxic, and thermal aspects of
entrainment do not apply. The Columbia station withdraws water from the
artificial cooling lake to cool the superheated steam in the turbines. It
is essentially a closed system, except that evaporative losses from the lake
require a constant input from the Wisconsin River. The "make-up" water is
presently pumped from the intake channel to the artificial lake by two
10,000-gal/min pumps. Water is drawn down an intake channel that connects
with the river approximately 3,000 ft from the cooling lake. The channel is
protected by two bar-grilles and a fish conservation traveling screen.
Studies of mechanical injury and mortality during entrainment have been
reported by Marcy (1973, 1976), Carpenter et al. (1974), Ginn et al.
(1974), King (1974), Davies and Jensen (1975), and Polgar (1975). Several
reviews such as those of Coutant (1970) and Hillegas (1977) have been
published. Although survival of damaged organisms is often quite low, it
does not appear that the numbers of organisms lost results in serious
effects on the aquatic systems.
Biocides such as chlorine are usually used at such low concentrations
that they pose no threat to entrained organisms or to the receiving body of
water (Marcy 1971, Bass and Heath 1975, Basch and Truchan 1976, Brungs 1976,
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Seegert and Brooks 1977). However, thermal shock, combined with small
amounts of chlorine, has a greater effect than either increased temperatures
or chlorine levels alone (Eiler and Delfino 1974, Ginn et al. 1974).
Cooling-systems designers now use predictive tools to minimize
impact. Curves and models predict for a given intake design the amount of
mechanical damage (Polgar 1975) and the extent of lethal and sublethal
thermal effects (Coutant 1970) expected. Models have also been developed by
Goodyear (1977), Christenson et al. (1977), and others to forecast effects
of removal on given fish populations.
POTENTIAL EFFECTS OF COOLING WATER INTAKE AT THE COLUMBIA SITE
Effects of entrainment of aquatic organisms from the Wisconsin River by
the Columbia Generating Station are different from effects seen at most
other generating stations. At Columbia there is no direct return of the
entrained water to the river. The analogy of the intake acting as a large
predator on the river ecosystem (Coutant 1970) is nore applicable than in
"once through" cooling situations. In assessing potential effects
researchers often draw a relationship between the percentage of water in the
river used and the resulting effect on the river. However, organisms in
riverine communities typically show "patchy" distributions (Whitton 1975),
and larger organisms can either avoid the intake channel or electively swim
into it.
Zooplankton and Drifting Macroinvertebrates
Zooplankton are too small to be screened out of the intake pumps and
are less able to avoid the influence of the pumping current than are larger
animals. The percentage of total river flow removed by the intake water at
Columbia presently averages 0.3% with a maximum of 1.08%. Assuming that the
number of organisms entrained by the Columbia intake is proportional to the
volume of river water used, we expect no significant loss of invertebrates
from the Wisconsin River. Several other entrainment studies at U.S. power
plants (King 1974, Davies and Jensen 1975, Hillegas 1977) did not
demonstrate measurable effects in downstream plankton communities even where
abundant data were available and generating stations in question diverted up
to 30% of the river flow.
Adult and Juvenile Fish
A 1-yr study of fish entrainnent at the Columbia site (Swanson
Environmental, Inc. 1977) reported the number, species, length, and
reproductive condition of fish impinged on the temporary screen box unit and
on the traveling screen unit currently in use. Sampling was conducted for a
24-h continuous period once a week. An estimated 14% of the total intake
volume was sampled. The catch numbers were extrapolated to estimate total
annual impingement as 668 + 387 fish/yr (mean +90% confidence limits). The
number of adult and juvenile fish impinged at Columbia is low, and even if
all impinged fish die, no effect on the river system should occur.
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Fish Eggs and Larvae
The Swanson Environmental, Inc., study (1977) also included sampling
for fish eggs and larvae. Submersible pumps were mounted behind the
traveling screen unit and pumped the sample water into 423-ym nets. Pump
rates were sufficient to prevent fish from avoiding the sampler. Estimated
annual entrainment of larval fish was 126,659 + 93,994 larvae/yr (mean +90%
confidence limits). No northern pike or walleye larvae were caught in the
samples. According to a summary of fish-census data for the Columbia site
(Wisconsin Department of Natural Resources 1973), northern pike and walleye
spawn in the wetland adjacent to Duck Creek. The mouth of Duck Creek is
located just upstream from the Columbia intake (Figure 1). Northern pike
larvae and fry remain on the spawning marshes until they attain a size of
20 mm at 16-24 days after hatching (Franklin and Smith 1963). Although
emigrating larvae of this size would not be able to avoid the intake
current, the river currents may be strong enough in early spring to sweep
larvae past the intake. Larval walleye are known to migrate from their
spawning marshes in intermittent pulses over a 10- to 15-day period (Priegel
1970). By sampling once every 7 days, the period of walleye larval
entrainment could have been missed. Walleye larvae may also avoid
entrainment by staying in the main currents as they enter the Wisconsin
River, therefore bypassing the shoreline by the intake. Newly hatched
walleye larvae emerging from similar spawning situations on the Wolf and Fox
Rivers in Wisconsin tended to stay in the strongest currents until they
reached more lacustrine situations where zooplankton were abundant (Priegel
1970).
In summary, as long as the Columbia intake continues to remove a small
percentage of the river flow, we expect no measurable effects of entrainment
on the river system. An exception might occur when organism distribution is
patchy near the intake, and a significant portion of one year-class (e.g.,
walleye larvae) is entrained. Aside from acting as a predator by removing
organisms from the Wisconsin River, the usual types of entrainment effects
(mechanical, toxic, and thermal) do not apply to the Columbia station.
BIBLIOGRAPHY FOR ENTRAINMENT
Basch, R. E., and J. G. Truchan. 1976. Toxicity of chlorinated condenser
cooling waters to fish. EPA-600/3-76-009, EPA Environmental Research
Laboratory, Duluth, Minnesota.
Bass, M. L., and A. G. Heath. 1975. Toxicity of intermittent chlorine
exposure to bluegill sunfish, Lepomis maorooh-irus' Interaction with
temperature. ASB Bull. 22:40.
Brungs, W. A. 1976. Effects of wastewater and cooling water chlorination
on aquatic life. EPA-600/3-76-098, EPA Environmental Research
Laboratory, Duluth, Minn.
Carpenter, E. J., B. B. Peck, and S. J. Anderson. 1974. Survival of
copepods passing through a nuclear power station on northeastern Long
Island Sound, U.S.A. Marine Biol. 24:49-55.
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Christenson, S. W., D. L. DeAngelis, and A. G. Clark. 1977. Development of
a stock progeny model for assessing power plant effects on fish
populations. In: Webster van Winkle (ed.) Proceedings of the
conference on assessing the effects of power plant-induced mortality on
fish populations, Gatlinburg, Tenn. Pergamon Press, Inc., New York.
Coutant, C. C. 1970. Biological aspects of thermal pollution. In:
Entrainment and discharge canal effects. Chemical Rubber Co. Grit. Rev.
Environ. Control 1(3):341-348.
Davies, R. M., and L. D. Jensen. 1975. Zooplankton entrainment at three
mid-Atlantic power plants. J. Water Pollut. Control Fed. 47:2130-2142.
Eiler, H. 0., and J. J. Delfino. 1974. Limnological and biological studies
of the effects of two modes of open-cycle nuclear power station
discharge on the Mississippi River (1969-1973). Water Res. 8:995-1005.
Franklin, D. R., and L. L. Smith, Jr. 1963. Early life history of the
northern pike, Esox lucius L. , with special reference to the factors
influencing the numerical strength of year-classes. Trans. Am. Fish.
Soc. 92:92-110.
Freeman, R. F., and R. K. Sharma. 1977. Survey of fish impingement at
power plants in the United States. Vol. II. Inland waters. ANL/ES-56
Vol. II. Argonne National Laboratory, Argonne, 111. 328 p.
Ginn, T. C., W. T. Waller, and G. L. Lauer. 1974. The effects of power
plant condenser cooling water entrainment on the amphipod Gammarus sp.
Water Res. 8:937-945.
Goodyear, C. P. 1977. Assessing the impact of power plant mortality on the
compensatory reserve of fish populations. In: Webster Van Winkle (ed.)
Proceedings of the conference on assessing the effects of power plant-
induced mortality on fish populations. Gatlinburg, Tenn. Pergamon
Press, Inc., New York.
Hillegas, J. M., Jr. 1977. Phytoplankton and zooplankton entrainment. A
summary of studies at power plants in the United States. Paper
presented at Savannah River Ecological Laboratory Symposium, Augusta,
Ga. (Preliminary draft).
King, J. R. 1974. A study of power plant entrainment effects on the
drifting macroninvertebrates of the Wabash River, M.S. Thesis, De Pauw
Univ., Greencastle, Ind.
Marcy, B. C. 1971. Survival of young fish in the discharge canal of a
nuclear power plant. J. Fish. Res. Board Can. 28:1057-1060.
Marcy, B. C. 1973. Vulnerability and survival of young Connecticut River
fish entrained at a nuclear power plant. J. Fish. Res. Board Can.
30:1195-1203.
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Marcy, B. C. 1976. Planktonic fish eggs and larvae of the lower
Connecticut River and the effects of the Connecticut Yankee Plant In:
D. Merriman and L. Thorpe (eds.) The impact of a nuclear power plant.
Connecticut River Ecological Study, Monogr. 1. Am. Fish. Soc. Bethesda,
Md.
Polgar, T. T. 1975. Assessment of near field manifestations of power
plants. Induced effects on zooplankton. In: Proceedings of the 2nd
Thermal Ecology Symposium, Augusta, Ga.
Priegel, G. R. 1970. Reproduction and early life history of the walleye in
the Lake Winnebago region. Wisconsin Dep. Nat. Resour. Tech. Bull. 45.
Seegert, G. L., and A. S. Brooks. 1977. The effect of intermittent
chlorination on fish: Observations 3 1/2 years, 17 species, and 15,000
fish later. Paper presented at the 39th Midwest Fish and Wildlife
Conference, Madison, Wis.
Swanson Environmental, Inc. 1977. Cooling lake make-up water intake
monitoring program, March 1976 - June 1977. Wisconsin Power and Light
Co., Columbia Energy Center, Portage, Wis., and Southfield, Mich.
Whitton, B. A. 1975. River ecology. Studies in ecology, Vol. 2. Univ.
California Press, Berkeley and Los Angeles.
Wisconsin Department of Natural Resources. 1973. Final draft environmental
impact statement for the Columbia Generating Station of the Wisconsin
Power and Light Company, Madison, Wisconsin.
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APPENDIX D
REVIEW OF LITERATURE ON ACID PRECIPITATION
Acid rainfall, the topic of this appendix, is not considered a
potential problem for aquatic ecosystems at Columbia because of the high
hydrogen-ion buffering capacity resulting from the calcareous nature of the
drainage basin.
Recent studies in both North America and Europe have documented the
occurrence of acid rains with a pH ranging from 2.1 to 5.0 (Likens and
Bormann 1974; Beamish 1974, 1976; Dickson 1975; Schofield 1976). Rainwater
is normally slightly acidic, with a pH of 5.7, as a result of the
equilibrium reaction between atmospheric carbon dioxide and water forming
carbonic acid (FUCO/). Both natural and anthropogenic processes, however,
can add three strong mineral acids, sulfuric, nitric, and hydrochloric, to
atmospheric water with a resulting sharp decrease in pH (Gorham 1976). The
most predominant of these acids is sulfuric (I^SO^), which can be formed in
substantial amounts from the sulfur dioxide (862) produced as sulfur in
fossil fuels oxidizes during combustion. Coal normally has between 1 and 3%
sulfur, but the percentage can go as high as 6%. Of less importance are
nitric acid (HNOo) and hydrochloric acid (HC1), which are also produced by
fossil-fuel combustion through the oxidation of organic nitrogen and
chlorine, respectively. These acids may then enter aquatic systems through
rainfall or, in northern latitudes, through spring ice and snow runoff.
The work of Cogbill and Likens (1974) illustrates that acid
precipitation is likely to remain a problem in certain areas. By graphing
isolines of rainfall pH falling over the eastern U.S., they have shown a
dramatic increase in the geographic area affected by acid rain, as well as
an increase in rainfall acidity for the 10-year period 1956-66.
The initial effects of acid input into lakes and streams depend largely
on edaphic characteristics that determine their buffering capacity. All
waters so far affected by acid precipitation have been in areas that are
geologically highly resistant to chemical weathering and usually have a low
concentration of major ions, particularly bicarbonate (HCO^), resulting in a
specific conductance less than 50 Jmhos/cm (Wright and Gjessing 1976). Acid
rainfall into such weakly buffered systems causes a loss of bicarbonate ion
and its replacement by sulfate; hence sulfate is the major anion in
acidified soft water, whereas bicarbonate predominates in non-acidified soft
water. Acidified lakes are frequently found to contain elevated aluminum
and manganese concentrations that are attributed to dissolution from
surrounding soils. Elevated levels of other heavy metals (Pb, Zn, Cu, Ni)
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may also exist downwind of major base-metal smelters (Van Loon and Beamish
1977).
Ecological studies concerned with acidification of aquatic ecosystems
have focused on fish poulations, since the loss of an exploitable fish
population is the most noticeable and economically important consequence of
acid precipitation. Fish loss is reported to be a gradual process resulting
not from acutely lethal pH changes, but rather from the failure to recruit
new year-classes into the population (Beamish 1974). At pH values above the
lower lethal level, interference with spawning has been demonstrated in both
laboratory and field studies (Mount 1973, Beamish 1976). The presumed
mechanism causing reproductive failure is disruption of normal calcium
metabolism that prevents females from releasing their ova (Beamish 1976).
Long-term effects of acidification on fish populations were summarized by
Beamish (1975) as follows: (1) failure to spawn, (2) low serum Ca levels
in mature females, (3) appearance of spinal deformities, (4) decreases in
the average size of year-classes, (5) reduction in population size, and (6)
disappearance of species from lakes.
Studies have indicated a genetic basis for acid tolerance at the
species level (Gjedrem 1976, Robinson et al. 1976, Schofield 1976) and
selective breeding of acid-tolerant fish strains has been proposed as a
means of stocking waters that have lost their natural populations. The
observed rates of population extinction indicate, however, that
acidification has been proceeding too rapidly for natural-selection
processes to be effective in maintaining fish populations under natural
conditions.
Equally as serious as damage to fish are the less conspicuous effects
of acid rain on aquatic organisms such as microdecomposers, primary
producers, zooplankton, and zoobenthos. Studies in six Swedish lakes, where
the pH decreased by 1.4 to 1.7 pH units in the last 40 years, have
demonstrated an inhibition of bacterial decomposition with a resultant
abnormal accumulation of coarse organic detritus (Hendrey et al. 1976a).
Rooted macrophytes, zooplankton, and benthic invertebrates are also stressed
by acidification of waters (Hendrey et al. 1976b). Some of the effects of
pH on aquatic organisms are summarized in Table D-l.
Although pH measurements of rainfall in the vicinity of the Columbia
Generating Station have not been made, it appears unlikely that acid
rainfall will noticeably affect nearby aquatic ecosystems for the following
reasons: (1) The Wisconsin River, Rocky Run Creek, and nearby waters are
well-buffered systems with total alkalinities in the range of 80 to 133
mg/liter CaCO-j and conductivities of 178 to 273 ^mhos/cm; (2) winds are
predominately from the west and south (Stearns et al. 1977) and, therefore,
power plant emissions should miss most of the nearby aquatic systems that
are located mainly west and south of the plant; (3) the present pH values of
the Wisconsin River (7.6 to 8.2) and Rocky Run Creek (7.6 to 8.2) are well
within the recommended safe range of 6.5 to 9.0 for natural waters and have
not changed noticeably since the plant began in 1975.
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TABLE D-l. SUMMARY OF pH EFFECTS ON AQUATIC ORGANISMS
pH
Effect
Reference
< 3.5 Unlikely that fish can survive for more
than a few hours
A few invertebrates (midges, mosquito,
caddisfly) have been found
Few plants (only mosses and algae) have
been found
EIFAC 1969
Lackey (1938)
Hendrey et al.
1976a
3.5-4.0 Lethal to salmonids and bluegills, limit
of tolerance of pumpkinseed, perch, and
pike, but reproduction is inhibited
Cattail (Typha) is the only higher plant
4.0-4.5 Only a few fish species survive, including
perch and pike
Lethal to fathead minnows
Flora is restricted
Some caddisflies and dragonfiles are found,
and midges are dominant
4.5-5.0 Salmonids may survive, but do not
reproduce
Benthic fauna are restricted; mayflies
are reduced
Fish populations are severely stressed;
a viable fishery is nonexistent
Snails are rare or absent
The fish community is decimated with
virtually no reproduction
White suckers and brown bullheads fail
to spawn, but perch do spawn
5.0-6.0 Rarely lethal to fish except some
salmonids, but reproduction is reduced
Larvae and fry of sensitive species may
be killed
Bacterial species diversity is decreased,
benthic invertebrates are reasonably
diverse, but sensitive taxa such as
mayflies are absent and molluscs are rare
Fathead minnow egg production and ability
to hatch are reduced
Smallmouth bass, walleye, and burbot stop
reproducing
Roe of roach (Rutelus rutelus) fail
to hatch
(continued)
U.S. EPA 1973
U.S. EPA 1973
U.S. EPA 1973
Hendrey et al.!976a
Beamish 1974, 1975
Beamish 1975
U.S. EPA 1973
flount 1973
Beamish 1976
tlilbrink and
Johansson 1975
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TABLE D-l (continued)
pH Effect Reference
6.0-6.5 Unlikely to be harmful to fish unless U.S. EPA 1973
free CO^ exceeds 100 ppm
Good invertebrate fauna except for
reproduction of Gammarus and Daphnia
Aquatic plants and microorganisms
relatively normal
6.5-9.0 Harmless to fish and most invertebrates
although 7.0 is near the lower limit for U.S. EPA 1973
Gammarus reproduction
Microorganisms and plants are normal
Toxicity of other substances
may be affected by pH shifts within
this range.
Future considerations should be given to the effect of added sulfur
emissions when Columbia II begins operation and on the contributions, if
any, of the Columbia plant emissions to acid rainfall over distant waters
such as northern Wisconsin lakes, some of which are poorly buffered and more
subject to acidification.
BIBLIOGRAPHY FOR ACID RAIN
Beamish, R. J. 1974. Loss of fish populations from unexploited remote
lakes in Ontario, Canada as a consequence of atmospheric fallout of
acid. Water Res. 8:85-95.
Beamish, R. J. 1975. Long-term acidification of a lake and resulting
effects on fishes. Ambio 4(2):98-102.
Beamish, R. J. 1976. Acidification of lakes in Canada by acid
precipitation and the resulting effect on fishes. Water Air Soil
Pollut. 6:501-514.
Cogbill, C. V., and G. E. Likens. 1974. Acid precipitation in the
northeastern United States. Water Resour. Res. 10:1133-1137.
Dickson, W. 1975. The acidification of Swedish lakes. Institute of
Freshwater Research, Drottningholm, Sweden. Rep. No. 54. p. 8-20.
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European Inland Fisheries Advisory Commission Working Party on Water
Quality. 1969. Water quality criteria for European freshwater fish:
Extreme pH values and inland fisheries. Water Res. 3:593-611.
Gjedrem, T. 1976. Genetic variation in tolerance of brown trout to acid
water. SNSF-project, Norway, FR5/76. 11 p.
Gorham, E. 1976. Acid precipitation and its influence upon aquatic
ecosystems: An overview. Water Air Soil Pollut. 6:457-481.
Hendrey, G. R., K. Baalsrud, T. S. Traaen, M. Laake, and G. Raddum.
1976a. Acid precipitation: Some hydrobiological changes. Ambio 5(5-
6):224-227.
Hendrey, G. R., R. Borgstrom, and G. Raddum. 1976b. Acid precipitation in
Norway: Effects on benthic faunal communities. Paper presented at the
39th Annual Meeting, Am. Soc. Limn, and Oceonography, Savannah, Ga.
Lackey, J. B. 1938. The flora and fauna of surface waters polluted by acid
mine drainage. Public Health Rep. 53:1499-1507.
Likens, G. E., and F. H. Bormann. 1974. Acid rain: A serious regional
environmental problem. Science 184:1176-1179.
Milbrink, G., and N. Johansson. 1975. Some effects of acidification on roe
of roach, Rutilus rutilus L., and perch, Peraa fluviatilis L., with
special reference to the Avad System in eastern Sweden. Institute of
Freshwater Research, Drottningholm, Sweden. Rep. No. 54.
Mount, D. I. 1973. Chronic effect of low pH on fathead minnow survival,
growth and reproduction. Water Res. 7:987.
Robinson, G. D., W. A. Dunson, J. E. Wright, and G. E. Mamolito. 1976.
Differences in low pH tolerance among strains of brook trout (Salvelinus
fontinalis) J. Fish Biol. 8:5-17.
Schofield, C. L. 1976. Acid precipitation: Effects on fish. Ambio 5(5-
6):228-230.
Stearns, C. R., B. Bowen, and L. Dzamba. 1977. Meteorology, p. 171-183.
In: Documentation of environmental change related to the Columbia
Electric Generating Station. 10th Semi-Annual Progress Report.
Institute for Environmental Studies, Univ. Wisconsin-Madison, Madison,
Wis. Rep. 82.
Van Loon, J. C., and R. J. Beamish. 1977. Heavy metal contamination by
atmospheric fallout of several Flin Flon area lakes, and the relation to
fish populations. J. Fish Res. Bd. Can. 34:899-906.
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Wright, R. F., and E. T. Gjessing. 1976. Acid precipitation: Changes in
the chemical composition of lakes. Ambio 5(5-6):219-223.
U.S. Environmental Protection Agency. 1973. Acidity, alkalinity, and pH,
p. 140-141. In: Water quality criteria. Ecol. Res. Ser., R3-73-033.
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APPENDIX E
REVIEW OF LITERATURE ON ALTERNATIVE DISPOSAL OF FLY ASH
Increased national emphasis on the use of coal to meet energy
requirements may result in a doubling of coal-ash production from 1975
levels by the year 1995 (PEDCO-Environmental, Inc. 1976). Annual coal-ash
production is currently estimated to be 61.9 x 10 tons (Davis and Faber
1977) and may be 100 x 106 tons by 1985 (Harriger 1977). About 20% of the
ash is used for commercial purposes in cement, asphalt and concrete,
fertilizer, fire control, road-bed stabilizer, soil aeration, and sanitary
landfill cover (PEDCO-Environmental 1976, Theis 1976a, Harriger 1977).
Research continues into additional uses for coal ash such as water
reclamation, sewage-sludge conditioning, and supplementation of soil sewage
micronutrients (Theis 1976a, Furr et al. 1977). Fly ash and lime cause
precipitation of phosphorus from natural waters, and the ash seals the
nutrients in the sediment; however, the side effects of such treatment may
be severe (Theis and De Pinto 1976). Fly ash concentrations of 10 to 20
g/liter were toxic to Stone Lake, Mich., fish. High pH, dissolved oxygen
depletion, heavy-metal release, and physical clogging and crushing of
organisms are other effects that have not been adequately investigated. Fly
ash applied to soils can neutralize acid soils and supply calcium and trace
elements (PEDCO-Environmental, Inc. 1976); however, the high conductivities
of fly-ash-water solutions may result in injuriously high salt
concentrations for many sensitive crops (Olsen and Warren 1976). Theis
(1976a) suggests the extraction of the following quantities of rare metals
from ash: 53.2 kg As/day, 5.2 kg Pb/day, 5.0 kg Cu/day, 49 kg Zn/day, 12.3
kg Cr/day, 730 g Cd/day, and 18.9 g Hg/day.
Despite continuing research the large excess of fly-ash production over
demand is likely to continue (Theis 1976a) and, coupled with an average rate
of ash production of 0.5 kg/kWh (PEDCO-Environmental, Inc. 1976), will
result in large amounts of ash to be disposed of in an environmentally sound
manner. The new source performance standards (NSPS) applicable to new power
plants prohibit discharges into natural waters from ash-settling ponds
(Dvorak and Pentecost 1977). To comply with these regulations, ash from
Unit II of the Columbia Generating Station is currently being held in a
segregated portion of the ash basin while a site for permanent land disposal
is sought and prepared.
Many concerns remain regarding the landfill disposal of coal ash. In
addition to the continued threat of surface contamination due to
precipitation and overland runoff, ground-water contamination and landfill
erosion are significant concerns. Although many of the principles of
sanitary landfilling are applicable if consideration is given to the
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different nature of the contaminants, an expanded study of coal-ash
landfills is needed. Information on the leaching and mobility of ash trace
constituents is limited (Dvorak and Pentecost 1977), and because of the
newness of the disposal method, little is known of the long-term effects of
such disposal. Such studies are needed for the creation of standards for
land disposal of toxic sutbstances, which is virtually unregulated at the
federal level (Fields and Lindsey 1975).
The most widespread concern about coal-ash landfilling is the potential
for ground-water contamination by leachate produced when water percolates
through the landfill. High salt concentrations in leachate may be a
significant problem, especially if it reaches ground-water supplies that are
already high in salt. Increased pH due to ash leachate may be a localized
problem (Olsen and Warren 1976), but pH is more important because of its
effects on metal solubilities and adscription. This potential for metal and
other trace-element contamination has received the greatest attention and
concern.
The ability of the soil to attenuate contaminants in the leachate is of
primary importance in preventing ground-water contamination by any kind of
landfill. Waldrip (1975) found that inorganic and organic materials from
sanitary landfill leachate are adsorbed by the soil, and many undesirable
ions are replaced by desirable ones in an ion-exchange process. He
concluded that most ground-water contamination is limited to_ the immediate
vicinity of the landfill because of slow movement of the ground water. The
low velocity allows sufficient time for ion exchange, dilution, and
dispersion to occur. The landfill contribution to ground-water supply is
significantly diminished within a few hundred feet of the landfill.
Griffin et al. (1976) studied the attenuation of metals and other
leachate constituents run through laboratory sediment columns. Clay was
relatively poor in reducing concentrations of Cl~, Na , and water-soluble
organic compounds, but K, NH,, Mg, Si, and Fe were moderately reduced in
concentration, probably by cation exchange with Ca in the soil. Low
leachate concentrations were strongly attenuated by small amounts of clay
possibly because of precipitation of the metals upon formation of metal
hydroxides or carbonates (caused by high pH and high bicarbonate
concentration in the^leachate). Low leachate concentrations of Al, Cu, Ni,
Cr, As, SO,, and PO, precluded interpretation for those substances. Suarez
(1974) describes the chemical reactions involving metals leached from
sanitary landfills and discusses their relationship with Eh, pH, and
dissolved oxygen.
A comparison of fly-ash landfill investigations is necessary to
determine the applicability of these sanitary landfill results to the
landfill designed expressly for fly ash. Theis (1976b) and Theis and Marley
(1976) discuss the potential for ground-water contamination from land
disposal of fly ash. They determined the important characteristics of ash
to be initial trace-metal concentration, acid-base characteristics, fly-ash
concentration in the aquatic system, and the size-fraction distribution of
the ash. A combination of field and laboratory studies demonstrated that
Cr, Cu, Hg, Pb, and Zn were either released from leachate in insignificant
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amounts or were rapidly sorbed onto soil particles. The metals As, Ni, and
Se, however, occurred in ground water at higher concentrations and appeared
able to migrate a greater distance. Sorptive processes could explain the
metal leachate behavior in the initial desorption of metals from the ash
into water and subsequent adsorption onto the soil phase.
The investigation of a landfill for fly ash from combustion of eastern
coal (Harriger 1977, Harriger et al. 1977) is the most comprehensive study
to date. The presence of clay-rich soil was determined to be the most
important factor affecting water quality. Other factors include composition
and quality of the ash, duration of exposure to leaching, pH, oxidation
conditions, and surface and ground-water flow patterns. Clay soils were
relatively impermeable and found to adsorb or exchange large quantities of
ions. Ground-water wells away from the landfill were lower in
concentrations of many trace substances, attesting to the benefits of
leachate percolation through the soil. Landfill wells often had
concentrations of As, Se, Fe, Mn, and SO- above the U.S. Public Health
Service drinking-water recommendations. Landfill wells also exhibited
higher concentrations of Zn, Ca, Cr, Cu, Mg, and K than the off-site
wells. The metals Ca, Cr, and Cu were fairly low, however, because of low
concentrations in the ash itself, good attenuation by clay, and the
prevailing pH conditions.
Analysis of surface waters (streams flowing across the landfill, runoff
from the landfill, and ponds formed from precipitation) indicated few
effects of the landfill once the water left the site. A stream enclosed by
pipe as it crossed the site appeared to receive some ground water and
ash-leachate seepage downstream. Concentrations of Fe, Mn, and SO^ exceeded
drinking-water standards, but decreased rapidly downstream. Levels of Ca,
Cd, Cu, Fe, Mg, Na, Se, Zn, and SO* were higher and pH was lower in ponds on
the landfill (especially those with exposed ash deltas) than in control
ponds away from the site. Even higher concentrations of metals occurred in
the sediments of the landfill ponds, indicating that the contaminants were
precipitating out of the water. Metal concentrations were high in runoff
water from the landfill, and concentrations higher than in ground water for
Cr, Cu, and Zn were evidence of attenuation by clay and restricted metal
mobility in ground water. Thus surface runoff must be contained to permit
these mechanisms to operate.
The pH and oxidation states of materials in the landfill influence the
effectiveness of the attenuation mechanisms. The solubility of most metal
ions is increased at lower pH values (Harriger 1977), and thus in acidic
leachate metals are not removed as readily by the attenuation processes.
Generally, high pH greatly decreases solubility, and only Zn and Cd are
considered soluble in the pH range 7 to 8.5 (Theis 1976a). Most Cr is
released from ash into the leachate at pH 3, although some is released at pH
6, 9, and 12 (Theis and Wirth 1977). Iron and Mn precipitate at pH greater
than 7.5 (Harriger 1977). Fields and Lindsey (1975) conclude that low pH
affects ion exchange and adsorption properties of soil. Clays are more
effective in adsorbing most metals when the pH is high, although a low pH is
best for adsorption of organics. They state that it is best to maintain
landfill soils at pH 7.0-8.0. Frost and Griffin (1977) found, however, that
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As and Se adsorption by clays is decreased at high pH. Oxidation causes the
formation of iron oxides and hydroxides; these precipitate from the leachate
and can adsorb other ions (Harriger 1977), thus increasing the purification
capacity of the soil.
The relative amounts of lime and amorphous iron oxides in the ash
determine the pH of the leachate. Western coals have high amounts of lime
(Theis and Wirth 1977), which account for the basic nature of the ash from
the Columbia station. The greatest environmental concern with ion pH ashes
is the large amount of surface leachable Fe (Theis and Wirth 1977). Theis
(1976a) states that a greater amount of metal is likely to be released from
ash into ground water than into surface water, because of the lower pH and
high CO^ content of ground water and the consequently greater likelihood of
ion exchange from ash into this water.
Research continues into the principles of site selection and design to
reduce as much as possible the threat of ground- and surface-water
contamination. Little is known about the potential environmental effects of
landfills in Wisconsin (Zaporozec 1974) and there have been few long-term
studies of solid waste disposal in the United States. Leachate production
occurs even in well-designed landfills, especially in humid areas such as
Wisconsin (Fields and Lindsey 1975, Zaporozec 1974), but this production can
be minimized or controlled with proper site selection and design.
Many investigators suggest the use of liners, either impervious to
retain all leachate, or permeable ones to supplement the ability of the soil
to attenuate pollutants (Fields and Lindsey 1975, Griffin et al. 1976,
PEDCO-Environmental, Inc. 1976, Dvorak and Pentecost 1977). Where clay in
native soils is insufficient, a clay liner can satisfactorily mitigate the
contamination threat. It has been suggested that ash landfills nay have the
capacity to seal themselves against leachate loss. As soluble CaO moves into
the soil and forms CaCO-j, the permeability of the soil may be significantly
reduced (Olson and Warren 1976). Fly ash is often deliberately applied to
sanitary landfills because of its moisture-adsorbing characteristics (PEDCO-
Environmental, Inc. 1976).
Other suggestions to reduce the potential of contamination include
vegetating the landfill to reduce erosion by wind or water. Harriger (1977)
found that erosion remained a problem when the ash was covered with bare
soil. PEDCO-Environmental, Inc. (1976) suggests the use of species tolerant
to high pH, boron, and salt. Recommendations for sanitary landfills in
southern Indiana include: use of upland sites to avoid runoff from upland
areas, sites with soils or intervening materials with high exchange and
adsorption capacities, and sites where the water table is much below the
bottom of. the waste; use of leachate lagoons to prevent surface-water
contamination, and avoidance of areas subject to flooding (Waldrip and Ruhe
1974). PEDCO-Environmental, Inc. (1976) presents a detailed discussion of
geological, chemical, and engineering aspects of landfill site selection and
design. A literature review by Heidman and Brunner (1976) lists references
concerning site locations, investigation, monitoring, and management for
sanitary landfills. Much of the information in both reports is also
applicable to coal-ash landfills.
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Several states and agencies have criteria and regulations that should
be considered in the construction of coal-ash landfills in Wisconsin. The
California State Water Resources Control Board (1975) lists the following:
(1) Underlying geological formations with questionable permeability must be
permanently sealed, or ground-water conditions must prevent hydrologic
continuity; (2) leachate and subsurface flow must be self-contained; (3)
sites must not be located over zones of active faulting; (4) limitations are
applied if the area is in a 100-yr (or more frequent) flood-frequency class.
The U.S. Environmental Protection Agency (1973) recommends the following
criteria: (1) Low population density; (2) low alternate land-use value; (3)
low ground-water-contamination potential; (4) away from flood plains,
excessive slopes, and natural depressions; (5) soil with high clay content;
(6) adequate distance from human and livestock water supplies; (7) areas of
low rainfall and high evaporation rates, where possible; (8) sufficient
elevation over the water table; (9) no hydrologic connection with ground or
surface water; (10) use of encapsulation, liners, waste detoxification, or
solidification/fixation where necessary; (11) adequate monitoring.
Consideration of all these suggestions will significantly reduce, if not
avoid entirely, the adverse effects that a fly-ash landfill might have on
environmental quality.
It appears that the high pH of the ash expected from Columbia II will
substantially reduce the pollution potential from a landfill. The landfill
site must be chosen carefully, however, to avoid direct connection with the
ground water. A clay or other type of liner will probably be beneficial, if
not required, to avoid ground-water contamination. Pipes to collect and
recirculate leachate should be used if there is any likelihood of less than
complete metal attenuation by the time the leachate reaches the ground
water.
SUMMARY
1. Fly ash may be used commercially for a variety of purposes but supply
will probably continue to exceed demand (Theis 1976a, PEDCO-
Environmental, Inc. 1976, Theis and De Pinto 1976, Harriger 1977).
2. Although recent air and water pollution standards prohibit the discharge
of ash or its leachate into surface waters, considerable concern has
arisen over the potential adverse effects of the dry disposal of fly ash
in landfills.
3. Metal and trace-element contamination of water, particularly ground
water, is the most serious concern. Soils vary widely in their
abilities to attenuate these pollutants.
4. Clay soils have the greatest capacity for metal adsorption and ion
exchange (Griffin et al. 1976, Theis 1976b, Theis and Marley 1976,
Harriger 1977).
5. Because of these mechanisms, and the dilution and dispersion in slow-
moving ground water, most ground^water contamination is limited to the
immediate vicinity of the landfill (Waldrip 1975, Harriger 1977).
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6. With proper precautions direct surface-water contamination is usually
minimal (Harriger 1977). Appropriate precautions include containment of
surface runoff and avoidance of low sites and steep slopes.
7. Better attenuation of metals is usually obtained when the leachate has a
high pH. This is caused by the reduced solubility of metals and
improved properties of clay under these conditions (Fields and Lindsey
1975, Theis 1976a, Harriger 1977). Fortunately, coal burned at the
Columbia Generating Station produces basic conditions in its ash.
8. Where natural soils are not sufficient, clay or impervious liners should
be applied to the landfill (PEDCO-Environmental, Inc. 1976, Dvorak and
Pentecost 1977). Fly ash appears to have some capacity to form a seal
itself (Olson and Warren 1976).
9. Other recommendations to reduce the potential environmental
contamination include covering with soil; encouraging vegetation;
containing leachate; adequate monitoring; and avoiding sites with high
ground water; flooding potential; active faulting; or low elevations.
BIBLIOGRAPHY FOR FLY ASH
California State Water Resources Control Board. 1975. Disposal site design
and operation information. Sacramento, Calif, p. 19-21.
Davis, J. E., and J. H. Faber. 1977. Annual report: National Ash
Association. National Ash Assoc., Washington, D.C.
Dvorak, A. J., and E. D. Pentecost. 1977. Assessment of the health and
environmental effects of power generation in the Midwest. Vol. II.
Ecological effects. Draft. Argonne National Laboratory, Argonne,
111. 169 p. (Permission obtained.)
Fields, T., and A. W. Lindsey. 1975. Landfill disposal of hazardous
wastes: A review of literature and known approaches. U.S.
Environmental Protection Agency, EPA/530/SW-165. Cincinnati, Ohio.
36 p.
Frost, R. R., and R. A. Griffin. 1977. Effect of pH on adsorption of
arsenic and selenium from landfill leachate by clay minerals. J. Soil
Sci. Soc. Am. 41:53-57.
Furr, A. K., T. F. Parkinson, P. A. Hinrichs, D. R. Van Campen, C. A. Bache,
W. H. Gutenmann, L. E. St. John, Jr., I. S. Pakkala, and D. J. Lisk.
1977. National survey of elements and radioactivity in fly ashes.
Environ. Sci. Technol. 11:1194-1201.
Griffin, R. A., K. Cartwright, N. F. Shimi, J. D. Steele, R. R. Ruch,
W. A. White, G. M. Hughes, and R. H. Gilkeson. 1976. Attenuation of
pollutants in municipal landfill leachate by clay minerals. Part 1:
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Column leaching and field verification. Illinois State Geol. Surv.
Environ. Geol. Notes, No. 78. 34 p.
Harriger, T. L. 1977. Impact on water quality by a coal ash landfill in
north central Chautaqua County, New York. Ph.D. Thesis, State
University College, Fredonia, N. Y. 192 p.
Harriger, T. L., W. M. Benard, D. R. Corbin, and D. A. Watroba. 1977.
Impact of a coal ash landfill on water quality in north central
Chautaqua County, New York. Symposium on Energy and Environmental
Stress in Aquatic Systems. Savannah River Ecology Laboratory.
(Abstracts).
Heidman, J. A., and D. R. Brunner. 1976. Solid wastes and water quality.
J. Water Pollut. Control Assoc. 48:1299.
Olson, R. A., and G. Warren. 1976. Aquatic pollution potential of fly ash
particles, p. 91-112. In: Toxic effects on the biota from coal and oil
shale development. Nat. Res. Ecol. Lab., Colorado State Univ.,
Internal Proj. Rep. No. 7, Ft. Collins, Colo.
PEDCO-Environmental, Inc. 1976. Residual waste best management
practices: A water planner's guide to land disposal. U.S.
Environmental Protection Agency, EPA/440/9-76/022, Cincinnati, Ohio.
SuareE, D. L. 1974. Heavy metals in waters and soils associated with
several Pennsylvania landfills. Ph.D. Thesis, Pennsylvania State Univ.,
University Park, Pa. 222 p.
Theis, T. L. 1976a. Potential trace metal contamination of water resources
through disposal of fly ash. Notre Dame Univ., CONF-750530-3, South
Bend, Ind. 21 p.
Theis, T. L. 1976b. Contamination of ground water by heavy metals from the
land disposal of fly ash. Tech. Prog. Rep. 1 June 1976 to 31 August
1976. Prepared for U.S. Energy Research and Development Administration,
Notre Dame Univ., South Bend, Ind. 44 p.
Theis, T. L., and J. V. DePinto. 1976. Studies on the reclamation of Stone
Lake, Michigan. U.S. Environmental Protection Agency, Ecol. Res. Ser.,
EPA-600/3-76-106, Cincinnati, Ohio. 84 p.
Theis, T. L., and J. J. Marley. 1976. Contamination of ground water by
heavy metals from the land disposal of fly ash. Tech. Prog. Rep.
1 June 1976 to 29 February 1976. Prepared for U.S. ERDA, Notre Dame
Univ., South Bend, Ind. 21 p.
Theis, T. L., and J. L. Wirth. 1977. Sorptive behavior of trace metals on
fly ash in aqueous systems. Environ. Sci. Technol. 11:1096-1100.
U.S. Environmental Protection Agency. 1973. Acidity, alkalinity, and pH,
p. 140-141. In: Water quality criteria. Ecol. Res. Ser. R3-73-033.
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Waldrip, D. B. 1975. The effect of sanitary landfills on water quality in
southern Indiana. Ph.D. Thesis, Indiana Univ., Bloomington, Ind.
160 p.
Waldrip, D. B., and R. V. Ruhe. 1974. Solid waste disposal by land burial
in southern Indiana. Water Resour. Res. Center, Tech. Rep. No. 45.
Purdue Univ., West Lafayette, Ind. 110 p.
Zaporozec, A. 1974. Hydrogeologic evaluation of solid waste disposal in
south central Wisconsin. Wisconsin Dep. of Nat. Resour., Tech. Bull.
78. 31 p.
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TECHNICAL REPORT DATA
(Please read Instructions on the reverse before completing)
1. REPORT NO.
EPA-600/5-80-078
2.
3. RECIPIENT'S ACCESSION NO.
4. TITLE AND SUBTITLE
Ecological Studies of Fish Near a Coal-Fired
Generating Station and Related Laboratory Studies
Wisconsin Power Plant Impact Study
5. REPORT DATE
July 1980 issuing date
6. PERFORMING ORGANIZATION CODE
7. AUTHOR(S)
John J. Magnuson, Frank J. Rahel, Michael J. Talbot,
Anne M. Forbes, Patrica A. Medvick
8. PERFORMING ORGANIZATION REPORT NO.
9. PERFORMING ORGANIZATION NAME AND ADDRESS
Department of Limnology
University of Wisconsin
Madison, WI 53706
10. PROGRAM ELEMENT NO.
1BA820
11. CONTRACT/GRANT NO.
R803971
12. SPONSORING AGENCY NAME AND ADDRESS
Environmental Research Laboratory
Office of Research and Development
U.S. Environmental Protection Agency
Duluth, Minnesota 55804
13. TYPE OF REPORT AND PERIOD COVERED
14. SPONSORING AGENCY CODE
EPA/600/03
15. SUPPLEMENTARY NOTES
16. ABSTRACT
Construction of a coal-fired electric generating station on wetlands adjacent to the
Wisconsin River has permanently altered about one-half of the original 1,104-ha site.
Change in the remaining wetlands continues as a result of waste heat and ashpit efflu-
ent produced by the station. Leakage of warm water from the 203-ha cooling lake is
causing a shift in the wetlands from shallow to deep-water marsh. Coal-combustion
byproducts enter the wetlands from the station's ashpit drain. Since this area was
known to have a diverse fish community and to be a spawning ground for Wisconsin River
game fish, we studied the effects of this habitat loss and degradation on fish popula-
tions. In laboratory experiments we investigated the use of temperature preference
and activity as a sublethal bioassay. In selection experiments we examined the
potential of fish to evolve metal-tolerant populations in chronically contaminated
environments.
17.
KEY WORDS AND DOCUMENT ANALYSIS
DESCRIPTORS
b.lDENTIFIERS/OPEN ENDED TERMS
c. COS AT I Field/Group
Thermal pollution
Fish
Sublethal effects
Ashpit effluents
Wisconsin power plant
study
Fish habitats
06/F
07/B
07/C
18. DISTRIBUTION STATEMENT
Release to Public
19. SECURITY CLASS (This Report)
unclassified
21. NO. OF PAGES
121
20. SECURITY CLASS (Thispage)
unclassified
22. PRICE
EPA Form 2220-1 (Rev. 4-77)
PREVIOUS EDITION IS OBSOLETE
•ft U.S. GOVERNMENT PRINTING OFFICE: 1980--657-165/0076
109
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