xvEPA
United States
Environmental Protection
Agency
ilh MN 51
EPA-600 3-80-081
August 1 980
Research and Development
Responses of
Stream
Invertebrates to an
Ashpit Effluent
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 foster technology transfer and a maximum interface in related fields.
The nme 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-600/3-80-081
August 1980
RESPONSES OF STREAM INVERTEBRATES TO AN ASHPIT EFFLUENT
Wisconsin Power Plant Impact Study
by
John J. Magnuson
Anne M. Forbes
Dorothy M. Harrell
Judy D. Schwarzmeier
Institute for Environmental Studies
University of Wisconsin-Madison
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
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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 or commercial products constitute
endorsement or recommendation for use.
ii
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FOREWORD
The U.S. Environmental Protection Agency (EPA) was designed to
coordinate our country's efforts toward protecting and improving the
environment. This extremely complex task requires continuous research in a
multitude of scientific and technical areas. Such research is necessary to
monitor changes in the environment, to discover relationships within that
environment, to determine health standards, and to eliminate potentially
hazardous effects.
One project, which 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. This interdisciplinary study,
centered mainly around the Columbia Generating Station near Portage, Wis.,
involves investigators and experiments from many academic departments at the
University of Wisconsin and is being carried out by the Environmental
Monitoring and Data Acquisition Group of the Institute for Environmental
Studies at the University of Wisconsin-Madison. Several utilities and State
agencies 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.
During the next year reports from this study will be published as a
series within the EPA Ecological Research Series. These reports will
include topics related to chemical constituents, chemical transport
mechanisms, biological effects, social and economic effects, and integration
and synthesis.
Since Columbia I began operating the ashpit drain has become an
unsuitable habitat for aquatic invertebrates. Upstreara-downstream
differences in invertebrate communities in Rocky Run Creek were observed
when ash effluent concentration was high. The effects of Columbia I were
undetectable from natural variation in the Wisconsin River. The major
effect of Columbia I on aquatic invertebrates is hypothesized to be
continued habitat alteration and, in particular, reduced substrate quality
and avoidance of unpreferred habitat.
Norbert A. Jaworski
Director
Environmental Research Laboratory
Duluth, Minnesota
iii
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ABSTRACT
Fly ash from the 527-MW coal-fired Columbia Generating Station Unit I
(Columbia Co., Wisconsin) is discharged as a slurry into an adjacent
ashpit. Water from the ashpit is pumped to a ditch that joins the ashpit
drain and Rocky Run Creek before they reach the Wisconsin River. Habitat
alterations have been noted as relatively minor changes in water quality
parameters (e.g., alkalinity, hardness, pH, and turbidity), as increased
amounts of some dissolved trace elements (Cr, Ba, Al, Cd, and Cu), and as
the precipitation of trace elements (Al, Ba, and Cr) into a floe that coats
the stream bottoms.
The ashpit drain became an unsuitable habitat for aquatic invertebrates
after Columbia I began operating. Upstream-downstream differences in
invertebrate communities in Rocky Run Creek were observed when ash effluent
concentration was high. The effects of Columbia I were undetectable from
natural variation in the Wisconsin River. The collection of data in only 1
preoperational year severely limited the analysis of generating station
impact.
Crayfish caged downstream from the ash effluent survived at the same
rate as those caged at upstream control sites, but they contained higher
levels of metals (Cr, Ba, Zn, Se, and Fe) and had lower metabolic rates.
Crayfish fed food containing Cr in the laboratory accumulated less than 3%
of the amount ingested. However, Cr in food may be an important factor
affecting invertebrate populations at the Columbia I site because its
concentration on particulates was high.
Survival of winter-generation Asellus r>acovitzai was similar when
exposed to control and ash-effluent water and to control food and food
exposed to ash effluent. R>or late-winter condition of the isopods
precluded detection of any sublethal effects. However, young-of-the-year
ins tars of Ganrnarus pseudolirnnaeus were more sensitive to the ash effleunt
than were adults.
The conductivity of the effluent increased in January 1977 when sodium
bicarbonate was first used to increase the efficiency of the electrostatic
precipitators. Since then conductivity measurements have indicated effluent
concentration at distances downstream from the generating station.
Thresholds for field and laboratory responses to the effluent, as measured
by conductivity, were estimated at 800 to 1,459 ymhos/cm and averaged about
IjlOOymhos. The annual record of conductivity was used to observe how
often the threshold was exceeded.
iv
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Rocky Run Creek is still a suitable habitat for many aquatic
invertebrates, but evidence of sublethal stresses and habitat avoidance
exists. The major effect of Columbia I on aquatic invertebrates is
hypothesized to be continued habitat alteration and, in particular, reduced
substrate quality and avoidance of unpreferred habitat. The susceptibility
of early life stages of crustaceans to the ash effluent may also be
important. Acute toxicity to adult forms is unimportant.
This report was prepared with the cooperation of faculty and graduate
students in the Laboratory of Limnology at the University of Wisconsin-
Madison.
Most of the funding for the research reported here was provided by the
U.S. Environmental Protection Agency (U.S. EPA). Funds also were granted by
the University of Wisconsin-Madison, Wisconsin Power and Light Company,
Madison Gas and Electric Company, the Wisconsin Public Service Corporation,
and the Wisconsin Public Service Commission. This report is submitted
toward fulfillment 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 U.S. EPA. The
report covers the period July 1, 1975 to July 1, 1978 and work was completed
as of April 10, 1979.
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CONTENTS
Foreword ill
Abstract iv
Figures viii
Tables x±
1. Introduction to the Generating Station Site 1
Purposes of this study 1
The generating station site 2
Generating station operation and adjacent
habitats 2
The aquatic sampling sites adjacent to the
generating station 8
2. Conclusions 14
3. Effects on Community Structure of Macroinvertebrates.......... 20
Introduction 20
Materials and Methods 21
Results 26
Discussion. 45
4. Effects on Individual Organisms 51
Introduction 51
Exposure of crayfish to ash effluent and
chromium-contaminated food 52
References 105
Appendices
A. Review of Literature on Entrainment from Cooling
Lake Intake Structures 114
B. Review of Literature on Acid Rain 119
C. Review of Literature on Alternative Disposal of
Fly Ash 125
D. Literature Review: The Dynamics and Effects of
Chromium and Other Metals in Organisms... 133
E. Wet Weight-Dry Weight Relationship 147
F. Accuracy of University of Wisconsin Nuclear Reactor Data 149
vii
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FIGURES
Number Page
1 Location of invertebrate sampling stations in streams near the
Columbia Generating Station 3
2 Conductivity (pmhos/cm) gradient in streams adjacent to the
Columbia Generating Station in September 1977 5
3 Groundwater flows before and after construction of the Columbia
cooling lake (Stephenson and Andrews 1976) 7
4 Summary of the effects of the ash effluent in field and
laboratory experiments 16
5 Annual conductivity (]amhos/cm) of water upstream (sampling stations
Al and SI) and downstream (sampling stations A2, A3, and A4) of
the ash effluent in the ashpit drain and upstream (sampling
stations Rl and R2) and downstream (sampling stations R3 and
R4) in Rocky Run Creek in 1977-78 17
6 ft>lar ordination of invertebrate samples from basket-type
artificial substrates in the ashpit drain (sampling station A3)
using numeric data, relative-abundance data, and presence-
absence data 27
7 Measures of community diversity in invertebrate samples from
basket-type artificial substrates in the ashpit drain
(sampling station A3) 28
^'
8 Seasonal abundance of dominant invertebrate taxa (numeric data)
in basket-type artificial substrates in the ashpit drain
(sampling station A3) 29
9 Measures of community diversity in invertebrate samples from
basket-type artificial substrates in Rocky Run Creek
(sampling station R5) near the mouth of the
Wisconsin River 33
10 Balar ordination of invertebrate samples from basket-type
artificial substrates in Rocky Run Creek (sampling station R5)
near the mouth of the Wisconsin River 34
viii
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11 Seasonal abundance of dominant invertebrate taxa (numeric data)
in basket-type artificial substrates at the downstream station
in Rocky Run Creek (sampling station R5) 35
12 Seasonal abundance of dominant invertebrate taxa (numeric data)
in basket-type artificial substrates at the upstream station
in Rocky Run Creek (sampling station Rl) 36
13 Number of invertebrate taxa (S) and number of individuals (N)
colonizing modified Bendy samplers upstream and downstream
from the ash effluent in June and September 1977 .............. 38
14 Polar ordination of invertebrate samples from modified Bendy
substrates in Rocky Run Creek upstream (sampling station R2)
and downstream (sampling station R3 and R4) from the ash
effluent in September 1977 40
15 Polar ordination of invertebrate samples from modified Bendy
substrates in Rocky Run Creek (sampling station R2, R3, and R4)
and the sedge meadow flow (station SI) in September 1977 40
16 Measures of community diversity in invertebrate samples from
basket-type artificial substrates at sites upstream (Wl) and
downstream (W2) from the Columbia Generating Station 42
17 Polar ordination of invertebrate samples from basket-type
artificial substrates at sites upstream (Wl) and downstream (W2)
of the Columbia Generating Station 44
18 Seasonal abundance of dominant invertebrate taxa (numeric data)
in basket-type artificial substrates at the upstream station
in the Wisconsin River (Wl) from May (M) through October (0) of
1974 and 1975 and the downstream station in the Wisconsin River
(W2) from May (M) through October of 1974, 1975, and 1976 46
19 Water levels in the Wisconsin River at the Columbia Generating
Station site during 1974-76 and times when basket-type
artificial substrate samples were placed and emptied 50
20 Top view schematic of modified "Trophy" No. 20737 minnow trap used
for caging crayfish at sites in ash basin drainage system 54
21 Schematic of the 465-ml glass jars used as respirometers
(Klinger 1978) 55
22 Concentrations of five metals in the tissues of crayfish caged at
four sites 65
23 The relationship between chromium concentration and duration of
soaking for leaf discs soaked in 1.0 ppm chromium 75
IX
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24 Mean chromium concentration over time in chromium-51 labeled
crayfish 80
25 Relationship between whole-body chromium concentration and total
chromium ingested by chromiura-51 labeled crayfish 81
26 Median number of actions performed by the three groups of
crayfish on four dates during the experiment 82
27 Survival of Asellus racovitsai exposed to leaf litter and water
from locations upstream (sampling station Al) and downstream
(station A4) of the ash effluent 91
28 Survival of Asellue racovitzai exposed to filtered (sampling
station E) and unfiltered (station C) ashpit drain water 93
29 Percent survival of young and adult Gamarus peeudolirmeaue
exposed to the ash effluent for 96 h 96
30 Itercentage of nymphs collected at monthly (1974 and 1975) or
twice-monthly (1976) intervals from the Wisconsin River and
Rocky Run Creek 99
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TABLES
Number Page
1 Summary of fliysical and Chemical Measurements of the Wisconsin
River at the Upstream (Wl) and Downstream (W2) Stations 8
2 Summary of Hiysical and Chemical Measurements Upstream and
Downstream of the Ash Effluent in Rocky Run Creek 10
3 Summary of Hiysical and Chemical Measurements Upstream and
Downstream of the Ash Effluent in the Ashpit Drain and
Rocky Run Creek on 1 and 9 September 1977 12
4 Concentrations (ppm) of Selected Trace Elements in
Dissolved and Suspended Particulate Fractions of the
Ashpit Drainage System • 13
5 Estimated Thresholds for Biological Responses to Ash Effluent
(from Figure 4, a-g) 18
6 Monthly Sampling Schedule for Basket-Type Artificial
Substrates .......................................... 22
7 Sampling Schedule for Modified Dendy Samplers and Number of
Samplers Placed Upstream and Downstream of the Ash Effluent
for 1 Week Colonizations 23
8 Statistical Differences Among the Four Sets of Six Monthly
Samples (1974-1977) in Rocky Run Creek Downstream from
the Ash Effluent (R5) 30
9 Statistical Differences Among Three Sets of Six Post-Operational
Samples (1976-1977) in Rocky Run Creek Downstream from
the Ash Effluent (R5) as Compared to the Corresponding
Set of Six Control or Pre-Operation Samples (1974) 32
10 Significance of Differences in the Numbers of Organisms in Rocky
Run Creek, Above and Below the Ashpit Drain. ...«•.. 39
11 Statistical Differences Among the Three Sets of Six Monthly
Samples (1974-1976) from the Downstream Wisconsin River
Station (W2) 41
XI
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12 Monthly Conductivity at Downstream Rocky Run Greek Station
R5 in 1977
13 Results of Macroinvertebrate Community Study
14 List of Metals Analyzed in Leaf and Crayfish Tissue Samples
for the Crayfish Caging and Chromium Ingestion Experiments ..... 59
15 Chemical and Hiysical Parameters of Site Water During Crayfish
Caging ................. ... ....... ....... ....................... 60
16 Chemical Parameters of Water Used for Measurement of
Metabolic Rates .......... . ..... ..... ........................... 61
17 Survival, Length of Exposure to Ash Effluent, and Mortalities
Among Crayfish Caged at Field Sites ............................ 62
18 Weight -Independent Metabolic Rates (K = mg 02 Consumed/H/l.Og
Sized Crayfish, Wet Weight) for Crayfish Caged at Four Sites... 63
19 Differences in Metabolic Rates Among Crayfish Caged at
Treatment and Control Sites.... ........ .. ............. . ....... 63
20 Significance Tests for Differences in Tissue Metal
Concentrations in Crayfish Caged at Four Sites ................. 66
21 Concentrations of Metals in Sugar Maple Leaves Soaked at Five
Sites in the Ash Basin Drainage Systems. ........ ....... ......... 67
22 Mean Metal Concentrations in Tissues of Frogs, Caged Crayfish,
and Laboratory Crayfish ............... . ................. . ...... 69
23 Lethal Effects of Chromium Ingestion ............................. 83
24 Metal Concentration in Tissues of Crayfish Fed Chromium in
the laboratory ................................................. 84
25 Metal Concentrations in Leaf Discs Fed to Control (C) and
Chromium-Fed (Cr) Crayfish ..................................... 85
26 Preliminary Exposure of Aeellue racovitaa-i to Mixtures of
Ashpit Drain (A3) and Control (Al) Water ....................... 90
27 Trace Element Concentrations (ppm ± 1 S.D. ) in Leaves Prepared
for Feeding Asellus March 1978 ........... . ...... . ..... . ........ 92
28 Summary of Water in Physical and Chemical Measurements of
Water in the Aeetlus Laboratory Experiment from 23 December
1977 to 17 March 1978 .......................................... 94
29 Comparison of life Histories of Heptageniidae from Wisconsin ..... 102
xii
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B-l Summary of pH Effects of Aquatic Organisms..... • 121
D-l Summary of Work Done to Determine Concentrations of Chromium
that are Lethal to Organisms 138
D-2 Summary of Work Done to Determine Sublethal Effects of
Chromium on Organisms 140
2
E-l Regression Equations, Coefficients of Determination (r ), and
Sample Sizes for the Relationships Between Wet Weight and
Dry Weight of Crayfish Used in Experiment . 148
F-l Comparison of Metal Concentrations in Standards Analyzed by
the University of Wisconsin Nuclear Reactor with their
Reported Actual Concentrations of the Standards.. 150
xiii
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ACKNOWLEDGMENT
We are grateful to Stephanie Brouwer for her efforts in editing this
report. Steven Horn and Cheryle Hughes prepared many of the figures.
Countless individuals have helped in the laboratory and in the field. In
particular we recognize Michael Talbot, Samuel Sharr, Walter Gauthier, and
Katharine Webster. The staffs at the Laboratory of Limnology and at the
Institute for Environmental Studies at the University of Wisconsin-Madison
deserve special thanks for their time and patience. Dr. Riilip Helmke and
his staff in the Soil Science Department of the University of Wisconsin-
Madison provided valuable help for trace-element analysis.
The investigators responsible for different aspects of this report
under the direction of Dr. John J. Magnuson are: Anne M. Forbes (community
structure of macroinvertebrates, exposures of isopods and amphipods to ash
effluent, compilation of final report); Dorothy M. Harrell (exposure of
crayfish to ash effluent and chromium-contaminated food, received M.S.
1978); Judy D. Schwarzmeier (community structure of macroinvertebrates and
variability in life histories of Heptageniidae, received M.S. 1976). The
literature reviews were prepared by Anne M. Forbes, Walter A. Gauthier,
Dorothy M. Harrell, and Frank Rahel.
xiv
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SECTION 1
INTRODUCTION
PURPOSES OF THIS STUDY
The original objective of this project was to study the impact of the
Columbia Electric Generating Station on aquatic macroinvertebrates in the
Wisconsin River, Rocky Run Creek, and the ashpit drain. Broad spatial and
temporal changes in invertebrate communities were observed using
multivariate analyses. Samples were collected from artificial substrates at
points upstream and downstream from the generating station 1 yr before
(1974) and 3 yr after (1975, 1976, and 1977) operation began. This work is
documented in Section 3 of this paper, "Effects on Community Structure on
Macroinvertebrates."
The changes in the physical layout and chemical inputs at the Columbia
site have altered the aquatic habitats. Of special interest to this project
were the effects of habitat alterations on aquatic invertebrates living
downstream from the ash effluent discharge. Negative effects of the
effluent could be considered in light of acute toxicity, sublethal toxicity,
avoidance of less favorable habitat, or a combination of the above. Changes
in the trace element content of organisms living in the ashpit drain were
measured by Helmke et al.
Midway through the study, several more specific problems were focused
on: (1) Completing the determination of trace-element levels in aquatic
organisms upstream and downstream from the ash effluent (Trace Elements
Subproject) and gathering data on the biological significance of the changes
observed (Aquatic Invertebrates Subproject); (2) evaluating the effects of
the ashpit effluent as a whole on local invertebrate fauna using a
combination of laboratory and field experiments (Aquatic Invertebrates
Subproject) (supporting data on the effluent contents came from the Aquatic
Chemistry and Trace Elements subprojects); and (3) continuing to monitor
aquatic invertebrate communities in Rocky Run Creek upstream and downstream
of the generating station.
Five studies were initiated to satisfy these objectives: (1) Exposure
of crayfish in cages upstream and downstream from the ash effluent; (2)
follow-up measurements of metabolic rates and trace element body burdens for
crayfish caged in the field; (3) a laboratory feeding study of the uptake
and effects of chromium on crayfish; (4) laboratory exposures of amphipods
and isopods to the ash effluent for data on survival, growth, and
reproductive success; (5) an examination of spatial and temporal variability
in life histories of mayfly nymphs collected in the study area. The
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materials, methods, results, and discussion of these five studies are
presented in Section 4, "Effects on Individual Organisms."
Overall conclusions for this paper are given in Section 2, and the
bibliography is presented in Section 5. The appendices contain literature
reviews: Appendix A, entrainment from cooling lake intake; Appendix B, acid
rain; Appendix C, alternative disposal of fly ash; Appendix D, effects of
chromium and other heavy metals. Appendices E and F contain the supporting
data analysis from the crayfish experiments.
THE GENERATING STATION SITE
The Columbia Generating Station Unit I is a 527-MW coal-fired electric
generating station located on the eastern floodplain of the Wisconsin River,
5 km south of Portage, Wisconsin (Figure 1). The generating station building
is 73 m (240 ft) high with a 150-m (500-ft) boiler chimney equipped with two
electrostatic precipitators. Construction of Columbia I began in 1971;
operation began in April 1975. Columbia II, a second unit of similar size,
with a 195-m (650-ft) stack, cooling towers, and sulfur-removal scrubbers,
began operating in spring 1978.
The 2,726-acre site of these dual generating stations covers a range of
plant and animal communities, including aquatic, wetland, and forested
areas. The installation has permanently altered 1,100 acres which includes
a 500-acre cooling lake, 70-acre ash basin, coal-handling facilities, roads,
and various other structures. The cooling lake, designed to recycle the
thermal effluent from the generating station, was built on 500 acres of
native wetlands. Water from the Wisconsin River was used to fill the
cooling lake and is still pumped almost continuously into the lake to make
up for evaporation and leakage losses.
GENERATING STATION OPERATION AND ADJACENT AQUATIC HABITATS
Fly Ash
Columbia I burns about 5,000 tons/day of low-sulfur pulverized coal
from Colstrip, Montana, with a typical ash content of 7 to 8%. The high
energy electrostatic precipitators installed to reduce particulate emissions
collect approximately 98% of this "fly ash" residue and discharge it as a
slurry with cooling pond water into the ashpit adjacent to the plant. Water
entering the ashpit flows through a series of lagoons where the ash
particles settle out. The water is then pumped to the ashpit drain and
eventually combines with the water of Rocky Run Creek (Figure 1).
The chemistry of the ashpit has been studied (Andren et al. 1977). The
metal oxides composing the major reactive portions of the ash result in a
water pH of 10 to 11. Since Wisconsin water quality standards prohibit the
release of water > pH 8, sulfuric acid is added before the ash effluent is
discharged. Adding acid causes elements such as Ba, Al, and Cr to
precipitate into a floe that coats the bottom of the ashpit drain and is
carried with the current into Rocky Run Creek and the Wisconsin River.
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W2
old bridge abutment
Figure 1. Location of invertebrate sampling stations in streams near the
Columbia Generating Station.
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Beginning in January 1977, sodium bicarbonate was routinely added to
the pulverized coal to increase efficiency of the electrostatic
precipitators. This increased conductivity in the ashpit drain and Rocky
Run Creek (Figure 2). Conductivity became a useful tool for measuring ash
effluent concentration downstream from the generating station. Habitat
alteration and water chemistry changes due to the ash effluent are discussed
in the Rocky Run Creek and Ashpit Drain site descriptions (p. 8-12).
Fly ash and bottom ash produced during the operation of Columbia I are
pumped into the ashpit. When Columbia II is operating, only bottom ash is
added to the total volume of ash; all fly ash from Columbia II must be
disposed of dry. The ashpit structure was altered in preparation for the
handling of dry ash. The ashpit will continue to receive demineralizer and
blow-down waste. How the effluent to the ashpit drain will change is not
known, but it is possible that the volume of effluent will decrease and its
concentration will increase.
Intake Water from the Wisconsin River
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 by impingement on intake screens or
abrasion in pumps, pipes, and condensers.
3. The toxic effects of biocides used to reduce the fouling
of pipe systems by microorganisms.
4. The effects of thermal shock during condenser passage.
Only the removal aspect of cooling water intake is relevant to the
Columbia site; mechanical, toxic, and thermal aspects of entrainment do not
apply because the entrained water is not directly returned to the river. The
total river flow removed at Columbia presently averages 0.3%, with a maximum
of 1.08%.
A 1-yr study of egg and larval fish entrainment and of juvenile and
adult fish impingement at the Columbia site (Swanson Environmental, Inc.
1977) reported insignificant numbers of fish losses. As long as the
Columbia intake continues to remove a small percentage of the river flow, no
measureable effects of entrainment on the river system are expected. An
exception might occur if an organism with patchy distribution becomes
concentrated near the intake and a significant portion of one year-class
(i.e., walleye larvae) becomes entrained. Aside from removing organisms
from the Wisconsin River, the usual entrainraent effects (mechanical, toxic,
and thermal) do not occur at the Columbia station.
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federating
Station
Figure 2. Conductivity (ymhos/cm) gradient in streams adjacent to the
Columbia Generating Station in September 1977.
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Leakage from the Cooling Lake and Ash Basin
The cooling lake was built by constructing dikes on a sedge meadow, a
wetland plant community that builds a deep, peaty soil and is maintained by
groundwater discharge. The cooling lake created a 2.75-m (9-ft) hydrostatic
head above the remaining wetlands (Figure 3). Seepage through the bottom of
the cooling lake has altered the wetland habitat adjacent to the west dike
and has helped to dilute the ash effluent as it flows to Rocky Run Creek.
Before the generating station was built, groundwater from the adjacent
uplands flowed at about 1 ft^/sec to the sedge meadow (Stephenson and
Andrews 1976). Now 1 ft^/sec seeps into the ashpit drain and mostly into
the sedge meadow on the west.
There is evidence that the wetland west of the cooling lake is being
altered by a combination of higher and more stable water levels, increased
surface water flow and substrate erosion, warmer water temperatures, and
perhaps dissolved components (Bedford 1977). Emergent aquatic species and
annuals have replaced the previously dominant sedge meadow communities. An
equilibrium state has not been reached and effects on vegetation are
expected to spread (Bedford 1977). The sedge meadow habitat being replaced
has been described as an ideal spawning habitat for northern pike (Priegel
and Krohn 1975, McCarraher and Thomas 1972).
Leakage from the ashpit was substantial after it was filled, but has
continued only along the west dike where it was not sealed by ash deposits
(Stephenson and Andrews 1976). Well-water samples taken in the fall of 1976
indicated the ashpit had some impact on local groundwater chemistry (Andren
et al. 1977).
Potential for Acid Rain Damage
Although the pH of rainfall in the Columbia Generating Station vicinity
has not been measured, 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 140
mg/liter CaC03 and conductivities of 180 to 280 Vmhos/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 located west and south of the plant.
3. The present pH of the Wisconsin River and Rocky Run Creek (7.6 to 8.2)
is well within the recommended safe range of pH 6.5 to 9.0 for natural
waters and has not changed noticeably since the the plant began
operation in 1975.
The effect of added sulfur emissions when Columbia II begins operation
should be considered in the future. The contributons, if any, of Columbia
plant emissions to acid rainfall over distant waters—such as northern
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River
UJ
Cooling Lake
River
500
1000
METERS
1500
2000
Figure 3. 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 (from Stephenson and Andrews 1976).
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Wisconsin lakes, some of which are poorly buffered and more subject to
acidification—will also need to be considered.
THE AQUATIC SAMPLING SITES ADJACENT TO THE GENERATING STATION
The sampling sites selected for this study have been used for field
sampling of invertebrate communities (Section 3), for field and laboratory
studies on individual species (Section 4), and for associated monitoring of
physical and chemical parameters.
The Wisconsin River
The Wisconsin River flows along the western boundary of the Columbia
Generating Station (Figure 1) approximately 5 km south of Portage,
Wisconsin. Throughout the study area, the Wisconsin River has a sandy
bottom, about 25 to 50 m wide (at low flow) with frequent sandbars and with
scattered submerged logs and fallen trees. The river basin is characterized
by extensive seasonal flooding. Water discharge is partly controlled by a
series of dams. Major vegetation types surrounding the study area are
floodplain forest and sedge meadow. Floodplain consists of silver maple
(Acer sacaharinum'), cottonwood (Populus deltoides), willow (Salix nigrd) and
a ground layer of grasses and shrubs. Land use is primarily agricultural
and recreational.
Two sampling sites were located on the Wisconsin River about 0.5 km
upstream (Wl) and 3.0 km downstream (W2) from the intake of the generating
station (Figure 1). The two sites were chemically similar, but current and
water depth at the upstream site were about twice as great as at the
downstream site (Table 1).
TABLE 1. SUMMARY OF PHYSICAL AND CHEMICAL MEASUREMENTS OF THE WISCONSIN
RIVER AT THE UPSTREAM (Wl) AND DOWNSTREAM (W2) SAMPLING SITES*
Measurement
Depth (cm)
Current (cm/sec)
Temperature (°C)
Dissolved oxygen (ppm)
Conductivity ( mhos at 25°C)
PH
Total alkalinity (ppm)
Total hardness (ppm)
18 Aug.
Wl
110.0
67.0
24.0
8.6
184
8.2
85.0
136.0
1976
W2
56.0
38.0
21.9
8.8
211
8.3
91.0
128.0
12 Oct.
Wl
115.0
i
14.7
11.4
225
8.2
101.0
86.0
1976
W2
57.0
—
13.0
11.2
239
8.2
122.0
108.0
* Measurements were made downstream 4 to 5 h earlier than upstream.
8
-------
Rocky Run Creek
Rocky Run Creek originates in a marsh lake and flows through about 20
km of agricultural lands before reaching the generating station. The creek
is spring fed at several points. The substrate is silt with a high organic
content. Aquatic vegetation includes water stargrass (Heteranthewz dubia),
pondweed (Potamogeton spp.), and coontail (Cer>atophy11im sp.)
The two most widely separated Rocky Run Creek sites—one upstream from
the effluent near Co. Hwy. JV (Rl) and one downstream from the effluent near
the mouth of the creek (R5)—are separated by a large slough where the
ashpit drain discharges into Rocky Run Creek (Figure 1). A main creek
channel connects the two sites, but numerous backwaters and small channels
are present in the adjacent sedge meadow and flood-plain forest. Flood
waters from the Wisconsin River inundate the region in the spring, but water
levels gradually drop until only the main creek channel is visible in
summer.
Two sampling sites (Rl and R2) in Rocky Run Creek are located upstream
from the ash effluent and three sites (R3, R4, R5) are located downstream
(Figure 1). The water upstream from the ash effluent is usually higher in
alkalinity, hardness, and pH than water at the downstream sites. Table 2
compares stations Rl and R5 and Table 3 presents data for sites R2, R3, and
R4. Current speeds and dissolved oxygen are reduced at downstream site R5,
but not at the other downstream sites. Conductivity is high near the ash
effluent entry and decreases only slightly as it flows to the Wisconsin
River (Figure 2). These high conductivities can not be traced into the
river itself. Precipitated elements from the ashpit drain form a floe which
coats the creek bottom and results in slightly elevated turbidity.
Ashpit Drain
The creek that receives the ashpit effluent originates in wetlands that
are part of a mint farm east of the power plant. The creek originally
crossed the area now occupied by the cooling lake and entered Duck Creek
just above the north knoll (Figure 1). During construction of the Columbia
facility, this creek was diverted to run parallel to the east dike of the
cooling lake and then turn westward to join the backwaters of Rocky Run
Creek. The substrate of the drainage ditch was originally silt with a high
organic content. After generating station operation began, some areas
(particularly those near site A3) gradually became more sandy as the silt
washed away. The artificial drainage ditch banks are steep and grassy for
about 0.75 km beyond the railroad tracks. At the end of this diking, sedge
meadow and flood-plain forest form the creek banks.
Sedge meadow or drained sedge meadow still border the mint farm creek
for several kilometers upstream from site Al. The substrate consists of
silt with a high organic content primarily from decomposing sedge.
One sampling station (Al) is located in the mint drain upstream from
the ash effluent (Figure 1) and five sites (A2, A3, A4, A5, and A6) follow
the effluent gradient to its entry into Rocky Run Creek. Conductivity is
-------
TABLE 2. SUMMARY OF PHYSICAL AND CHEMICAL MEASUREMENTS UPSTREAM AND
DOWNSTREAM OF THE ASH EFFLUENT IN ROCKY RUN CREEK. AVERAGES OF MONTHLY
SAMPLES (MAY THROUGH OCTOBER), + 1 S.D.; (RANGE): AND NUMBER OF SAMPLES
Measurement
Temperature (°C)
Current speed
(cm/speed)
Dissolved oxygen
(mg/liter)
Conductivity
(ymhos/cm)
at 25°C
Alkalinity
phenol (ppm)
Upstream
1976
19.7±7.4
(7.3-25.9)
6
25.716.2
(15.4-32.1)
5
12.1±0.8
(10.8-13.3)
6
472±14
(454-482)
9.4±12.6
(0.0-28.5)
7
(Rl)
1977
17.5±5.9
(8.8-24.0)
6
23.7±2.3
(20.8-26.1)
4
11.6±0.7
(11.0-12.8)
5
485±14
(475-512)
4.9±9.4
(0.0-23.4)
6
Downstream
1976
20.4±6.4
(11.1-26.5)
6
9.311.3
(7.7-11.2)
6
437130
(386-477)
3
(R5)
1977
17.216.9
(9.1-27.0)
6
< 5
8.611.7
(5.7-10.0)
5
8041280
(528-1188)
6
Total (ppm)
260.3121.4 250.1115.4 225.21H9.6 195.1125.5
(222.3-283.0) (235.0-278.8) (200.0-256.5) (158.0-228.2)
767 6
Hardness (ppm) 269.512.9 270.914.0 244.112.6 209.5121.0
(266.0-273.0) (264.8-275.2) (240.6-247.0) (175.6-234.8)
464 6
PH
79510.20 7.8710.18 7.6410.21 7.6110.26
(7.75-8.20) (7.65-8.05) (7.25-7.06) (7.36-7.90)
666 6
Turbidity (JTU)
4.541.9
(2.2-7.0)
5.
7.0+3.9
(5.0-14.0)
5
10
-------
outfall (Table 3). Current speeds in the ashpit drain (A2 through A6) are
faster than in the mint drain (Al) and alkalinity, hardness, and usually pH
are lower. Temperature is usually several degrees higher in the ashpit
drain, probably because of leakage from the cooling lake. Some dissolved
trace-element concentrations are high in the ashpit, but precipitate after
the pH is lowered and before the effluent is discharged to surrounding
waters (Table 4). The precipitated elements form the floe that slightly
elevates turbidity and coats the bottom of the ashpit drain (Table 3).
Although levels of these dissolved trace elements are lower in the ashpit
drain than in the ashpit, they are higher than the levels in the mint drain
or Wisconsin River. They sometimes exceed the estimated no-effect
concentrations which are based on single elements in laboratory bioassays
(Water Quality Criteria 1973). There are also some fly-ash particles and
perhaps some organic byproducts of coal combustion in the ash effluent.
11
-------
TABLE 3. SUMMARY OF PHYSICAL AND CHEMICAL MEASUREMENTS UPSTREAM AND
DOWNSTREAM FROM THE ASH EFFLUENT IN THE ASHPIT DRAIN AND ROCKY RUN CREEK
ON SEPT. 1 AND 9 1977. AVERAGE, + 1 S.D.; (RANGE); AND NUMBER OF SAMPLES
N)
Measurement
Temperature (°C)
Current speed
(cm/sec)
Dissolves
oxygen ,
(mg/liter)f
Conductivity
(pmhos/cm)
Alkalinity
phenol (ppm)
Total (ppm)
Hardness (ppm)
PH
Turbidity (JTU)
Al
18.0*0.0
(18.0)
2
+
1
7.6
1
454±20
(439-468)
2
0.0
1
250.2
1
244.4
1
7.45
3.8
1
A2
21.5±0.7
(21.0-22.0)
2
11.4
1
— — .
2882±68
(2834-2930)
2
0.0
1
59.2
1
120.0
1
7.25
27.0
1
A3
21.1±0.1
(21.0-21.2)
2
21.4
1
8.6
1
2515±11
(2507-2523)
2
0.0
1
93.8
1
130.0
1
7.35
27.0
1
Site
A6
22.8±1.1
(22.0-23.5)
2
11.1
1
2376*136
(2279-2472)
2
0.0
1
103.8
1
145.2
1
7.45
21.0
1
SI
22.3±1.1
(21.0-23.6)
2
6.1
1
420±12
(429-412)
2
0.0
1
220.9
1
197.6
1
7.25
21.0
1
R2
20.9±1.3
(20.0-21.8)
2
21.6
1
10.9
1
480±11
(472-488)
2
4.8
1
282.8
1
272.4
1
7.95
6.2
1
R3
22.5±0.7
(22.0-23.0)
2
29.7
1
1119±147
(954-1238)
3
0.0
1
193.8
1
204.8
1
7.65
18.0
1
R4
22.2±0.2
(22.0-22.3)
2
18.2
1
10.3
_— —
1
1108±67
(1060-1155)
2
0.0
1
207.2
1
210.0
1
7.65
17.0
1
+Strong wind was reversing the current on this date; typical values range from 7 to 9 cm/sec.
*0xygen meter malfunctioned. Dissolved oxygen values reported are for 7 Oct. 1977. Temperatures were 9.2, 9.0, 8.0, and
8.0°C, respectively.
-------
TABLE 4. CONCENTRATIONS (PIM) OF SELECTED TRACE ELEMENTS IN DISSOLVED
AND SUSPENDED PARTICIPATE FRACTIONS OF THE ASHPIT DRAINAGE SYSTEM*
Wisconsin
River
Ashpit
discharge**
Mint drain
(Al)
Ashpit drain
(A2) (A4)
Dissolved particulates
Cr
Bat
Al
Cd
Cu
<0.001-0.001
<0.050
0.022-0.093
0.0001
0.014-0.077
0.730
0.1-11.4
<0.001
<0.075
0.003-0.165
0.035-0.065
0.380
0.03-4.0
0.006-0.028
0.045-0.476
0.0021-0.0031 0.0001-0.0002 0.0024-0.0029 0.0001-0.0012
<0.001-0.002 0.004-0.045 <0.0003-0.002
Suspended particulates
Cr
Ba
Alt
134
17,240
107
0.004-0.043
740
1,350
0.002-0.024
1,294
1,225
Dissolved analyses were from monthly samples in November 1976 through April 1977
by Andren et al.(1977). Suspended particulates were measured in the fall of 1975
^by Helmke et al. (1976a).
Before addition of sulfuric acid.
tHelmke et al. (1976b).
fAluminum is an important component of the precipitate in the ashpit drain (Andren
et al. 1977), but its concentration cannot be measured by neutron activation
analysis.
-------
SECTION 2
CONCLUSIONS
Fly ash from the 527-MW coal-fired Columbia Generating Station Unit I
(Columbia Co., Wis.) is discharged as a slurry into an adjacent ashpit.
Water from the ashpit is pumped to a ditch that joins two streams—the
ashpit drain and Rocky Run Creek—before reaching the Wisconsin River.
Relatively minor changes in water quality parameters (e.g., alkalinity,
hardness, pH, and turbididty), increased amounts of some dissolved trace
elements (Cr, Ba, Al, Cd, and Cu), and the precipitation of trace elements
(Al, Ba, and Cr) into a floe that coats the bottom of the streams have
caused habitat alterations.
Effects of the fly-ash effluent from Columbia I on aquatic invertebrate
communities decreased as distance from the generating station increased. The
ash effluent concentration changed on a seasonal basis depending on the
volume of water pumped from the ashpit and on the amount of dilution from
the mint drain creek, sedge meadow flow, Rocky Run Creek, and groundwater
discharge. The conductivity of the effluent increased in January 1977 when
sodium bicarbonate was first used to increase the efficiency of the
electrostatic precipitators. Since then, conductivity measurements have
been used to indicate effluent concentration at distances downstream from
the generating station.
After Columbia I began operating in 1975, the ashpit drain—the creek
that directly receives the ash effluent—became an unsuitable habitat for
aquatic invertebrates. There was a 3-month delay before community changes
were observed in 1975; some invertebrate taxa thrived late in the fall.
However, upstream and downstream comparisons of invertebrate communities in
1977 revealed a lack of organisms colonizing artificial substrates
downstream.
Community differences were also observed in Rocky Run Creek—0.5 km
downstream from the ash effluent entry—when compared to upstream samples,
but only when conductivity was over 1,000 ymhos/cm. Upstream-downstream
differences were not detected in Rocky Run Creek when the conductivity was
near 800 ymhos/cm. Invertebrates drifting from the upstream station to
downstream stations were suddenly exposed to the ash effluent and apparently
did not colonize artificial substrates when the effluent concentration was
above a threshold level. Invertebrates appeared unaffected 1 km downstream
in Rocky Run Creek where pre-operational data were compared to post-
operational data.
Effects of the operation of Columbia I were undetectable in the
Wisconsin River from 1974 to 1977. Natural variation in seasonal cycles of
14
-------
invertebrate communities in the river were documented; these data will be
useful for long-term monitoring of the river.
It was possible to examine the reasons for the differences observed in
the field by controlling exposures of individual populations of crustaceans
to the ash effluent. Crayfish caged downstream from the ash effluent
survived at the same rate as those caged at upstream control sites, but they
contained higher levels of five metals (chromium, barium, zinc, selenium,
and iron) in their body tissues and had lower metabolic rates. The lowered
metabolic rates of exposed crayfish were influenced by one or more of the
following: Reduced quantity or quality of food; increased metal
concentrations in tissues; or possibly a combination of water quality
parameters affected by coal-combustion byproducts.
Concentrations of chromium in potential food sources for invertebrates
at the Columbia site increased as much as four-fold in leaf litter and nine-
fold in suspended particulates downstream from the ash effluent. Laboratory
crayfish exposed to chromium in their food accumulated less than 3% of the
amount ingested. However, chromium in food may be an important factor
affecting invertebrate populations at the Columbia site because of its high
concentration of particulate sources.
Survival of winter-generation Aeellus rviaovitzai. was similar for
exposure to control and ash-effluent water and to control and ash effluent-
exposed food. Boor late winter condition of the isopods precluded detection
of any sublethal effects. However, young-of-the-year Gammarus
peeudolimnaeus were more sensitive to the ash effluent than were adults of
the species.
Results of these studies can be considered in relation to effluent
concentration at any one time and place (Figure 4). Thresholds for field
and laboratory responses to the ash effluent were estimated by averaging no-
effect and lowest-effect conductivities (Table 5). The threshold for
effects fell between 800 and 1,459 ymhos/cm with an average of about 1,100
ymhos/dm. The conductivity gradient downstream of the generating station
can be used to predict the extent of the effects.
Conductivity measurements were above the threshold of effects in the
ashpit drain before it enters Rocky Run Creek during all of 1977-78 (Figure
5). Exceptions occurred when the generating station was not operating for
short periods in the spring and fall of 1977. Conductivity in Rocky Run
Creek was lower than in the ashpit drain, usually near the 1,100 ymhos/cm
threshold and higher on one occasion. Responses of invertebrate communities
were more subtle in the creek than in the ashpit drain and were more
difficult to assess. Although Rocky Run Creek is still a suitable habitat
for many aquatic invertebrates, evidence of sublethal stress and habitat
avoidance exists.
It is hypothesized that the major effect of the operation of Columbia I
on aquatic invertebrates is through habitat alteration and in particular,
through reduced substrate quality and avoidance of unpreferred habitat.
Susceptibility of early life stages to the ash effluent may also be
15
-------
1000 2000
Conductivity
» S«pt. c Jun« d S«pt.
z —
? &600-
a i
•* DC
X "D
8 S
-20 | |«0.
Q)' ®
3 S.
Q- 1
3) 2
.108 |200-
-r
V
T
z _
r s
S E
8 1
^ to
-20 1 IJ400-
3 2
1 I
? ^
-1" o |200-
2. £
o
6
Z
(n.2-7) |
I
I
I
1
1
1
1
1
I '
s. 1
" r^S"s^l 1
\ iH~ *-
V 1
1 1 1 1 1 1 1
1000 2000 3000 500 1000 500 1000
Conductivity Conductivity Conductivity
—I—
500
1
1000
ASELLUS(n*70)
GAMMARUS(n
a 1°°
5
g
(n.10)
Conductivity
1000 2000
Conductivity
500 1000
Conductivity
Figure 4. Summary of the effects of the ash effluent in field and laboratory experiments. Ash-
effluent concentration is expressed by conductivity (ymhos/cm) of the water. Modified Bendy
samplers were placed upstream and downstream from the ash effluent in the ashpit drain (a
and b) and Rocky Run Creek (c and d). Crayfish (e) were exposed in the field for 62 days
and their metabolic rates were measured in the laboratory (K* mg02 h~l g~^). Asellus and
Gcormapus (f) and northern pike eggs (g) were exposed to ash effluent in the laboratory.
-------
3000 r-
2500 -
ISOOi-
ioa
500
° R4
• R1-2
M
1977
M
O
N
D J
1978
Figure 5. Annual conductivity (umhos/cm) of water (a) upstream (sampling
stations Al and SI) and downstream (sampling stations A2, A3, and
A4) from the ash effluent in the ashpit drain and (b) upstream
(sampling stations Rl and R2) and downstream (sampling stations
R3 and R4) in Rocky Run Creek in 1977-78.
17
-------
TABLE 5. ESTIMATED THRESHOLDS FOR BIOLOGICAL RESPONSES
TO ASH EFFLUENT (FROM FIGURE 4, a through g)*
Type of experiment From Figure 4 Threshold conductivity
( mhos/cm)
Field a
b
c**
d
Average
Laboratory e
ft
R
1,042
1,448
—
800
1,096
1,148
1,459
825
Average 1,144
* Thresholds were estimated by averaging the conductivity of the control
water and the conductivity of the most dilute ash effluent water to elicit
a response.
**No response observed.
t'Gammarus only; no response for Aeellue.
important. Acute toxicity to adult forms of crustaceans is unimportant.
Guthrie et al. (1974) studied the effects of coal-ash effluent on a
stream that emptied into a swamp, then into smaller streams, and finally
into the Savannah River in Georgia. Mechanisms that would return effluent
water to acceptable standards before the water entered the river were
emphasized. Important mechanisms were the settling of particulates and the
recycling of chemical elements by aquatic food webs. This research
demonstrated the importance of entire food webs for pollutant removal and
suggested the selective introduction of resistant consumers to increase the
cycling efficiency of the biotic system (Guthrie and Cherry 1976).
The drainage system for ash effluent at the Columbia site also flows
through small streams and wetland habitat before reaching the Wisconsin
River. In contrast to the Georgia study, the Columbia wetland has been
valued as a habitat for spawning game fish (Magnuson et al. 1980) and
resident and migratory birds (Willard et al. 1977). Inputs from the coal-ash
effluent and changes in the characteristics of the wetland vegetation
adjacent to the cooling lake (Bedford 1977) threaten habitat quality for
wildlife. Our concern with long-term effects of the generating station is
18
-------
focused equally on habitat loss and on the quality of water entering the
Wisconsin River. Wetland habitats available for spawning fish populations
in this section of the Wisconsin River were inventoried (Magnuson et al.
1980) and it was determined that the generating station site provides at
least 30% of the available habitat.
In summary, it was concluded that the 3.6-km long ashpit drain has
become an unsuitable habitat for aquatic invertebrates. The habitat quality
of a localized area of Rocky Run Creek at least 0.5 km downstream of the
ashpit drain entry has also been reduced. These localized effects will
increase as more effluent is discharged and effects will be most evident in
low-water years when effluent dilution is minimal.
19
-------
SECTION 3
EFFECTS ON COMMUNITY STRUCTURE OF MACROINVERTEBRATES
INTRODUCTION
The purpose of this chapter is to document the impact of the Columbia
Generating Station on aquatic invertebrates in the Wisconsin River, Rocky
Run Creek, and the ashpit drain. Of particular interest is the community of
invertebrates inhabiting the two streams that receive the ash effluent—the
ashpit drain and Rocky Run Creek (Figure 1). Sample sites were selected
upstream (Al, Rl, and R2) and downstream (A2 through A6, and R3 through R5)
from the ash effluent in both streams to observe effects of a dilutional
gradient. Effects of the ash effluent were not measureable in the Wisconsin
River (Wl and W2), but seasonal cycles of macroinvertebrates as baseline
data for long-term change were documented.
Aquatic invertebrates have often been used to indicate environmental
quality and change (Wilhm 1975). Analytical techniques provide indices of
community diversity from data on numbers and kinds of species in a series of
samples. Traditional techniques produce diversity indices for one sample at
a time. Multivariate techniques present similarities and differences
between many samples simultaneously; therefore, assemblages of many species
can be compared in space and time.
Ordination as a multivariate tool was used to compare samples because
it gives a graphic representation of complex relationships. Patterns in the
data are not readily discernable with the more classical analytical methods.
Ordination has been used in macroinvertebrate studies to observe community
change along physical gradients such as salinity, temperature, or substrate
type (Hughes and Thomas 1971; Erman 1973; Hocutt 1975). It also has been
used to demonstrate the ordering of a series of macroinvertebrate stations
along a gradient of many pollutional disturbances (Beckett 1978). In this
study ordination was used to examine:
1. Similarities between different stations on a single date
2. Changes in the communities at different stations on a seasonal basis
3. Changes in the seasonal cycle of organisms from year to year.
For comparison, more traditional methods of representing community
diversity were used: the Shannon-Weaver index (Shannon and Weaver 1949),
evenness (Pielow 1966), equitability (Lloyd and Gherlardi 1964), and the
number of taxa and individuals.
20
-------
MATERIALS AND METHODS
The Invertebrate community was sampled using artificial substrates so
that colinization at different stream sites could be assessed without the
complication of variable substrate type. The resulting samples are quan-
titative but are not measures of actual standing crops. Two types of
artificial substrate samplers—basket-type with limestone and a modified
Dendy-type—were used to sample the invertebrate community. Both upstream-
downstream and before-after operation sampling designs were used (Tables 6
and 7).
Basket-Type Artificial Substrates
Five sampling sites were selected in 1974 for the pre- and post-opera-
tional study (Figure 1). Three were downstream from the generating
station. The first (A3) was in the ashpit drain about 1.8 km downstream
from the confluence of the ash effluent with the mint farm creek, the second
(R5) was near the mouth of Rocky Run Creek, and the third and farthest down-
stream site (W2) was in the Wisconsin River about 0.5 km from the mouth of
Rocky Run Creek. Two sampling sites were located upstream from the
generating station, one (Wl) on the Wisconsin River about 0.4 km upstream
from the intake channel and the other (Rl) on Rocky Run Creek near Co. Hwy.
JV.
Monthly samples were taken after spring floods until freeze-up in 1974
and 1975. Sampling continued at the downstream Wisconsin River site in 1976
and at both Rocky Run Creek stations in 1976 and 1977 (Table 6).
Organisms were collected from basket-type artificial substrate samplers
(Mason et al. 1970) which consisted of 20- x 29-cm chicken barbeque baskets
(or similar wire replicas) filled with 4.5 kg of limestone gravel (average
diameter of 7.6 cm) and suspended from overhanging branches 5 to 10 cm above
the substrate. Three samplers were placed at each station. At
approximately 1-month intervals, the organisms were removed by shaking the
samplers about 12 times inside an aquatic D-frame net (1-mm mesh). In the
Wisconsin River samplers, nets of Hydropsychidae larvae often held rocks
together; when this occurred, the basket was shaken until the rocks were
loose.
All samples were preserved in 70% alcohol in the field. Insects and
crustaceans were sorted, identified, and counted in the laboratory. Samples
were divided into four subsamples before sorting. If more than 2.5 h was
needed to sort the first 25% of a sample, the sample was subsampled.
Downstream Rocky Run Creek samples (R5) were never subsampled.
Identification was usually to the generic or sometimes family level;
specific identifications were made where possible. The smallest instars of
the Hydropsychidae and Corixidae families could not be identified to genus;
separate categories were created for them. Pupae were keyed to family.
21
-------
TABLE 6. MONTHLY SAMPLING SCHEDULE FOR BASKET-TYPE ARTIFICIAL
SUBSTRATES*
Location Month Pay
1974 1975 1976* 1977
Wisconsin River
upstream (Wl) and
downstream (W2)
May 10 28 24
June 10 26 22
July 14 23 21
August 15 21 17
September 13 19 15
October 11 17 12
Rocky Run Creek
upstream (Rl) and
downstream (R5)
May 10 28 24 31
June 10 26 22 28
July 14 23 21 27
August 15 21 17 25
September 13 19 15 23
October 11 17 12 21
Ashpit drain (A3)
May — 28
June 14 26
July 15 23
August 13 21
September 10 19
October 12 17
November 19
*Dates shown are midpoints between sampler placement and
removal. 1974 was the pre-operational year. Samplers were
placed after the peak of the spring flooding each year.
**Downstream only for the Wisconsin River.
22
-------
Modified Dendy Samplers
The invertebrate community was sampled upstream and downstream of the
ash effluent in June and September of 1977 (Table 7). Artificial substrates
were modified from the multiple-plate Dendy sampler (Hester and Dendy
1962). Each of these modified Dendy samplers was made from a single Tuffy
brand mesh ball held between two 8- x 8-cm masonite plates by an eye bolt
and wing nut. Samplers were held in place with floats and weights. Because
modified Dendy samplers could be placed anywhere in the stream, it was
possible to randomly sample the creeks at sites closer to the ash
effluent. Placement of the basket-type artificial substrates had required
overhanging branches. The smaller size of the Dendy samplers made it
possible to process more replicates. Three samplers were randomly placed at
sites Al, A2, A3, R2, and R3 in June; four samplers at Al, A2, A3, A5, A6,
and SI in September; and eight samplers at R2, R3, and R4 in September
(Figure 1). After a 1-week exposure, the samplers were removed and placed
in freezer containers with 70% alcohol. The 1-week exposure time was
selected to minimize the buildup of detritus around the sampler. Longer
exposure times were not practical because of this buildup. At the
laboratory, samples were concentrated through a 0.07-mo net and organisms
were identified and counted.
TABLE 7. SAMPLING SCHEDULE FOR MODIFIED DENDY SAMPLERS AND
NUMBER OF SAMPLERS PLACED UPSTREAM AND DOWNSTREAM OF
THE ASH EFFLUENT FOR 1-WEEK COLONIZATIONS
Date removed
Location Station 6 June 19779Sept. 1977
Mint drain
Ashpit drain
Al
A2
A3
A4
3
3
3
0
4
4
4
4
Sedge meadow flow SI 0 4
Rocky Run Creek
Upstream R2 3 8
Downstream R3 3 8
R4 0 8
Although eight samplers were placed at stations R2, R3, and R4 in
September, several samplers were missing when collections were made. Seven
samplers remained at R2, two at R3, and four at R4. For some analyses,
23
-------
samples from the two downstream sites, R3 and R4, were combined to provide a
total of six downstream replicates. Samplers were also missing at stations
A5 and A6 on the same date. These losses were due to sudden, extremely high
ashpit drain flow or were removed by boaters.
Supporting Riysical-Chemical Data
Water temperatures were taken whenever samplers were placed or
removed. Dissolved oxygen (YSI Model 54A oxygen meter or winkler; American
Riblic Jfealth Association 1971), conductivity (YSI Model 33), pH (Fisher
Accuinet Model 150 or Hach flienol Red kit), alkalinity and hardness (American
Riblic Health Association 1971) and current speed (Drogue, Ocean Equipment
Model 451, or Neyrpic Midget) were measured irregularly in 1974 and 1975 and
whenever samplers were placed or removed in 1976 and 1977. Turbidity was
measured in 1977 (Hach Model 2100A). All conductivity readings were
adjusted to 25°C.
Methods of Analysis
Polar Ordination—
Many dominant taxa and numerous seasonal or spatial changes make
comparisons among stations and years difficult to describe. Ordination
techniques simplify these comparisions by accounting for all species and
stations simultaneously. Ordination plots can illustrate temporal
differences as trajectories connecting samples in time (Bartell et al.
1977). They can also show spatial changes in species composition along
ecological gradients or reveal spatial or temporal patterns by clumping
similar samples. Samples that represent important changes or that are
notably different are often evident as endpoints of the axes or as isolated
points. In interpreting ordinations, it is useful to remember that the
first axis usually illustrates the greatest (and possibly most important)
differences between samples. Successive axes usually account for less
variation.
Analyses were performed using three data transformations (numeric,
relative abundance, and presence-absence) to assess the influence of taxa
with varying numerical importance. Numeric (raw) data were averaged from
the three basket-type samplers in each collection. Numeric data used in
ordinations emphasized dominant taxa the most, as it also did ^n relative
abundance form, but the latter put samples on an equal numerical basis. The
presence-absence form weighted the presence or absence of each species
equally, regardless of numerical abundance. Three axes were constructed for
each ordination, but the third axis seldom revealed useful information.
Polar ordination was chosen for this study because it involves less
ecological distortion and because the results tend to be more ecologically
interpretable than results from other mathematically more sophisticated
methods of ordination (Gauch and Whittaker 1972, Beals 1973). In polar
rdination, distances are calculated between all pairs of samples based on
their compositional similarity using the Bray-Curtis Dissimilarity
24
-------
Coefficient. Axis endpoints were selected with the variance/regression
method (Beals et al. unpublished).
Other community measures—
Other indices of species structure in communities are based on
calculations from single samples. Some of these methods were applied to
this data:
Number of taxa
Number of individuals
The Shannon function (Shannon and Weaver 1949, Wilhm 1970), based on
s
information theory: H = -£ pilog oi; where s = total number of taxa
1=1
in a sample and pi is the proportion of individuals in the i[th]
taxon. This index was recommended by the U.S. Environmental
Protection Agency (U.S. EPA) for purposes of establishing uniformity
between different studies (Weber 1973). H is influenced by the number
of taxa (s) and by the evenness (e) with which individuals are
apportioned between the species; it is relatively independent of total
sample size (N) (Odum 1971).
"iT
Evenness, measured by the function e = (Pielow 1966) and by
log s g,
the recommended U.S. EPA method, the equitability ratio E = — ; where
S
s1 is the hypothetical number of species based on MacArthur's model
(Lloyd and Gherlardi 1964).
Distances are calculated between all pairs of samples based on their
compositional similarity using the Bray-Curtis dissimilarity coefficienta:
n
E X., - X
d(Xi, Xj) = '
** 1 /•»?•• Tr • \ K—-L
n
E xik + E xik
k=l 1K k=l J
where d = distance between sample pairs and n = number of species.
25
-------
Statistical Analyses—
Statistical tests were used where appropriate to compare invertebrate
distributions. The position of samples on ordination axes and the other
measures of community diversity listed above were compared between years
using a Friedman two-way analysis of variance by ranks (Siegel 1956).
Samples collected upstream and downstream from the ash effluent in 1977 were
analyzed with the Mann-Whitney U test (Siegel 1956).
RESULTS
Pre- and tost-operational Sampling of the Ashpit Drain and Rocky Run Creek
Samples of invertebrate communities from the basket-type artificial
substrates enabled us to compare a pre-operational year (1974) with post-
operational years in the ashpit drain (1975) and Rocky Run Creek (1975,
1976, and 1977). In the ashpit drain, considerable changes in the community
colonizing the substrates were observed in the post-operational year,
1975. In Rocky Run Creek, community variations during 1975, 1976, and 1977
were subtle and impossible to relate to generating station operation.
Ashpit Drain—
When numbers of organisms were important with numeric .data in an
ordination and when dominant taxa were emphasized with relative abundance
data, the first axis distinctly separated August, September, and October of
the post-operational year from all other months (Figures 6a and b).
Reductions in number of taxa (s) and individuals (N) occurred in these 3
months (Figure 7). Only the September 1975 sample separated because a
number of taxa were absent; this is seen on the first axis of the presence-
absence ordination (Figure 6c). A reduction in number of taxa was also
apparent in July of the pre-operational year. This sample separated on the
second axis of the presence-absence ordination. However, the reduction was
temporary; larger numbers of taxa appeared during the rest of the year
(Figure 7).
Examples of taxa that were common in 1974 and early 1975, but were
absent or greatly reduced during the last 3 months sampled in 1975 (August,
September, and October) are Stenacvan •Lnterpunatatwnt Hydropeyahe sp.,
Cheumatopsyche sp., juvenile Hydropsychidae and Chironomidae (Figure 8).
Other taxa such as Coenagrionidae, Lepidoptera, and Simuliidae appeared in
the fall of 1974, but did not reappear in similar numbers in the fall of
1975. New taxa did not replace these losses or reductions. A few taxa were
unchanged in number (for example, Hyalella azeteca, sp., and Asellue
raeov-Ltzai).
The Shannon index H for pre- and post-operational ashpit drain samples
revealed lower diversity in August, September, and October of the first
post-operational year (Figure 7). In this case, H gave a useful
representation of large community changes in response to the ash effluent.
H could be used in combination with the numbers of taxa, numbers of
individuals, and the numeric data to draw conclusions similar to those
26
-------
NUMERIC
RELATIVE ABUNDANCE
PRESENCE/ABSENCE
8 - »-«)975
Figure 6. Polar ordination of invertebrate samples from basket-type arti-
ficial substrates in the ash pit drain (sampling station A3) using
numeric data, relative abundance data, and presence-absence data.
Monthly samples were collected from June (J) through November (N)
in the pre-operational year (1974) and from May (M) through
October (0) in the post-operational year (1975).
27
-------
H
4.0
3.0
2.0
1.0
o'
1974
1975
1.0 r
e o.s
1.0 r
0.5
0
j i
N
40 r
20
0
2000
1500
1000
500
M J J A S 0 N
MONTH
Figure 7. Measures of community diversity in invertebrate samples from
basket-type artificial substrates in the_ash pit drain (sampling
station A3). The Shannon-Weaver index (H), eveness (e), equita-
bility (E) , number of species (S), and number of individuals (N)
were calculated for monthly samples from June (J) through November
(N) in the pre-operational year (1974) and from May (M) through
October (0) in the post-operational year (1975).
28
-------
1974
J N
1975
M 0
KS
VD
Hyalello azteca
Coenagrionidae
Hydropsychidae (pupae)
Chironomidae (pupae)
Caen is
Gammarus pseudolimnaeus
Stenocron
Lepidoptera
(Nymph. /Neo.)
Simuliidae
Hydropsyche
Cheumatopsyche
Hydropsychidae
(juvenile)
Asellus racovitzai
Chironomidae
TOTAL
1974 1975
J N M 0
T400
Figure 8. Seasonal abundance of dominant invertebrate taxa (numeric data) in basket-type artificial
substrates in the ash pit drain (sampling station A3). Data were collected from June (J)
through November (N) in the pre-operational year (1974) and from May (M) through October (0)
in the post-operational year (1975).
-------
resulting from the use of ordinations and numeric data. Evenness (e) and
equitability (E) were not particularly useful for showing pollution
responses; the degree of apportionment for individuals among the taxa was
similar in 1974 and 1975 even though H was reduced in 1975 (Figure 7). This
occurred because the number of taxa (s) was correspondingly reduced.
Statistical analyses comparing the 2 years were impossible because of
small numbers of paired monthly samples (n = 5). However, the differences
observed formed the basis for continued monitoring of the ash effluent (See
"Upstream and Downstream Sampling of the Ashpit Drain and Rocky Run Creek").
Rocky Run Creek—
Sets of six monthly samples (May through October) were taken from the
basket-type substrate samplers during 1 pre-operational (1974) and 3 post-
operational (1975-77) years at the downstream Rocky Run Creek station
(R5). Statistical analyses demonstrated that the 4 years were not different
from each other in the number of taxa (S) present, in the Shannon-Weaver
diversity index (H), in evenness (e) or equitability (E), or in axis
positions from relative abundance or presence-absence ordinations (Table 8).
TABLE 8. STATISTICAL DIFFERENCES AMONG THE FOUR SETS OF SIX
MONTHLY SAMPLES (1974-77) IN ROCKY RUN CREEK
DOWNSTREAM FROM THE ASH EFFLUENT (R5)"*"
--
Parameter Sum of ranks rL
S
N
H
e
E
Numeric ordination
Axis 1
Axis 2
Relative abundance
Axis 1
Axis 2
1974
15
19
15
14
17
8.5
18
ordination
13
20
1975
12
7
17
19
18
21
6
12
14
1976
20.5
19
14
12
11
14.5
15
15
16
1977
12.5
15
14
15
14
16
21
20
10
4.55
9.60
0.60
2.60
3.00
9.90
12.05
3.80
5.85
ns
*
ns
ns
ns
*
**
ns
ns
Presence/absence ordination
Axis 1
Axis 2
10
20.5
16
16
13
12
21
11.5
6.60
6.15
ns
ns
* p < 0.05.
**p < 0.01.
'^Friedman two-way analysis of variance by ranks,
^Ranked from lowest to highest values.
30
-------
Significant differences were observed in comparing the total number of
organisms or in comparing the axes from numeric ordination. From the sum of
ranks for total number of organisms, it is clear that the numbers were
lowest in 1975. The sum of ranks for axes from numeric ordination was
highest for 1975 on the first axis and lowest for 1975 on the second axis.
One might argue that generating station operation affected numbers of
invertebrates, but not taxonomic proportions in 1975, and that recovery
occurred before the 1976 samples were collected. However, there is data for
only a single pre-operational year, and there is no evidence that the drop
in numbers is beyond the range of natural variation. We believe that this
amount of fluctuation would occur in the absence of the generating station.
Since differences between years were based only on numerical data and
similarities were reflected in all other community measures tested, it is
concluded that generating station operation had no observable effect at this
sampling station.
A second approach to the analysis of Rocky Run Creek samples treated
the pre-operational year as a control by subtracting data for each month
from the respective months in the post-operational years. This was done for
the ordinations by directly measuring distances between pre- and post-
operational samples for each month from graphs of the first and second
axes. The result was one ranking procedure for each ordination.
All parameters for the post-operatonal years varied randomly from the
pre-operational year except for the Shannon-Weaver index (H) and the
relative abundance ordination (Table 9). The index H was most similar to
control in 1975, as shown by the low sum of ranks. The seasonal changes in
H were nearly identical in 1974 and 1975 (Figure 9). H sometimes varied
from 1974 in 1976 and 1977, but the extent or pattern of the fluctuations
did not indicate that the operation of the generating station was a cause.
Differences from the control year on the relative abundance ordination
axes were pronounced in 1977 (Table 9); the sum of ranks is larger than in
1975 or 1976. At this point, it is useful to look at the ordination axes
before attempting an explanation.
For each of the three ordinations (numeric, relative abundance, and
presence-absence), trajectories connecting the six monthly samples for each
year were separated for viewing (Figure 10). The first axis in the
numerical and relative abundance ordinations spread the samples on a
seasonal basis, with fall samples occurring near the origin. Fall peaks of
Pelooorie femoratus and Hyalella azteca populations (Figure 11) contributed
to this. The seasonal cycles in different years appear quite similar. The
midsummer samples in 1977 (July and August) show the greatest differences
from other years; these samples are at the endpoints of the second axis for
numerical data (Figure lOd) and are compressed nearer the origin (as
compared with midsummer 1974-76 samples) for relative abundance data (Figure
lOh). Changes in taxonomic composition in the three post-operational years,
as compared to the control year, were statistically significant in the
relative abundance ordination. The differences, however, are subtle and
cannot be attributed to generating station operation. The most obvious
31
-------
TABLE 9. STATISTICAL DIFFERENCES AMONG THREE SETS OF SIX
POST-OPERATIONAL SAMPLES (1976-1977) IN ROCKY RUN CREEK
DOWNSTREAM FROM THE ASH EFFLUENT (R5) AS COMPARED TO THE
CORRESPONDING SET OF SIX CONTROL OR PRE-OPERATIONAL SAMPLES (1974)'
± 2
Parameter Sun of ranks t r
1975 1976 1977
s
N
H
e
E
11.5
15
7
15
9
13.5
11
17
9
15
11
10
12
12
12
2.04
3.82
9.94
3.74
4.50
ns
ns
*
ns
ns
Ordinations
Numeric
Relative abundance
Presence/absence
15
8.5
13
10
9.5
8
11
18
15
3.82
10.70
5.52
ns
**
ns
p <0.001
*p < 0.05.
**p < 0.001.
1"Friedman two-way analysis of variance by ranks.
^Ranked from least to greatest difference from 1974.
SBased on distances from plots of Axis 1 vs. Axis 2.
differences in 1977 are midsummer population increases in Asellus
•naaovitsai, Dubiraphia sp., and Caen-is sp., and decreases in Berosus sp.
(larvae) (Figure 11).
Community diversity, as illustrated by H, e, and E, declined in the
fall of each year (Figure 9), a pattern that can be attributed to the
dominance of Peloovis femoratue and Hydlella azteca* Although the diversity
functions demonstrate a certain predictability in community structure
through the sampling season at this location, the ordinations were useful
for: 1) showing when yearly seasonal abundance cycles were similar because
of taxonomic composition and 2) encouraging us to examine the midsummer
samples for subtle changes in community structure in 1977.
Data were also collected at the upstream Rocky Run station (Rl) to
compare the magnitude of year-to-year change between upstream and downstream
stations. These data are being stored and may be used if changes caused by
the generating station are eventually observed downstream. The invertebrate
community at the upstream statipn (Figure 12) was quite different from the
downstream station because of habitat differences and especially because of
faster current upstream.
32
-------
4.0
3.0
H 2.0
1.0
0
1974
1975
1976
1977
1.0 P
6 0.5
1.0 r
C 0.5
N
40 r
20
1000
500
0
M J J A S O
MONTH
Figure 9. Measures of community diversity in invertebrate samples from
basket-type artificial substrates in Rocky Run Creek (sampling
station R.5) near the mouth of the Wisconsin River. The Shannon-
Weaver index (H), evenness (e), equitability (E), number of species
(S), and number of individuals (N) were calculated for each of six
monthly samples from May (M) to October (0) in pre-operational year
(1974) and 3 post-operational years (1975-77).
33
-------
197*
1975
1976
1977
.8-
. 9
X'
Figure 10. Polar ordination of invertebrate samples from basket-type arti-
ficial substrates in Rocky Run Creek (sampling station R5) near
the mouth of the Wisconsin River. Trajectories connect six
monthly samples from May (M) through October (0) for 1 pre-
operational year (1974) and 3 post-operational years (1975-77).
All four trajectories for numeric data (a through d) are based
on one ordination and are separated for clarity. The same is
true for all four trajectories for relative abundance data (e
through h) and for presence/absence data (i through 1).
34
-------
M
1974
1975 1976 1977
M 0 M 0 M 0
Crangonyx
Asellus
racovitzai
Caenis
Stenacron
Dubiraphia
Deronectes
Berosus
(larvae)
Coenagrionidae
Pelocoris
femoratus
Chironomidae
Hyalella
azteca
TOTAL
208
Figure 11. Seasonal abundance of dominant invertebrate taxa (numeric data)
in basket-type artificial substrates at the downstream station
in Rocky Run Creek (sampling station R5). Data were gathered
from May (M) through October (0) in 1 pre-operational year (1974)
and 3 post-operational years (1975-77).
35
-------
1974 1975 1976
MO M M
~
Perlesta placida
Pycnopsyche
Baetis
Ase/lus racovitzai
Stenonema exiguum
Deronectes
Chironomidae
Stenacron
-I
Hyalella azteca
Gammarus pseudolimnaeus
Cheumatopsyche
Hydropsychidae (juv.)
TOTAL
500
Figure 12. Seasonal abundance of dominant invertebrate taxa (numeric data)
in basket-type artificial substrates at the upstream station in
Rocky Run Creek (sampling station Rl). Data were gathered from
May (M) through October (0) in 1974, 1975, and 1976.
36
-------
Upstream and Downstream Sampling of the Ashpit Drain and Rocky Run Creek
The modified Bendy substrates were used to compare invertebrate
communities closer to the ash effluent entry than was possible with basket
samplers. The data were collected in the third post-operational year 1977
which was the first year conductivity could be used to indicate ash effluent
concentration.
Numbers of taxa and individuals per modified Dendy sampler were
extremely low in the ashpit drain, as compared to the upstream mint creek in
June and September 1977 (Figure 13). Conductivity was over 2,000 mhos/cm
in the ashpit drain, indicating high effluent concentration. During the
June sampling, water was being drawn out of the cooling lake into the sedge
meadow, and hence into Rocky Run Creek. This diluted the ash effluent to
less than 800 mhos/cm downstream in Rocky Run Creek and the invertebrate
community colonizing the artifical substrates was not affected. In
September, the dilution from the sedge meadow was significantly reduced and
the conductivity in downstream Rocky Run Creek was greater than 1,000
Umhos/cm. The number of individuals colonizing the substrates downstream as
compared to upstream declined.
Polar ordinations of the September Rocky Run data were used to
interpret the nature of community differences between upstream and
downstream samples. The differences in numbers of individuals in the
upstream and downstream samples was apparent on the first axis of ordination
using numeric data (Figure 14d). The percentage ordination demonstrated a
tendency for upstream and downstream samples to separate on the second axis,
but it can be concluded that the dominant species are similar at both
locations (Figure 14b). An ordinaton of presence/absence data resulted in
the separation of upstream and downstream samples on the first axis (Figure
14c). The second axis separated the two downstream samples occurring
closest to the ash effluent. The series of ordinations indicated that total
numbers of individuals were different downstream from the ash effluent, that
community structure was similar as far as dominant species are concerned,
but that total taxonomic composition differed.
2
We selected the 13 most abundant of 29 taxa from the above samples and
hypothesized that differences in number would be statistically lower
downstream using the Mann-Whitney test on each taxon (p < 0.05). The
hypothesis held true for 10 of the 13 taxa, as well as for the total numbers
of individuals (Table 10).
There are three upstream sources of organisms that can potentially
colonize Rocky Run Creek downstream from the ash effluent: 1) the ashpit
drain, 2) upstream Rocky Run Creek, and 3) the flow from the sedge meadow.
The ashpit drain can be excluded because of the near absence of organisms.
Ordinations of the upstream and downstream Rocky Run Creek samples and the
sedge-meadow samples demonstrate that community structure in the sedge-
Excluded taxa were represented by less than four individuals and were
present in only a few of the samples.
37
-------
June
Sept.
s
N
u mhos/cm
s
N
u mhos/cm
Derating
Station
7(6-8)
22(19-34)
1643
I 11(8-14)
j 34(47-75)
I
L.
1107
15(11-17)
203(114-382)
792
[ 10(8-12)
I 52(23-82)
u JJJJL.
i
(
!
i
i
12(10-13)
144(100-293)
493
15(14-18)
257(93-343)
480 !
0(0-2)
0(0-2) j
L_J25J5_J
Figure 13.
Number of invertebrate taxa (S) and number of individuals (N)
colonizing modified Bendy samplers upstream and downstream from
the ash effluent. Medians (and ranges) are reported for June
and September 1977. Conductivity (umhos/cm) is included as an
indication of effluent concentration.
38
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TABLE 10. SIGNIFICANCE OF DIFFERENCE IN THE NUMBERS OF
ORGANISMS IN ROCKY RUN CREEK, ABOVE AND BELOW THE ASHPIT DRAIN.
Crustacea
Ephemeroptera
Trichoptera
Gamma-rue pseudolirmaeus
Hyalella
Asellus racovitzai n.s,
*
*
Diptera
Total number of individuals
Baetis spp.
Stenaaron •Lnterpunatatwn
Stenonema exiguum
Hydroptilidae
Hydroptilidae pupae
Cheumatopsyehe sp.
Hydropeyohe sp.
Hydropeychidae
(small instars)
Chironomidae
Simuliidae
*
*
n.s.
n.s,
*
*
*Numbers significantly lower below the ashpit drain at p <0.05,
Mann-Whitney U, n^ = 6 (R3 and RA combined), n2 = 7.
meadow flow (relative abundance and presence/absence data) was different
from all Rocky Run Creek samples on the first axis (Figure 15b and 15c).
Numbers of dominant sedge-meadow taxa were intermediate between the two
Rocky Run Creek stations; differences in community structure did not appear
until the second axis of the numeric ordination (Figure 15a). It was
concluded that the sedge-meadow flow makes a relatively small contribution
to the downstream Rocky Run Creek invertebrate community.
Sampling of the Wisconsin River
Seventy-seven different invertebrate taxa, representing 12 orders, were
collected from the Wisconsin River. Dominant taxa were three genera of the
net-spinning caddisflies (Hydropsychidae—Hydropeyche, Cheumatopsyehe, and
Potomya , midges (Chironomidae), and the mayfly genera Baetis, Baetisca,
Caen-is, Isanychia, Stenonema, and Heptagenia. Many of the remaining taxa
were present only sporadically in the basket samplers; 21 taxa were caught
on only one or two dates at the upstream station, and 19 appeared only once
or twice downstream. Other taxa occured at low Ivels throughout most of the
season, never reaching more than 12% of the total community on any one date.
39
-------
NUMERIC
RELATIVE ABUNDANCE
PRESENCE/ABSENCE
.8
.6
CO
X
Q /
§
to
.2
(
|_ c. -8
D .6
.4
•- .2
0
5 Q*?° T
• b.
o -4
O Cfcl
oo •
o -2,
B ^
5 • B D <-
B.ii
c.
D
-
D
•
p o
) •
o _
^ 1 1
0 w .2 .4 .6 .8 0 .2 .4 .6 .8 0 .2 .4
O UPSTREAM-R2 FIRST AXIS
D DOVNSTREAM-R3
• DOWNSTREAM-R4
Figure 14. Polar ordination of invertebrate samples from modified Dendy
substrates in Rocky Run Creek upstream (sampling station R2) and
downstream (sampling stations R3 and R4) from the ash effluent
in September 1977.
NUMERIC
RELATIVE ABUNDANCE
PRESENCE/ABSENCE
8 -
- (9
0
°
J_
• •
•b
0
OUPSTREAM-R2
D DOWNSTREAM-R3
• DCWNSTREAM-R4
• SEDGE-si
n .B
D
.6
- (~} ^ .4
o
L tf B° •
• .2
o , , • ,
I.
• B
o «
" 0 0[§)
D D ••
(5)
-
2 .4 6 .8 .2 .4 6 .8
FIRST AXIS
Figure 15. Polar ordination of invertebrate samples from modified Dendy
substrates in Rocky Run Creek (sampling stations R2, R3, and R4)
and the sedge-meadow flow (station SI) in September 1977.
40
-------
Wisconsin River data were analyzed for two different purposes: 1) to
compare locations downstream from the generating stations for 3 years (one
pre-operational and two post-operational) and 2) to consider the variability
in the seasonal cycles of river invertebrate communities.
There were no significant differences between the 3 years sampled
downstream from the generating station when the numbers of individuals, the
Shannon-Weaver index, or most of the ordination axes were considered (Table
11). Number of taxa was significantly different; the sum of ranks was low
in 1974 as compared to 1975 and 1976. Evenness (e) and equitability (E)
also demonstrated significant differences between years, with a high sum of
ranks in 1974. This would be expected because diversity was not different
but the number of taxa was lower in 1974. In examining the raw data, it was
obvious that while the smaller number of taxa (s) occurred throughout 1974,
the actual numerical difference was small (Figure 16).
TABLE 11. STATISTICAL DIFFERENCES AMONG THE THREE SETS OF SIX MONTHLY
SAMPLES (1974-76) FROM THE DOWNSTREAM WISCONSIN RIVER STATION (W2)f
Parameter
S
N
H
e
E
Sum
1974
7
9
14
17.5
18
*
of ranks
1975
15
12
10
9.5
9
1976
14
15
12
9
9
2
r
7.90
4.50
2.80
9.18
10.62
A
ns
ns
*
A*
Numeric ordination
Axis 1 14 15 7 7.90 *
Axis 2 12 15 9 4.50 ns
Relative abundance ordination
Axis 1 14 10 12 2.80 ns
Axis 2 14 12.5 9.5 3.20 ns
Presence/absence ordination
Axis 1
Axis 2
15
10
11
13
10
13
3.80
2.46
ns
ns
*p < 0.05.
**p < 0.001.
tFriedman two-way analysis of variance by ranks.
^Ranked from lowest to highest values.
41
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4.0 •
3.0-
H 2.0- •
1.0-•
0 -
1.0-
« 0.5-
40-
S 20--
0
Upstream
H 1 (-
1974
1976
H 1 1 1 H
Ql 1 1 1 1 (-
M J
S
MONTH
Downstream
H 1 1 1 1 1
Figure 16. Measures of community diversity in invertebrate samples from
basket-type artificial substrates at sites upstream (Wl) and
downstream (W2) from the Columbia Generating Station. The
Shannon-Weaver Index (H) , evenness (e) , equitability (E) , number
of individuals (N) were calculated for monthly samples from May
(M) through October (0).
42
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Analyses of sample positions on the first axis of the numeric
ordination revealed significant differences between the 3 years (Table
11). None of the other ordination axes demonstrated differences between
years. Community structure, as indicated by relative abundance of taxa or
by presence and absence of taxa, was therefore similar in 1974, 1975, and
1976.
In summary, none of the above analyses indicated effects of generating
station operation. Sampling was limited in only one pre-operational year
(1974). Assuming that the variation observed would have occurred in the
absence of the generating station, the natural variations in seasonal cycles
of river community will now be considered.
All 30 samples (five sets of six monthly samples) were included in each
of the three ordinations (numeric, relative abundance, and presence-
absence). The five time-series trajectories were difficult to view on one
graph, so they have been separated for interpretation (Figure 17). Seasonal
change usually accounted for the greatest variation, as shown by the spread
of each series of dates on the first axis. Seasonal change often continued
to be important on the second axis, resulting in somewhat similar
trajectories in different years (for example, Figure 17h and j) or places
(for example, Figure 17f and h). The seasonal pattern appeared in about
one-half of the plots of the first and second axes as a loop, with the
communities at the end of the sampling season (fall) coming back near those
at the beginning (spring) (for example, Figure 17a, i, and k).
Numeric Data—
Ordinations using numeric data had samples with the greatest number of
organisms nearer the origin of Axis 1 and samples with the smallest number
located at the opposite end (Figure 17 a through e). Numerical differences
between samples were of continued importance on Axis 2 where spring samples
with extremely high numbers of juvenile Hydropsychidae (Figure 18) were
located at the end of the axis (Figure 17b and d).
Upstream and downstream seasonal cycles were similar in 1974 (Figure
17a and c), with seasonal loops- showing similarities between spring and fall
samples. The pattern (Figure 17b and d) was different in 1975 when the
total numbers of organisms varied widely during the sampling season (Figure
18). The pattern downstream in 1976 (Figure 17e) was much like the 1974
pattern (Figure 17 c), but the loop was closer to the origin of Axis 1;
there was a greater total number of organisms (N) throughout 1976 as
compared to 1974 (Figure 18).
Relative Abundance Data—
Transforming data to relativized form placed samples on an equal
numerical basis. Ordinations of upstream and downstream samples were again
similar in 1974 (Figure 17f and h); however, the 1975 samples not only
differed from 1974 but at midsummer they were at opposite ends of the first
axis (Figure 17g and i). The taxa with the greatest number of total
organisms in July and August 1975 were different upstream and downstream
43
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UPSTREAM
DOWNSTREAM
1974
I
1975
I
1974
1975
1976
1 1
c
111
z
a.
M
b.
. or
.8
ui
.8 .8
9' O
C.
e,
\. . . .
X
-*
M
.6
I
Mi ui
ZZ1
UJiu
k.
.6
M
.6
O.
Figure 17. Polar ordination of invertebrate samples from basket-type artificial substrates at sites
upstream (Wl) and downstream (W2) from the Columbia Generating Station. Trajectories
connect six monthly samples from May (M) through October (0) for each location and year.
All five trajectories for numeric data (a through e) are based on one ordination and are
separated for clarity. The same is true for all five trajectories for relative abundance
data (f through j) and for presence-absence data (k through o).
-------
(Figure 18). Downstream, the Chironomidae comprised 70 to 75% of the
samples. Upstream, Chironomidae contributed less than 5% and a combination
of juvenile Hydropsychidae (small instars), Eydropsyche, Baetis, Isonychia,
and Cheumatopsyche accounted for 80 to 90% of the total. The upstream
midsummer community was more similar to the 1974 community by having several
dominant taxa. The downstream 1976 samples (Figure 17j) formed a loop
similar to that of 1974 (Figure 17h), indicating that the taxonomic
succession was similar in both years.
Presence/Absence Data—
When rare taxa had equal weight to abundant taxa in the ordinations,
community differences upstream and downstream became apparent. The large
number of taxa (77), however, made the differences difficult to interpret.
In general, the upstream samples were separated less on Axis 1 than on Axis
2 (Figure 17k and 1) and the reverse occurred for downstream samples (Figure
17m, n, and o). There was a number of taxa, characteristic of areas with
slow current speeds, that were found primarily at the downstream station
where slower current speeds are typical. These included Stenaeron
interpunetatum, Trieorythodes sp. and Gomph-idae. Their presence indicated a
spatial difference in river communities that became apparent in the
ordinations only when the less abundant taxa were given equal importance.
DISCUSSION
Ashpit Drain and Rocky Run Creek
The ashpit drain has become an unsuitable habitat for aquatic
invertebrates since Columbia I began operating. There was a 3-month delay
before community changes were observed in 1975 and some invertebrate taxa
thrived late in the fall. However, upstream and downstream comparisons of
invertebrate communities in 1977 revealed an impressive lack of organisms
colonizing artificial substrates downstream. Conductivity could not be used
as a measure of effluent concentration until 1977. However, the effluent
probably became more concentrated as fly ash buildup increased in the ash
basin.
Community differences were also observed in an upstream-downstream
comparison in Rocky Run Creek when the downstream conductivity was over
1,000 ymhos/cm but not when the conductivity was about 800 ymhos/cm.
Invertebrates drifting from upstream station R2 to downstream stations R3
and R4 would have experienced a sudden exposure to the ash effluent and
apparently did not colonize artificial substrates when the effluent
concentration was above a threshold level. Also, invertebrates appeared
unaffected when compared to pre-operational data on two occasions when the
conductivity near the mouth of Rocky Run Creek was >1,000 ymhos/cm (Table
12). The taxa near the creek mouth were gradually exposed to the 1,000-
Umhos/cm effluent, as shown by conductivity measurements through the
season. Perhaps the taxa present during the low current flows at this
station were more tolerant to this poilutional stress.
45
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UPSTREAM
1974 1975
M 0 M 0
Empididoe
Simuliidae
Boat is
Isonychio
Stenonema
exiguum
Stenonema
terminatum
Potamyia
flava
Hydropsychidoe
pupae
Cheumatopsycht
Hydropsyche
1974 1975
M 0 MO
Hydropsychidoe
(juvenile)
Chironomidoe
TOTAL
r500
1974
DOWNSTREAM
1975
1976
Empididae
Simuliidae
Stenonema
terminalum
Hydropsychidae
pupae
Cheumotopsyche
Hydropsyche
0 M 0 MO
I 1 1 1 1 1 I I ''''
Hydropsychidoe
(juvenile)
Chironomidae
TOTAL
1974 1975 1976
M 0 M O M , , 0
r500
Figure 18. Seasonal abundance of dominant invertebrate taxa (numeric data)
in basket-type artificial substrates at the upstream station in
the Wisconsin River (Wl) from May (M) through October (0) of
1974 and 1975 and the downstream station in the Wisconsin River
(W2) from May (M) through October of 1974, 1975, and 1976.
46
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TABLE 12. MONTHLY CONDUCTIVITY
AT DOWNSTREAM ROCKY RUN CREEK
STATION R5 IN 1977
Month Conductivity
(ymhos/cm)
May 528
June 1,086
July 798
August 1,118
September 530
October 693
It is concluded that the effects of ash effluent decreased in severity
as distance from the generating station increased (Table 13). On the basis
of field observations, it is hypothesized that thresholds for ash effluent
toxicity and/or habitat avoidance exist between 800 and 1,000 ymhos/cm.
This is discussed in more detail along with the results of laboratory
experiments in Section 2 (Conclusions).
Wisconsin River
None of the community parameters tested for the Wisconsin River samples
indicated generating station effects on aquatic invertebrate communities.
The effects of the operation of Columbia I were not as yet measurable and
research was therefore focused on smaller streams closer to the generating
station. Seasonal cycles of invertebrate communities in the river were
documented and the data will be of use for long-term monitoring of the
river. The collection of data in only 1 pre-operational year (1974) in all
streams studied has been and will continue to be a limitation.
Few studies have examined long-term temporal variation in
macroinvertebrate communities. Ward (1975) found very little change over 29
years in a relatively undisturbed mountain stream. Richardson (1928)
documented more drastic community changes over 12 years in midwestern
streams affected by pollution. McConnville (1972) reported consistent
seasonal trends during a 2-year period on the Mississippi River.
Other studies have documented spatial variation among widely separated
stations within both a stream and single riffle (Needham and Usinger
1956). McConville (1972) found few community differences in a 5-mile
section of the Mississippi River. Beckett (1978) used polar ordination to
demonstrate faunal homogeneity in a series of 14 stations on a southwestern
Ohio river system during the June high water period. The same 14 stations,
however, were ordered along a gradient of pollutional disturbances during
low flow in August and September when pollutant concentrations were
47
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TABLE 13. RESULTS OF MACROINVERTEBRATE COMMUNITY STUDY
Location
Distance downstream
from ashpit (km)
Years
sampled
Detectable
differences ?
Ashpit drain
Ashpit drain
Rocky Run Creek
Rocky Run Creek
Wisconsin River
1,3,5
5.5
1974-1975
1977
1977
1974-1975-1976-1977
1974-1975-1976
Yes - before
& after
Yes - upstream
& downstream
Yes - upstream
& downstream
No - before
& after
No - before
& after
No - upstream
& downstream
maximized.
Similar seasonal cycles of aquatic invertebrates were observed in this
study at two stations 3.5 km apart in the Wisconsin River in 1974. The
time-series trajectories created by the first two axes were very similar for
ordinations of numeric and relative abundance data. The 1975 seasonal
pattern was quite different from that of 1974 and the relative abundance
ordination revealed differences between upstream and downstream samples in
midsummer. The 1976 seasonal trajectory was similar to that of 1974. The
presence-absence ordination was more difficult to interpret, but showed
spatial differences within the river by spreading upstream samples on the
second axis and downstream samples on the first axis.
Seasonal change was reflected on the first and/or second axes in the
ordinations regardless of the data transformation, but the comparison of
several transformations was still useful. For example, use of the numeric
data isolated samples with high or low total numbers of organisms.
Ordination of relative abundance data helped to show specific patterns in
seasonal cycles by eliminating the effect of total sample size without
removing the importance of dominant taxa. The presence/absence data made it
possible to assess the importance of less abundant taxa by giving all taxa
equal weight. The use of several transformations can be applied to
pollution studies using ordination. For example, samples with similar
48
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diversity but with different dominant taxa or with similar diversity but
different numbers can be separated. More ecological information is often
revealed than if traditional diversity indices are used alone.
The initial spring placement of our samplers into the Wisconsin River
was governed by the timing of receding floods (Figure 19). Samplers were
placed after the peak spring floods when danger of damage or loss was
reduced. Hence, our year-to-year comparisons of invertebrate community
structure were based around water level rather than temperature,
photoperiod, etc. One might argue that the differences in the invertebrate
community observed in 1975 were due simply to the later onset of the
sampling program. However, the 1976 samplers were placed at about the same
calendar time as in 1975, but the seasonal cycle was more similar to that of
1974. The year with the most unusual water level was 1975; there were
several fall floods. The fall water levels in 1974 were higher than those
of 1976, but mean daily river depths varied little in either year. This
could help explain the small differences in invertebrate communities
upstream between 1974 and 1975, but not the larger differences downstream.
A sand bar formed in front of the downstream station in 1975. The offshore
portion of the bar was exposed by early July and by early August it was
continuous with the shore. The altered current flow caused more siltation
at the downstream station and some taxa typical of this habitat were more
numerous (e.g. Stenaeron, Tr-icorythodes, Gomphidae). The sand bar was
covered with water again in 1976 which would explain the greater similarity
of 1974 and 1976 trajectories from the downstream site. Other sources of
variation in this study include natural influences, man-induced
perturbations, and sampling error, although none of the variations observed
wre attributed to man-made causes. The study was not designed to assess
variability in the river as a whole by randomly selecting many stations, but
rather to look at the predictability of patterns at only two locations. A
study encompassing both points of view would be valuable.
49
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Wisconsin River levels adjacent to
Columbia Electric Generating Station
Jwiuary F*bruwy
StpMmbw Odabw NowrtMr Oaomlwr
Figure 19. Water levels in the Wisconsin River at the Columbia Generating
Station site during 1974-76 and times when basket-type artificial
substrate samples were placed and emptied.
50
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SECTION 4
EFFECTS ON INDIVIDUAL ORGANISMS
INTRODUCTION
This section documents studies initiated in 1976 and 1977 to examine
the effects of the ash effluent on several invertebrate species. These
studies include: 1) exposure of crayfish in cages upstream and downstream
from the influence of the ash effluent; 2) follow-up measurements of
metabolic rates and trace-element body burdens for the crayfish caged in the
field; 3) a laboratory feeding study of the uptake and effects of chromium
on crayfish; 4) laboratory exposures of amphipods and isopods to the ash
effluent for data on survival, growth, and reproductive success; 5) an
examination of spatial and temporal variability in life histories of mayfly
nymphs collected in the study area. The mayfly study documented variability
in seasonal cycles as part of the baseline data rather than as a study of
specific effects of the ash effluent.
For the experiments on the effects of the ash effluent, three
crustaceans that occur on the site and are easily handled in the laboratory
were selected: an isopod, Asellus raeovitsai', an amphipod, Gammarus
pseudolinmaeus; a crayfish, Orconectes propinquus. Asellus is extremely
abundant in the mint drain and Rocky Run Creek and is still present in low
numbers in the ashpit drain. Ganvnarus is common to Rocky Run Creek,
particularly at stations upstream of the old bridge abutment (Rl, R2, R3,
and R4) (Figure 1). Several species of Oreoneetes have been collected at
the site. The crayfish were selected as a representative benthic
detritivore large enough for trace-element analysis of individual tissues,
for holding in field cages with large enough mesh to maintain current flow
and food supply, and for convenient behavioral observations.
The Trace Elements and Aquatic Chemistry subprojects supplied data on
heavy-metal and trace-element concentrations for various components and
locations in the aquatic system (described in detail in Section I). These
data established increased levels of several elements in aquatic
invertebrates and in both soluble and particulate forms in the water
column. Of particular interest were chromium and barium which were higher
in the water column, suspended particulates, and organisms in post-
operational years.
Considering the high concentrations of some elements in the particulate
fraction of the affected waters (see Table 4), we have hypothesized that
food might be an important source of trace-element exposure for benthic
detritivores such as crayfish and isopods. At the same time, our knowledge
of the forms and concentrations of all coal-combust ion byproducts in the ash
51
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effluent has been incomplete. Thus it has been necessary to examine habitat
modifications and organism responses caused by the ash effluent as a whole
using conductivity as an estimate of the effluent concentration at any one
time.
EXPOSURE OF CRAYFISH TO ASH EFFLUENT AND CHROMIUM-CONTAMINATED FOOD
Exposure of Crayfish to Ash Effluent
Introduction—
The ash basin effluent from the Columbia Generating Station contains
elevated concentrations of metals in both soluble and particulate forms
(Helmke et al. 1976a and 1976b, Andren et al. 1977). These increased metal
levels might have adversly affected aquatic organisms inhabiting the
drainage system. Organisms collected from the ashpit drain contained
significantly higher concentrations of barium, chromium, selenium, and
antimony than did organisms from unaffected sites (Schoenfield 1978).
Although many trace metals are essential to organisms in small
concentrations, exposure to high concentrations may interfere with important
physiological processes. When organisms accumulate excess quantities, their
ability to survive or maintain a population may be impaired.
The objectives of this study were: 1) to determine the effects on
crayfish of exposure to a coal-ash effluent containing elevated metal
concentrations and 2) to study the effects of the ingestion of chromium-
contaminated food. Mortalities, metabolic rates, and tissue metal uptake
were determined for crayfish exposed to ash effluent in the drainage
system. Metabolic rate measurement may be a particularly valuable means of
detecting sublethal effects since oxygen consumption can reflect many kinds
of tissue damage or enzyme impairment. Ingestion may be an especially
important mode of uptake for chromium and other metals because metals are in
particulate form in the ashpit drain and are consequently available for
consumption by detritivores such as crayfish. To assess the degree of metal
uptake by detritus, and hence the quality of this food source for crayfish,
leaf material was soaked at effluent-affected sites and analyzed for metal
concentration.
Materials and Methods—
Crayfish collection and holding facilities—The crayfish, Orconectee
propinquue (Girard), used in all experiments were collected from dense
populations in Trout Lake, Vilas County, Wisconsin, with liver-baited minnow
traps or by divers with hand nets. The crayfish inhabits streams affected
by the generating station (Forbes, personal communication) but not in high
enough numbers for intensive experiments. In the laboratory, crayfish were
held in a flow-through system of three 190-liter glass aquaria and were fed
trout pellets approximately twice weekly. Water temperature varied
seasonally from 12 to 26°C; photoperiod was 16 h light:8 h darkness.
Water Chemistry Analysis—On-site measurements were made as follows:
Dissolved oxygen was measured directly with a YSI Model 54A meter;
52
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conductivity and temperature were measured with a YSI Model 33 meter;
current speed was measured with a Midget C.M. Neyrpic current meter.
Conductivities were corrected to 25°C. Water samples were collected with a
Magnuson-Stuntz siphon (Magnuson and Stuntz 1970) and returned to the
laboratory for determination of pH, hardness, alkalinity, and turbidity.
Laboratory analysis of field and laboratory samples was performed as
follows: In the laboratory, pH was measured with a Fisher Accumet Model 150
meter; turbidity was measured with a Hach Model 2100A meter. Alkalinity was
determined using the phenolphthalein and methyl orange indicator methods and
hardness was measured with the EDTA titrimetric method (American Public
Health Association 1976). Laboratory measurements of dissolved oxygen were
from Broenkow and Cline's (1969) and Klinger's (1978) modifications to the
Modified Winkler Method (American Public Health Association 1976).
Statistics—Symbols indicating levels of statistical significance are:
*, P < 0.05, **, P < 0.01, ***, P < 0.001. Unless otherwise indicated,
means are reported with the standard deviation of the sample. Degrees of
freedom are given with the symbol d.f. For analysis of variance, d.f. are
given first for the numerator mean square, then for the denominator mean
square.
Field Exposure—Six male and six female crayfish collected in June and
August 1977 were caged at four sites in and near the ashpit drain (Figure 1)
between 16 Sept. and 17 Nov. 1977. The two treatment sites were in the
ashpit drain (A-4) about 3 km downstream from the ash basin and in Rocky Run
Creek (R-4) immediately downstream of its confluence with the ashpit
drain. Hie two control sites were in the mint farm drain (A-l) just before
it merged with the ashpit drain and in Rocky Run Creek (R-2) just upstream
from its confluence with the ashpit drain. Crayfish were caged in plastic
minnow traps (Figure 20). One male and one female crayfish occupied each
cage, one animal in each compartment. The cages, resting on the substrate
in the middle of each stream, were anchored perpendicular to the flow with
bricks. The soft sediment at each site partially filled the cages. Organic
matter and organisms in the sediment flowed into the cages providing food
for the crayfish. Cages were checked once or twice a week and animals found
missing or dead before 22 Oct. were replaced.
Temperature and conductivity were measured at each site on every
visit. Every other week current speed and dissolved oxygen were measured
and water samples were collected for laboratory determination of pH,
hardness, total and phenolphthalein alkalinity, and turbidity.
The generating station was shut down for maintenance from 3 to 19 Oct.
1977. During this time period, no ash effluent was pumped from the ashpit
into the drainage system. Consequently, crayfish surviving the entire
period were not as completely exposed to the effluent for 16 out of 62 days.
Respirometry—All crayfish were returned to the laboratory on 17
November and placed in closed system respirometers (Figure 21) randomly
assigned to positions in a cold room. The crayfish acclimated for 24 h in
continuously aerated water from their own caging sites. The water
53
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CURRENT FLOW
Brick
Neoprene
stop
16cm
21cm
•43 cm
Figure 20. Top view schematic drawing of the modified "Trophy" No. 20737
minnow trap used for caging crayfish at sites in the ash basin
drainage system. Neoprene stoppers blocked each entrance funnel
and a 30- x 30-cm piece of PVC-coated fiberglass screen divided
each trap into two compartments. Trap mesh size ranged from
4 x 5 to 2 x 3 mm.
54
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Sampling lube
No. 19 hypodermic
needle
No. 12 Neoprene
stopper
Figure 21. Schematic drawing of the 465-ml glass jars used as respirometers
(Klinger 1978). The sampling tube consisted of 15 cm of 5-mm I.D.
glass tubing and 30 cm of 6.4-mm I.D. "Tygon" tubing. The
stopper and sampling tube created an air-tight seal in the jar.
The hypodermic needle inserted through the stopper allowed air
to enter the top of the jar as water was siphoned from the bottom
during sampling.
55
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temperature was 5°C and photoperiod was 10.5 h light: 13 h darkness,
approximating November field conditions.
After crayfish were acclimated, airstones were removed and the water in
each jar was replaced gently with aerated, filtered (0.45 Millipore
filter) water from the field site where that crayfish originally had been
caged. A water sample was siphoned into a 30-ml reagent bottle, stoppered
immediately, and analyzed for dissolved oxygen. This water, taken from the
respirometer, was replaced with additional water, under the assumption that
the dissolved oxygen concentration was identical. The stopper was inserted
securely and the initial time recorded. About 24 h later, time was recorded
and a second water sample was removed for analysis. Airstones were provided
and the animals were undisturbed for another 24 h. To determine metabolic
differences between crayfish under identical conditions, all crayfish were
placed in filtered tap water; respiration was measured immediately using the
previously mentioned procedures.
Temperatures in 14 respirometers were measured with a mercury
thermometer before and after each experiment. Mean temperature was 4.9 -j-
0.5°C with no difference between treatments. Twelve respirometers with no
crayfish contained the four site waters and tap water and served as controls
to determine dissolved oxygen changes not caused by crayfish. Triplicate
oxygen determinations from seven additional respirometers showed analytical
error no larger than 0.13 rag 02/liter. Samples of filtered site and tap
water were analyzed for conductivity, pH, hardness, and alkalinity.
At the end of the tap water experiment, the crayfish were weighed,
measured (carapace length and maximum width), and frozen in individual
plastic bags. All animals were later dissected and analyzed for metal
concentrations.
A stepwise multiple regression was performed separately on the oxygen
consumption data (M = mg 0^ consumed/h) for experiments in site and tap
water. Caging location and log wet weight were the only variables that
related significantly to oxygen consumption. Sex and interactions between
variables were not important. The regression coefficients, b, of the
equations using site and log weight (b = 0.751 for site water; b = 0.939 for
tap water) were used to find a weight-corrected metabolic rate for each
crayfish in each experiment based on the assumption that M = K^ or logll =
logK + b logW (where W = weight and K and b are constants) (Prosser 1973).
The weight-independent metabolism, K (Prosser 1973), expresses the metabolic
rate of a unit-sized organism (in this case 1 g). Since the relationship
between weight and oxygen consumption is not linear, crayfish of different
sizes will have different metabolic rates (VQ = mg Oo consumed/h/g). Thus,
metabolic rates should not be compared without first correcting for the
effect of weight by determining weight-independent metabolism. This value
can not be obtained by multiplying K by the weight. Instead, the equation M
= KW must be used. Log K was determined for each crayfish and the means
for the groups were compared with analysis of variance and appropriate
56
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a priori tests: [ashpit drain (A-4) against its control, mint drain (A-l);
Rocky Run downstream (R-4) against Rocky Run upstream (R-2); pooled
treatments (A-4 + R-4) against pooled controls (A-l + R-2)].
Calculations were performed using crayfish wet weight. The
relationship between wet weight and dry weight is shown in Appendix E. This
relationship may be used to convert the individual metabolic rate data to a
dry weight basis.
Soaking of Leaves in Ash Effluent—To assess the quality of effluent-
exposed leaf material as a food source for crayfish, whole sugar maple (Acer
saccharwn) leaves were soaked at five sites in July 1977. Hint and ashpit
drain sites were the same as those in the crayfish caging experiment (A-4
and R-4), but the Rocky Run Creek sites were 0.3 km below the confluence
with the ashpit drain near the old bridge abutment and above the confluence
at R-l (Figure 1). An additional site, A-2, was also chosen immediately
downstream from the confluence of the ashpit drain with the mint drain.
Leaves in nylon mesh bags were anchored to the substrate with bricks and
left in the water for 2 weeks, then returned to the laboratory for metal
analysis. Samples were of two types: 1) a composite of leaves soaked at four
separate time periods and dried; 2) a sample of leaves soaked during one
time period and frozen. The leaf samples were analyzed by the University
of Wisconsin nuclear reactor as described in the section on metal analysis.
Metal Analysis—Crayfish were dissected under a laminar flow hood using
separate stainless steel tools for each tissue removed. Chrome-plated
stainless steel has a high amount of chromium, so scissors and scalpels were
used as little as possible to avoid metal-on-metal abrasion that might
contaminate the samples. Instruments were carefully cleaned after each
dissection by washing with "Micro" brand detergent and rinsing with a series
of solutions: distilled water, distilled-deionized water, methanol, and
distilled-deionized water. Exoskeleton, gill tissue, the hepatopancreas,
and the abdominal muscle were removed and placed in separate preweighed 0.4-
dram polyethylene vials, reweighed to obtain wet tissue weights, and oven
dried at 40°C to obtain dry weights. Carcasses were placed in preweighed
30-ml jars and oven dried. Carcass dry weight and the four tissue dry
weights were summed to obtain an approximate whole animal dry weight. Leaf
samples were torn into small pieces with forceps, placed in 0.4-dram vials,
and oven dried. Metal concentrations in exoskeleton and abdominal muscle
were determined by the University of Wisconsin nuclear reactor using neutron
activation. The neutron flux was 17 x 10 neutrons/cm^/sec. Gill samples
were not analyzed.
Hepatopancreas samples were too small to be analyzed by the nuclear
reactor. After drying, each hepatopancreas was transferred to preweighed
high purity quartz tubing (1.5 mm inner diameter), reweighed, and heat-
sealed. The following standards were prepared and placed in identical
tubes: 1) synthetic liquid standard containing 40.0 yg/ml Cr, 60.0 Pg/ml
Ba, 30.0 yg/ml Sb, and 5.00 y/ml Se in 0.4 M HN03; 2) National Bureau of
Standards - Standard Reference Material #1571, orchard leaves; and 3)
Canadian Certified Reference Materials Project, SO-4, Bottle 289 (described
57
-------
by Koons and Helmke 1978). These three standards were also included as
samples with the tissues sent to the nuclear reactor to determine the
accuracy of reactor analysis. This information is given in Appendix F. The
hepatopancreas samples and standards were irradiated at the nuclear reactor
•JO O
with a neutron flux of 8 to 10 x 10 neutrons/cm /sec and allowed to cool"
for approximately 2 weeks. The quartz tubes were then taped to the centers
of posterboard cards to permit replication of sample geometry and each was
radioassayed for approximately 6 h using two lithium-drifted geranium Li(Ge)
detectors as described for the chromium ingestion experiment. A Tracer
Northern Model TN-11 computer-based multichannel analyzer processed the
signals (Koons and Helmke 1978).
Where metal concentrations were high enough, peak areas were calculated
by computer, with adjustments for background and decay. In other cases,
peak areas were calculated by the same method used for assays of the live
crayfish in the chromium feeding experiment.
Although data on a large number of metals were obtained by both
methods, only five metals (chromium, barium, zinc, selenium, and iron) were
studied (Table 14). These were selected because of their elevated
concentrations in the ash effluent or organisms (Helmke et al. 1976a,
1976b). When two energy peaks were obtained for the same metal, a mean
concentration was calculated. Interference caused discrepancies in some
samples and these data were discarded.
Results—
Effects of Ash Effluent on Water Quality—Ash effluent inputs to the
mint drain and Rocky Run Creek resulted in consistent chemical and physical
differences in water quality (Table 15). Conductivity was greatly increased
in the ashpit drain (A-4) due to a high concentration of ions, principally
sodium, and was also much higher in Rocky Run Creek downstream from the
confluence with the ashpit drain (R-4). Alkalinity and hardness were much
lower at the effluent-affected sites (A-4 and R-4) than at the control sites
(A-l and R-2). In the ashpit drain, pH was somewhat higher than in the mint
drain and a very slight increase persisted at the downstream Rocky Run Creek
site. The very high^pH of the undiluted ashpit effluent is reduced before
it reaches the mint drain by the addition of sulfuric acid. At other times
of the year, pH values in the ashpit drain were consistently lower than in
the mint drain.
Concentration of dissolved oxygen was higher in the ashpit drain than
in the mint drain and was slightly lower at the downstream Rocky Run Creek
site (R-4) than at the upstream site (R-2). Oxygen variation was probably
due more to the effects of current speed than to the nature of the effluent
itself. Current speed was much greater in the ashpit drain (A-4) than in
the mint drain (A-l) because of its greater volume of water and was reduced
in Rocky Run Creek downstream from the confluence with the ashpit drain (R-
4). Groundwater inflow from the cooling lake slightly warmed the ashpit
drain water (Stephenson and Andrews 1976). This temperature increase
58
-------
TABLE 14. LIST OF METALS ANALYZED
IN LEAF AND CRAYFISH TISSUE SAMPLES
FOR THE CRAYFISH CAGING AND
CHROMIUM INGESTION EXPERIMENTS*
Metal
Half life
(days)
Gamma ray energy
peak location(s)
(KeV)
Soil Science Analyzers
Chromium 27.8
Barium 12.0
Zinc 243
Selenium 120
Iron 45.6
U.W. Nuclear Reactor
320
496
1,115
265, 280
1,099, 1,292
Chromium
Barium
Selenium
Iron
Zinc
Cadmium^
27.8
12.0
120
45.6
243
2.2
320
216
265
1,099
1,115
528
Hepatopancreas samples were analyzed using
detectors at the University of Wisconsin Soil
Science Department; all others were assayed at
the University of Wisconsin nuclear reactor.
TCadmium was below the detection limit in
almost all samples, except for some leaf
material.
sometimes persisted in Rocky Run Creek. Turbidity was greatly increased in
the ashpit drain, but was slightly higher at the downstream Rocky Run Creek
site than at the upstream site.
The dilution of the ash effluent by Rocky Run Creek water was
observed. The differences between control and downstream sites for all
parameters measured were much smaller in Rocky Run Creek (R-2 and R-4) than
in the ashpit drain system (A-l and A-4). When the generating station was
not operating, most parameters (conductivity, alkalinity, hardness,
dissolved oxygen, current speed, and temperature) in the ashpit drain (A-4)
and Rocky Run Creek downstream (R-4) returned to levels similar to those at
their control sites (A-l and R-2).
59
-------
TABLE 15. CHEMICAL AND PHYSICAL PARAMETERS OF SITE WATER DURING CRAYFISH CAGING+
Ashpit drain
(A-4)
Temperature (°C)
Current speed
(cm/sec)
Conductivity
(ymhos/cm
at 25°C)
Alkalinity, phenol-
g phthalein (ppm)
Alkalinity, total
(ppm)
Hardness (ppm)
PH
Turbidity (JTU)
Mint Drain
11.2±4.04
3.8-16.2
13
5.913.6
3.5±11.1
4
457112
438-476
13
0
4
232.52±18.37
211.07-251.00
5
253.18±1.56
251.60-255.37
5
7.16±0.15
7.02-7.35
5
6.2±2.6
4-13
4
Pumping
12.80±4.50
4.8-17.5
10
25.5±12.8
13.4-38.9
3
1.844±439
1,209-2,443
10
0
4
100.20±12.12
90.60-114.50
4
167.05119.39
148.19-190.41
4
7.39±0.15
7.29-7.59
4
1815.6
13-24
3
No pumping
11.60±4.07
9.2-16.3
3
7.4
1
553±121
482-693
3
0
1
227.81
1
244.00
1
7.55
1
17
1
Rocky Run
upstream
(R-2)
10.99±3.91
3.6-14.8
12
21.214.1
17.3-25.9
4
527±62
479-636
12
0
4
256.72±22.12
240.73-283.25
5
272.18±3.62
268.80-276.49
5
7.65±0.09
7.52-7.75
5
8.1±1.5
6-9.5
4
Rocky Run
(R-4)
Pumping
11.60*4.18
3.8-16.2
10
16.512.7
13.4-18.5
3
900±210
672-1,290
10
0
4
201.36116.76
176.73-213.50
4
233.40117.92
216.40-258.62
4
7.73±0.09
7.60-7.79
A
9.0±2.6
7-12
3
No pumping
9.3311.88
8.2-11.5
3
14.9
1
529187
473-630
3
0
1
232.94
1
239.2
1
7.75
1
9.5
1
+Separate analyses were performed for the ashpit drain (A-4) and downstream Rocky Run Creek (R-4) during the 16 days
that no pumping occurred from the ashpit. Mean 1 standard deviation, range, and sample size are given.
-------
The filtered waters used in the respirometers were chemically similar
to the unfiltered site waters (Table 16). The relative ordering of the
sites was identical for conductivity, alkalinity, and hardness, and the
differences for the pH measurements were minor. In the tap water,
conductivity and alkalinity were similar to the control site waters and pH
and hardness were higher than any of the site waters.
TABLE 16. CHEMICAL PARAMETERS OF WATER USED
FOR MEASUREMENT OF METABOLIC RATES
Site
Ashpit Rocky Run Rocky Run
Mint drain drain upstream downstream Tap water
(A-l) (A-4) (R-2) (R-4)
Conductivity
( mhos/cm) 446 1,519 484 734 521
at 25°C pH 7.73 7.78 7.95 7.78 8.21
Phenolphthalein
Alkalinity (ppm) 000 00
Total alkalinity
(ppm)
Hardness (ppm)
217.28
252.13
93.70
149.41
248.51
273.64
176.73
212.74
217.47
304.09
Length of Exposure and Mortality—Three mortalities occurred during the
caging experiment and eight individuals escaped from the cages (Table 17).
Escaped and dead crayfish were replaced during the first 5 weeks. One
control mortality took place in the first few days, while the two ashpit
drain crayfish died during the last 2 weeks of exposure. Visual examination
of the data indicated that the three ashpit drain crayfish with less than 42
days of effluent exposure were not consistently different from other ashpit
drain crayfish in metabolic rate or metal concentration (Harrell 1978).
Length of exposure was therefore not considered during further analysis.
Sublethal Effects—Metabolic Rate—For both experiments—site water and
tap water—the order of the mean weight-independent metabolic rates from
highest oxygen consumption to lowest was as follows: Rocky Run Creek
upstream (R-2), mint drain (A-l), Rocky Run Creek downstream (R-4), and
ashpit drain (A-4) (Table 18). Metabolic rates declined in all groups when
they were transferred to tap water (Table 18). The groups had significantly
different metabolic rates in tap wauer (P = 0.036) and approached
significant difference in site water (P = 0.060) (Table 19, Row 1). In both
61
-------
TABLE 17. SURVIVAL, LENGTH OF EXPOSURE TO ASH EFFLUENT, AND
MORTALITIES AMONG CRAYFISH CAGED AT FIELD SITES'1"
Survived until termination
of experiment
Mortalities
Site
A-l
A-4
R-2
R-4
No. of
crayfish
5
1
4
2
7
1
2
10
1
12
No. of
days caged
on site
62
58
51
48
62
27
48
62
58
62
No. of
days exposed
to full ash
effluent
0
0
0
0
46
27
32
0
0
46
No. of
crayfish
1
1
1
0
0
No. of
days
survived
4
55
62
No. of
days exposed
to full ash
effluent
0
39
46
+Length of exposure is less than time caged due to plant shutdown for 16
days. Crayfish caged for less than 62 days (full length of experiment)
were replacements for dead or escaped crayfish. All surviving crayfish
were used for respirometry. Sites were at the mint drain (A-l), ashpit
drain (A-4), Rocky Ran Creek upstream (R-2), and Rocky Run Creek downstream
(R-4).
experiments, the control crayfish (A-l and R-2, pooled) had significantly
higher metabolic rates than the effluent-exposed crayfish (A-4 and R-4,
pooled) (Table 19, Row 2). Metabolic rates of mint drain (A-l) and ashpit
drain crayfish (A-4) were significantly different in their site water, but
not when they were transferred to tap water. There were no differences in
metabolic rate between the two Rocky Run Creek sites (R-2 and R-4).
Some respirometers reached low levels of dissolved oxygen, leading to
concern that stress caused differences in oxygen consumption (Larimer and
Gold 1961, Wiens and Armitage 1961, McMahon et al. 1974). Final oxygen
concentrations in the site water experiment ranged from 8.00 to 0.78
mg/liter, with five of the 44 respirometers less than 3.00 mg/liter.
However, there was no difference in distribution of low oxygen respirometers
between groups of crayfish (Analysis of variance, F = 1.089, n.s., d.f. =
3,40) and, therefore, relative differences in metabolic rate were probably
not affected.
62
-------
TABLE 18. WEIGHT-INDEPENDENT METABOLIC RATES (K = mg 02 CONSUMED/H/l.Og
SIZED CRAYFISH, WET WEIGHT) FOR CRAYFISH CAGED AT FOUR SITES"1"
K
Site (Mean)
R-2 0.03327
A-l 0.03258
R-4 0.02931
A-4 0.02483
All crayfish
R-2 0.02009
A-l 0.01592
R-4 0.01556
A-4 0.01315
All crayfish
Log K + S.E.
Metabolism in site
-1.478 + 0.021
-1.487 + 0.024
-1.533 + 0.033
-1.605 + 0.063
Metabolism in tap
-1.697 + 0.031
-1.798 + 0.041
-1.808 + 0.048
-1.881 + 0.051
Wet
Weight (g)
water
5.97 + 1.24
5.52 + 2.05
6.40 + 2.30
5.92 + 1.96
5.95 + 1.89
water
5.97 + 1.24
5.52 + 2.05
6.44 + 2.20
5.92 + 1.96
5.96 + 1.87
No. of
Samples
11
12
10
44
11
12
12
10
45
+The 1.0-g crayfish is a hypothetical unit-size crayfish. Mean wet weight,
W + S.D., and sample size are also given. Values of log K were used for
statistical comparisons.
TABLE 19. DIFFERENCES IN METABOLIC RATES AMONG CRAYFISH CAGED AT
TREATMENT AND CONTROL SITES. ANALYSIS OF VARIANCE AND A PRIORI TESTING
WERE PERFORMED ON DATA COLLECTED IN WATER FROM THE CAGING SITES AND
SUBSEQUENTLY IN TAP WATER"4"
Comparisons
F - ratio
Site water
Tap water
Analysis of variance
Treatments vs. controls
(A-4 + R-4) vs.
(A-l + R-2)
A-4 vs. A-l
R-4 vs. R-2
2.67 n.s. (d.f. = 3,40) 3.12* (d.f. = 3,41)
5.937* (d.f. = 1,40)
5.741* (d.f. = 1,40)
1.258 n.s. (d.f. = 1,40)
4.871* (d.f. = 1,41)
1.945 n.s. (d.f. - 1,41)
3.660 n.s. (d.f. = 1,41)
+Degrees of freedom are given for the numerator mean square and the
denominator mean square.
63
-------
Nine of the 12 empty respirometers gained oxygen, resulting in a mean
gain of 0.15 + 0.37 mg O^/respirometer for the site water experiment. This
gain was randomly distributed among types of water (Kruskal-Wallis Test, H =
4.79, n.s. d.f. = 3); therefore, it is unlikely to have biased the results.
Metal Uptake—Crayfish exposed to ash effluent accumulated all metals
studied, but tissues differed in the degree to which they acquired the
metals (Figure 22). Chromium was located primarily in the hepatopancreas,
but there were smaller amounts in muscle and exoskeleton samples. Barium
appeared in the hepatopancreas and exoskeleton, selenium in the
hepatopancreas and muscle, and iron in the hepatopancreas. Iron in the
exoskeleton was .the only element significantly lower in the ashpit drain as
compared to the mint drain. Zinc occurred in all three tissues analyzed.
The analytical method did not distinguish between metals adsorbed onto the
exoskeleton surface and those actually assimilated into the tissue. Taking
this into consideration, the hepatopancreas incorporated the greatest
concentrations of most metals, with the exception of zinc, which was highest
in muscle. These results were quite similar to those obtained from tissue
analysis of the chromium-fed crayfish (see Table 24).
Differences in tissue metal concentrations among the four crayfish
groups were usually significant (Table 20, Col. 1). Only chromium in the
muscle and barium in the hepatopancreas did not show differences. Results of
the a priori, tests (Table 20, Cols. 2 through 4) were more variable. The
only difference between the two Rocky Run Creek crayfish groups was the
higher chromium in the hepatopancreas of the downstream group (Table 20,
Col. 4). Effluent-exposed crayfish (A-4 and R-4, pooled) had higher metal
concentrations in most cases than did the controls (A-l and R-2, pooled)
(Table 20, Col. 2). This difference between controls and treatments
appeared to be caused by the elevated metal levels in ashpit drain crayfish
as compared to their controls in every case except for chromium in the
hepatopancreas (Col. 3). Even in those tissue-metal combinations where most
values were below the detection limit, any values above the limit were
usually ashpit drain samples (Table 20). This may indicate that with low
enough detection limits, a significant difference between crayfish groups
might have been found.
Leaves soaked at effluent-exposed sites generally had higher concen-
trations of chromium, barium, and selenium than the control site leaves
(Table 21), with the downstream Rocky Run Creek (R-4) leaves higher than
those from the ashpit drain (A-4). Values for iron were similsfr at the
control (A-l) and Rocky Run Creek (downstream of A-4) sites but were lower
in the ashpit drain (A-4). There was no obvious pattern for zinc.
Differences between single and composite samples could be due to seasonal
variations or to the fact that single samples were frozen before they were
dried.
Discussion—
Mortality—Exposure to the .effluent had no significant lethal
effects. Total mortality during the exposure of the crayfish to ash
64
-------
HEPATOPANCREAS MUSCLE
EXOSKELETON
2O
~ 2 15
£* tn
% o IO
3 or
5 5
0
25
<£ 1 20
o. or 15
-g 10
5
0
50
^ § 40
|z 30
3 !j 20
(o 10
0
200
-go 150
c ^
Q. ~ 1 OO
~ 50
-
.
i
1
i i
«
"T i •
" i T
- T
-
j j
-
_
-
4 9
-
-
- I I '
-
.
800
? Z600
Q. 0
3-400
200
n
-
~
LH
3
k 2
. i •
c
i
j
r T I0
T 8
u 6
: I - 4
-A f 1 ! 2
1 1
T
300
not
detectable 200
100
i i i i «
T
_ J_ t 1 I
W -. - -
1
,
06
0.5
k 0.4
0.3
. O.2
0.1
f\
T
^ I
T s^ i l
not
detectable
i i i t
\J "
, 600
400
200
i ^\
t
i 600
- T J- T
t T
-IT o 400
- T O
- 1 1 200
-
- , T
:f }
W v/
I200h
i
not 90°
detectable gOO
300
_j 1 1 1— O
•
O
1,1
R-2 A-l R-4 A-4
SITE
10 12 12 II
SITE
12 12 12
SITE
12 12 12
Figure 22. Concentrations of five metals in the tissues of crayfish caged
at four sites. Means are given in ppm (dry weight) with 95%
confidence intervals. Where individual samples were below the
detection limit, they were assigned a concentration of 0 ppm
and included in the calculation of the mean. Sites are shown in
order of increasing effluent concentration, with R-2 and A-l as
control sites with no effluent. Sample sizes are listed below
the site names.
o = A-l (mint drain)
• = A-4 (ash pit drain)
= R-2 Rocky Run Creek Upstream
A= R-A (Rocky Run Creek Downstream
65
-------
TABLE 20. SIGNIFICANCE TESTS FOR DIFFERENCES IN TISSUE METAL CONCENTRATIONS IN CRAYFISH CAGED AT
FOUR SITES. WHERE THE DATA APPEARED NOT TO BE NORMALLY DISTRIBUTED (CHROMIUM IN MUSCLE AND
HEPATOPANCREAS, BARIUM AND ZINC IN EXOSKELETON), A NON-PARAMETRIC ANALYSIS OF VARIANCE
(KRUSKAL-WALLIS TEST) GAVE THE SAME DEGREE OF SIGNIFICANCE AS DID ANALYSIS OF VARIANCE, AN
INDICATION THAT THE ANALYSIS WAS NOT AFFECTED BY DISTRIBUTION PROBLEMS
F - ratios
Metal
and
tissue
Cr in muscle
Cr in exoskeleton
Cr in hepatopancreas
Ba in muscle
Ba in exoskeleton
Ba in hepatopancreas
Se in muscle
Se in exoskeleton
Se in hepatopancreas
Zn in muscle
Zn in exoskeleton
An in hepatopancreas
Fe in muscle
Fe in exoskeleton
Fe in hepatopancreas
Analysis
Variance
(1)
0.4557 n.s.
13.265***
12.986***
Undetectable
3.897*
1.317 n.s.
17.042***
Undetectable
10.672***
4. 708**
3.474*
4.039*
Undetectable
10.097***
7.367***
A-4 + R-4
vs.
A-l + R-2
(2)
0.8106 n.s.
16.53***
4.777*
3.008 n.s.
2.474 n.s.
A priori Tests
A-4
vs.
A-l
(3)
0.5958 n.s.
27.22***
0.0290 n.s.
1.336 n.s.
1.536 n.s.
32.69*** 42.33**
except in three A-4 samples
19.759*** 24.450***
8.612**
6.910*
3.108 n.s.
except in one
7.306*
9.824**
12.085**
9.131*
5.447*
A-4 sample
14.454***
10.386**
R-4
vs. H
R-2 (Kruskal-Wallis
(4) (5)
0.2711 n.s 0.550 n.s.
0.3488 n.s.
22.89*** 147.075***
1.455 n.s. 10.523*
1.313 n.s.
2.574 n.s.
2.343 n.s.
0.4715 n.s.
0.4870 n.s. 8.259*
0.1472 n.s.
0.206 n.s.
2.225 n.s.
-------
TABLE 21. CONCENTRATIONS OF METALS IN SUGAR MAPLE LEAVES
SOAKED AT FIVE SITES IN THE ASH BASIN DRAINAGE SYSTEMS"1"
Site
Kind of sample
Metal concentration (ppm dry weight)
Ba Cr Fe Se Zn
A-l
A-4
0.3 km
downstream
Single t
Composite
Single "f
Composite
Single ^
Composite
175.1
122.0
355.6
176.8
371.6
266.6
10.78
7.49
39.73
14.35
51.93
28.49
14,080
5,981
1,795
3,938
12,720
6,524
b.d.
0.1482
1.1060
0.4981
0.9121
0.9658
417.3
217.4
245.2
238.8
244.5
193.7
from R-4
R-l
Single T
Composite
121.4
197.6
14.73
23.84
5,933
1,690
b.d.
0.8892
336.0
237.6
Single samples consisted of leaves frozen after one 2-week soaking
period (Removed 20 July 1977). Composite samples consisted of leaves
from several 2-week soaking periods (15 and 23 June, 1 and 7 July 1977).
These were dried and combined for analysis, b.d. = below detection limit.
fSingle sample treated as above for 20 July. Composite sample from
1 and 7 July only.
effluent was low. The only non-ashpit drain death was in the mint drain and
was possibly due to poor initial health of that particular crayfish. Both
additional deaths occurred in the ashpit drain near the end of the
experiment and may indicate the beginning of a trend, but sublethal effects
most likely will limit the crayfish population in the ashpit effluent.
Metal Uptake—Crayfish living in the ashpit drain accumulated all of
the metals studied (chromium, barium, zinc, selenium, and iron). By the
time the effluent is diluted by Rocky Run Creek, environmental metal
concentrations are reduced, and with the exception of chromium, the elements
do not appear in the crayfish tissues in statistically significant
amounts. It appears anomalous that chromium concentrations in the
hepatopancreas were elevated over control levels in downstream Rocky Run
Creek crayfish (R-4), but not in ashpit drain crayfish (A-4). Helmke et al.
(1976a) reported that chromium concentrations in suspended particulates in
the ashpit drain increase with distance from the ash basin. If this
tendency toward increased chromium precipitation continues in Rocky Run
Creek, there may be more chromium available to crayfish in the Rocky Run
Creek sediments than in the ashpit drain. The higher chromium
concentrations in effluent-exposed leaves from Rocky Run Creek than in those
from the ashpit drain further support this hypothesis.
67
-------
The five metals accumulated in different body tissues. All except
barium were found in the hepatopancreas and all but selenium were found in
the exoskeleton; selenium and zinc were also found in the abdominal
muscle. These data do not indicate whether metals observed in exoskeleton
samples are due to surface adsorption or to actual tissue incorporation.
Schoenfield (1978) discussed this in detail and suggested using
metalrscandium ratios to answer this question. He found that whole Asellus
racovitsai, a benthic detritivore from the ashpit drain, had high levels of
scandium indicating inorganic contamination. Ingestion of sediment hydrous
iron oxides precipitated on the body surface could cause this
contamination. Thus, the high concentrations of many metals in crayfish
exoskeletons may be largely due to surface adsorption rather than tissue
assimilation.
The importance of this surface contamination as a route for metal
incorporation in other tissues is unclear. Exoskeleton permeability is a
significant factor in metal assimilation by crustaceans and polychaetes
(Bryan and Hummerstone 1973), but the crayfish exoskeleton may be relatively
impermeable to some metals (Wiser and Nelson 1964). In either case, a
tendency for metal adsorption to the surfaces of organisms indicates a
potential source of metal contamination for crayfish in the ash effluent
drainage system, whether from direct transport through the integument or
from ingestion of surface deposits on detritus and prey organisms.
The high concentration of chromium in the hepatopancreas agreed with
most findings reported in the literature (discussed in the section on
chromium concentrations in laboratory-exposed crayfish). The hepatopancreas
is important in dynamics and storage of many other metals as well, such as
lead and copper in Asellus (Brown 1977), copper and zinc in the shrimp,
Crangon (Bryan 1971), zinc in crabs, lobsters, and freshwater crayfish
(Bryan 1966, 1967), and cobalt in crayfish (Wiser and Nelson 1964). Luoma
(1976) found that mercury concentrations in the crab, Fhalamita orenata,
were higher in the viscera than in the body muscle. Thus, the
hepatopancreas has a most important function in storing excess amounts of
metals taken into the body and releasing them to other tissues or excreting
them.
It is important to compare the data with Schoenfield's (1978) data on
metal concentrations in organisms collected in the Columbia ash effluent
system. Tissues of similar physiological function in frogs, Rana pipiene,
and crayfish were similar in metal concentrations (Table 22). The only
order-of-magnitude discrepancies were for selenium in the liver and
hepatopancreas, and for zinc in the muscle. This may be due to different
physiological mechanisms for dealing with metals in an amphibian vs. a
crustacean or to differences in the amount of time animals were exposed to
the ash effluent. Crayfish fed chromium in the laboratory also contained
similar tissue chromium concentrations. With the exception of high zinc in
the crayfish muscle, both studies implicate the hepatopancreas and liver as
tissues that concentrate metals. This is expected when the function of
these organs in metal detoxification, storage, and elimination is
considered.
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TABLE 22. MEAN METAL CONCENTRATIONS IN TISSUES OF FROGS,
CAGED CRAYFISH, AND LABORATORY CRAYFISH4"
Liver (frogs) or
Hepatopancreas (crayfish) Muscle
Chromium (Cr)
Frogs 4.9 1.9
Caged crayfish 6.2 1.8
Laboratory crayfish
Cr§ 2.3 t
Cr11 5.9 1.9
Barium (Ba)
15 t
Caged crayfish 20 ^
Iron (Fe)
Frogs 770 30
Caged crayfish 640
Selenium (Se)
Frogs 2.3 0.9
Caged crayfish 33 0.4
Zinc (Zn)
Frogs 106 21.4
Caged crayfish 163 625
+Frogs were collected from site A-5 in Figure 1 (Schoenfield 1978). Caged
crayfish were collected from site A-4 in Figure 1.
$ Indicates metal concentration below the detection limit.
§ Crayfish in the laboratory exposed to Cr.
1TCrayfish in the laboratory exposed to 51Cr-labeled Cr.
Ash Effluent Characteristics Affecting Crayfish Metabolism—Long-term
exposure to effluent in the ashpit drain reduces the metabolic rate of
crayfish and this reduction is more pronounced before further dilution
occurs in Rocky Run Creek. There are three ash effluent characteristics
that might play some role in the decreased metabolic rate of crayfish held
in the ashpit drain and Rocky Run Creek below the confluence: Increased
ionic concentration, reduced food supply, and increased heavy metal
concentration. These parameters are among those frequently found to alter
metabolic rates (Wiens and Armitage 1961, Vernberg et al. 1973, Rice and
Armitage 1974, Frier et al. 1976, Nelson et al. 1977). Other important
factors such as time of day, season, and activity were constant among
69
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treatments. Differences in weight were corrected for and differences in sex
did not affect metabolic rate.
The increase in conductivity at sites receiving ash effluent may appear
to explain the variations in metabolic rates of the crayfish based on both
the substantial conductivity increase in the ashpit drain and on the widely
reported effects of salinity changes on metabolism. However, a comprehensive
study of the experimental results and the literature suggests that increased
metal levels and decreased food supply are more valid explanations.
Conductivity is important primarily as an indicator of the concentration of
an ash effluent containing coal combustion byproducts, including trace
elements, organic contaminants, fly-ash particles, and salts.
Recent work on the effects of environmental variables on invertebrate
metabolic rates has involved marine or brackish water species. Although
ionic composition may differ, salinity and conductivity are both expressions
of ion concentrations. Metabolic rate changes in the ash effluent might be
compared with results obtained in sea water, particularly since the high
conductivity of the effluent is due primarily to sodium ions, a major ionic
component of sea water. There is disagreement over the effects of salinity
on metabolic rate (Nelson et al. 1977). Some authors suggest that increased
metabolic rates at salinities differing from the organism's isosmotic point
(point of equal osmotic pressure) indicate an increased energy cost due to
osmoregulation (regulation of osmotic pressure in the body of an
organism). Others report either a decrease in metabolic rate in non-optimal
salinities, or else no correlation between them. The work of Nelson et al.
(1977) with the prawn, Macrobraehium roseribergii indicates a reduced
metabolic rate when certain salinity levels are exceeded at various
temperaures. Frier (1976) notes a marked increase in isopod oxygen
consumption in low salinity water and Taylor et al. (1977) report the same
increase for marine crabs, Caroinus maenas, exposed to 50% seawater at
10°C. Taylor et al. (1977) also found that this increase did not occur at
18°C and attributes this to quiescence and failure to osmoregulate in warmer
waters. In the cooler water, however, the crabs used oxygen through
osmoregulation and hyperactivity—possibly an avoidance mechanism. Vernberg
et al. (1973) also report a decline in metabolic rate in less optimal
temperature and salinity conditions for larval crabs, Uea pugilator>.
Thus, in many marine species, oxygen consumption is lowest in solutions
isosmotic with the blood. This can perhaps be extended to freshwater
organisms that must continually use oxygen to osmoregulate. In an
environment with higher conductivity, the water may be closer to the osmotic
level of the blood and, consequently, less energy would be needed for
osmoregulation. If this is true, the ashpit drain may be a beneficial
environment rather than a hazard; however, there are reasons to doubt
this. Most freshwater animals can not tolerate high salt concentrations,
especially when only one salt is present (Hynes 1960), as is the case for
the sodium in the ash effluent. The increased ionic content could adversely
affect the organism's osmotic balance. If this reduced enzyme activity or
impaired ability to obtain and transport oxygen, a lower rate of oxygen
consumption would result. Prolonged exposure could cause cellular
starvation or hypoxia and death could occur. Some of the data indicate that
70
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the mechanism of increased conductivity resulting in decreased metabolic
rate is not entirely applicable. In the mint drain where the lowest
conductivity is found, crayfish do not have the highest metabolic rates;
crayfish from the Rocky Run Creek control site do. In addition, the ashpit
drain and downstream Rocky Run Creek crayfish respond to tap water of lower
conductivity with lower oxygen consumption rather than the hypothesized
increase. The reduction in metabolic rate in tap water instead may be a
response to the stress of acclimating to a new chemical environment. Further
evidence of stress lies in the change in the regression coefficient for the
relationship between log weight and log 02 consumed/h. The value of b in
site water (crayfish of all treatments pooled), 0.751, agrees well with the
reported values near 0.74 (Prosser 1973). When exposed to tap water, b is
considerably higher—0.939. Vernberg and Vernberg (1969) report that the
value of b may change with temperature, salinity, environmental history, or
geographic population. If the change from site to tap water is indeed a
stress to the animal, the value of b as well as the metabolic rate could be
considerably altered while the organism acclimates to its new environment.
In summary it appears that conductivity does not control metabolic
rate, but indicates changes in the concentration of the ashpit effluent as
it progresses downstream, an effluent that contains some other factor(s)
causing reduced metabolism in crayfish. Decreased food supply and increased
metal concentrations are potential causes of the changes in oxygen
consumption.
No attempts were made to assess quantity and quality of food available
for crayfish at the various sites. However, from visual observations, it
appeared that quantity and perhaps quality are much lower in the ashpit
drain than at the other sites. The most food is available at the upstream
Rocky Run site followed by the mint drain. The mint drain is a narrow
drainage ditch with much overhanging vegetation, primarily sedge grasses,
and considerable duckweed, Lernnat on the surface. Rocky Run Creek drains a
marsh system but it also receives detritus from macrophyte beds and flood-
plain forest. Both the mint drain and Rocky Run Creek have dark brown
sediments with high organic content. The ashpit drain, in contrast, is
diked for all of its length upstream from the caging site. There is little
overhanging bank vegetation, and even less aquatic vegetation. The current
is rapid, possibly removing detritus from the area. The sediment is lighter
brown with more sand and contains much less organic matter. Higher
turbidity may reduce photosynthetic activity. This reduces habitat
diversity for aquatic vegetation and for animals that crayfish eat. High
metal concentrations may reduce photosynthetic activity in plants, further
reducing the food supply in the ashpit drain and downstream Rocky Run
Creek. Clendenning and North (1960) found a 50% reduction in kelp,
Maorooystis pyrifera, photosynthesis when 5 ppm of hexavalent chromium was
added to the water.
Insufficient food may explain the deaths of two ashpit drain crayfish
after 2 months at the site. Reduced food supply may also result in lower
oxygen consumption. In a review of the literature, Newell (1973) reports
that starvation is associated with reduced metabolic rate in many intertidal
71
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invertebrates. Lower feeding rates may reduce total activity levels (hence,
lower metabolic rates). This may be an adaptation to use less energy when
less food is available. The highest concentrations of chromium, barium, and
selenium occur in leaves soaked at the two effluent-affected sites (A-4 and
R-4); therefore, the quality of the food may be lower at these sites as
well.
The third explanation suggested for the reduced metabolic rates at
contaminated sites is the exposure of the crayfish to heavy metals. Although
many trace metals are essential in small concentrations, such as in enzyme
complexes and respiratory pigments, exposure to high concentrations of these
metals may interfere with a wide variety of physiological processes. These
effects have been extensively documented in the literature (Becker and
Thatcher 1973, Eisler 1973, Eisler and Wapner 1975). Metals may decrease
oxygen consumption by interfering with enzymes and oxygen transport
molecules, resulting in a reduced ability to utilize oxygen. Cellular
metabolism may become less efficient. Effects of metals on the gills may
lead to reduced oxygen exchange capabilities. De Coursey and Vernberg
(1972) found a reduced metabolic rate in Uca pugilator larvae upon exposure
to 0.18 ppm of mercury. Vernberg et al. (1973) determined that 1.8 ppm of
mercury reduced metabolism at 25° and 30°C and increased it at 20°C. They
concluded that suboptimal conditions of temperature and salinity reduced
metabolic rate and that the direction of the added stress from the mercury
was temperature dependent. Fromm and Schiffman (1958) exposed largemouth
bass to hexavalent chromium and observed reduced oxygen consumption after a
brief initial increase. They attribute this to a gradual decrease in
cellular metabolism caused by chromium accumulation in various tissues,
rather than to direct impairment of respiration. They found no significant
change in the respiratory epithelium or in opercular movements. Two species
of crabs, Maoropoda ro8tr>ata and Paahygrapsus nKnamor>atusf consumed less
oxygen when exposed to chromium (Chaisemartin and Chaisemartin 1976).
Therefore, it appears that high metal levels in the ash effluent reduced
crayfish metabolic rates by becoming incorporated into tissues and
interfering with cellular metabolism.
Visual food supply assessment and concentrations of three metals in the
hepatopancreas (chromium, selenium, and iron) follow the same ranking as
does metabolic rate of crayfish exposed to those factors in ash effluent.
Metabolic rate decreased from sites R-2 to A-l to R-4 to A-4, as did food
supply. Hepatopancreas metal levels increase in this same order. Thus,
both factors probably interact to reduce the desirability of the effluent-
affected habitats for crayfish. The metals in the effluent may be the
ultimate cause of the observed sublethal effects, responsible not only for
direct effects on crayfish metabolism, but also for the reduced food supply
available to the crayfish. Severely reduced animal populations in the
ashpit drain were observed after the power plant began operating (See
Section 3). Animal material that may be a significant portion of the
crayfish food supply in the non-contaminated environments would now be
limited in quantity in the ashpit drain.
The modification of the mint drain by the addition of ash effluent has
reduced its value as a habitat for crayfish, as well as for many other
72
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species. Heavy metal inputs, reduced food supply, altered ionic
composition, turbidity, and the precipitation of barium and aluminum floes
all play some role in the impairment of a sizeable portion of the sedge
meadow drainage system. This area is of particular importance to man as a
spawning area for game fish (Priegel and Krohn 1975, Magnuson et al. 1980).
Although dilution of the effluent by ground and surface water
substantially reduces the amount of contamination in Rocky Run Creek, the
sublethal effects observed in the ashpit drain persist there. Metabolic rate
was lower and metal concentrations were higher in crayfish from downstream
Rocky Run Creek than in crayfish caged at the upstream control site. Even
though these differences were not statistically significant (except for
chromium in the hepatopancreas), long-term sublethal effects in Rocky Run
Creek may gradually become apparent.
Other organisms and life cycle stages may be much less tolerant of
environmental impairment than are mature crayfish. Juvenile crayfish are
more susceptible to metal contamination than the adults tested in this
experiment (Doyle et al. 1976, Hubschman 1967, Van Olst et al. 1976).
Young-of-the-year Garrmarus were more susceptible to ashpit drain water than
were adults (See Experiment II, page 92). Eggs, juveniles, and recently
molted individuals may have more permeable surfaces or less efficient metal
storage and elimination systems. The crayfish, Oreoneates propinquus, which
is resistant to some metals—i.e., cadmium (Gillespie et al. 1977)—is
important in food chains (Neill 1951) and may contribute significant amounts
of metal to less tolerant organisms at higher trophic levels. Thus, the
effects of metals on other organisms inhabiting the sedge meadow drainage
system may be much greater than the effects observed in the adult crayfish
caged in the effluent.
Exposure of Crayfish to Chromium-Contaminated Food
Introduction—
Waters receiving ash basin effluent have elevated concentrations of
barium and chromium in the suspended particulate fractions (Helmke et al.
1976a). In addition, concentrations of barium, chromium, selenium, and
antimony in organisms collected from the ashpit drain are much higher than
in those from unaffected sites (Schoenfield 1978). Crayfish caged at
effluent-exposed sites accumulate significant amounts of chromium, barium,
selenium, iron, and zinc, and it is suspected that ingestion of high
concentrations in particulate forms contributes substantially to the
organisms' body burdens of the metals. Chromium was fed to crayfish in the
laboratory to determine metal uptake and tissue accumulation, as well as
mortality and sublethal behavioral effects.
Materials and Methods
Crayfish collection in September 1976 followed the procedures of the
crayfish caging experiment and the animals were held in the laboratory under
the same conditions until the experiment began. Crayfish were divided into
three experimental groups. The first group (designated Cr*) was fed leaf
73
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discs soaked in a 1.0-ppm solution of chromium in distilled water that
included a tracer amount of chromium-51. The second group (designated Cr)
was fed leaf discs soaked in a 1.0-ppm chromium solution with no radioactive
tracer. This group served as a control to detect any effects due solely to
radiation and not to the chromium. The third group of crayfish served as
controls (designated C) and were fed leaf discs soaked only in distilled
water.
Food preparation—In October 1976, yellow leaves were removed from a
single sugar maple tree, Acer sacaharum, growing along the shore of Lake
Mendota, Dane County, Wisconsin. They were oven dried for 3 days at 40°C
and stored at room temperature in large polyethylene bags. After soaking
leaves in water for 10 min to soften, a cork borer was used to cut discs 1
cm in diameter. Only entire discs without parts of the three primary veins
were used. Discs were dried at 40°C and stored in covered glass jars.
A 1.0-ppm chromium solution was selected for leaf soaking because
accumulation of chromium occurred and the concentration leveled off within 2
weeks (Figure 23). Discs soaked in 0.1 ppm chromium accumulated very little
chromium, and those soaked at 50 ppm chromium had not reached a stable
concentration after 3 weeks of soaking. It was suspected that allowing the
concentration to stabilize would reduce the variation between individual
discs. Two weeks appeared to be the optimal duration, since leaf matter may
reach its maximum nutritive value for the detritivore, Tipula, after 2 weeks
of stream conditioning (Cummins 1974). Leaf discs were pre-leached in Lake
Mendota water because this procedure nearly doubled chromium uptake (Figure
23).
The food supply for each week was prepared by soaking 300 discs for
each crayfish group in 500 ml of Lake Mendota water for 1 week, with water
changes after 2 and 4 days. Discs were transferred to the appropriate
soaking solutions: 500 ml of distilled water, 500 ml of distilled water
containing 1 ppm chromium (as potassium chromate, t^ Cr 0^), and 500 ml of 1
ppm chromium to which 80yCi of chromium-51 was added (obtained from the
University of Wisconsin Radiopharaacy as 2 yCi 51Cr in 1 ml of saline
solution). The chromium solutions contained sufficient chromium atoms to
allow each disc to attain maximum concentration. After 2 weeks in the
treatment solution, each set of 300 discs was dried at 40°C for 2 days and
stored in a separate glass jar. Before feeding to the crayfish, an
appropriate number of discs was placed in 200 ml of distilled water on a
magnetic stirrer for 1 h so that they would sink and be within reach of the
animals.
Holding and feeding procedures—In February 1977, 1 month before
beginning the experiment, six male and six female crayfish for each
treatment began acclimating to the experimental aquaria and feeding
procedures. Three males and three females were placed in each of six 38-
liter glass aquaria. A Nytex screen divided each aquaria in half. An
airstone was placed in each half in a perforated PVC plastic tube with a
cotton plug at the top to prevent breaking air bubbles from spraying
radioactive chromium into the air. Three short lengths of opaque PVC pipe
74
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500-•
Q.
CO
U
CO
LU
U
8
o£
U
400--
300 •-
200--
PRE-LEACHED
O NOT PRE-LEACHED
0
0
NUMBER OF DAYS
Figure 23. The relationship between chromium concentration and duration of
soaking for leaf discs soaked in 1.0 ppm chromium. One set of
discs was leached in lake water for 1 week prior to treatment.
Discs used in the crayfish feeding experiment contained 263 ppm
Cr (95% confidence limits: 172 to 354 ppm).
75
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served as shelter for the crayfish. A nine-square grid on the bottom of
each tank half was used for recording crayfish location.
The aquaria were placed in three livestock watering troughs (two
aquaria per tank) with plexiglass observation windows on one side. The
troughs served as thermistor-controlled water baths to maintain 20°C water
temperatures and as safety features to contain any radioactive chromium if
an aquarium broke. A 12 h light:12 h darkness photoperiod was maintained
with a 15-W white light bulb centered over each aquarium.
Temperature in each aquarium was recorded daily. Mean temperature was
21.6 ± 1.3°C. Each week, before cleaning the aquaria and changing half of
the water, conductivity was measured with a YSI model 33 meter and water
samples were collected. Samples were analyzed for pH, phenolphthalein and
total alkalinity, hardness, and dissolved oxygen using the laboratory
methods used in the crayfish caging experiment.
During acclimation and the experiment (March 7 to May 12 1977),
crayfish were weighed weekly and individually fed five leaf discs three
times per week in 0.47-liter (1-pint) freezer containers with water at about
20°C. Each crayfish was allowed to feed for 1.5 h, then returned to its
aquarium. The number of discs consumed was recorded to the nearest 0.25
disc. Feeding in individual containers permitted the amount eaten by each
animal to be determined. In addition, since chromium could leach back out
of the leaf discs into the water, the discs were kept out of the aquarium
water and the exposure of the crayfish to soluble chromium was minimized.
During acclimation, crayfish were fed discs prepared in the same way as
the control diet. Most of the male crayfish never fed well, possibly
because they were too confined in the small freezer containers. Thus, all
the males were replaced by randomly selected females and the experiment
began 1 week later.
Whole-body chromium assay—Crayfish in the Cr* group were analyzed
weekly for chromium uptake.Total chromium concentration was not determined
since this method can only determine chromium added to the sample over and
above the amount present before the experiment began. Whole-body radioassays
were performed using high resolution gamma ray spectroscopy on a (Ge(Li)
detector. The detector had a resolution of 2.0 KeV for the 60Co gamma ray
at 1.33 KeV. The signals were routed to a Nuclear Data Model 2200
multichannel analyzer (Koons and Helmke 1978). Chromium-51 releases gamma
rays at an energy of 320 KeV. By assaying each crayfish for a known time
and comparing the peak size to the peak of a standard solution of known
chromium-51 concentration, the amount of chromium-51 in the crayfish was
determined with appropriate corrections for radioactive decay.
The standard was prepared by diluting a portion of the original 2 C
51Cr solution to 40 Ci in 2 ml of distilled water (2.500 x 10~4g of Cr).
Knowing the original ration of chromium-51 to chromium in the leaf soaking
solution, and making certain assumptions, the total chromium concentration
in the crayfish was calculated. The assumptions are that both forms of
76
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chromium behave the same way in biological systems (that is, the original
Cr:Cr ratio will remain the same during all biological processes) and that
there is no isotopic replacement of chromium-51 for chromium already present
in the crayfish before the experiment began. The following equation was
used to determine chromium concentration:
Cr concentration in sample:
-4
counts/min of sample 2.500 x 10 g of Cr in standard
X
counts/min of standard sample wet weight (g)
The error introduced by dissimilar sample geometries was reduced by
immobilizing each crayfish by attaching it to a heavy posterboard card with
rubber bands. The card was placed in a thin plexiglass box with the suture
between the animal's thorax and abdomen clearly centered. The box could be
placed inside the detector in a reproducible position. Water samples, leaf
disc samples, and the standard samples were assayed in 25-ml scintillation
vials taped into the center of the box.
The box was lined with a plastic bag that was discarded after each
crayfish to prevent contamination. Each crayfish was wrapped in water-
soaked cheesecloth to minimize dehydration and assayed for 0.5 h. This was
2
not long enough to reduce analytical uncertainty to 1% but was the maximum
amount of time feasible without injuring the crayfish. Crayfish in the Cr
and C groups were also attached to cards and wrapped in wet cheesecloth for
0.5 h every week; however, they were not taken to the building that housed
the detector nor were they placed in the plexiglass box. Each Cr and C
crayfish was assayed at least once during the experiment; none indicated any
chromium-51 contamination. Additional monitoring included counting
chromium-51 labeled leaf discs and water samples from various stages in the
food preparation and feeding process to indicate final chromium content of
discs and chromium loss due to leaching. The water in the two aquaria
holding Cr* crayfish was assayed weekly and showed no contamination,
supporting the assumption that the animals were received no chromium from
the water.
error was introduced into the results because the crayfish varied in
size and were of very different dimensions from the standard in a cylindrical
vial. It is particularly important to center each sample in front of the
detector and to maintain the same distance from sample to detector.
2Analytical uncertainty is expressed as ^ , where b is the total number of
gamma rays detected in the peak. It is an estimate of the precision of
multiple analyses (Koons and Helmke 1978) and does not include error
introduced by sample weighing and handling or biological variation.
77
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Behavioral observations—Weekly behavioral observations alternated so
that any one crayfish was observed only once in 2 weeks. The location,
position with respect to shelter, activity, interaction with other crayfish,
and dominance in the interaction were recorded. The 10 possible locations
included nine squares on the bottom of the tank and the screen divider.
Activities were classified as:
L. Non-interactive: walking, climbing (on glass sides, aeration tube,
or screen), swimming, feeding, grooming, and motionless.
2. Interactive: touching, aggression, and retreating.
All patterns except motionless and retreating were "volitional," active
behavior patterns not initiated by another animal. The ethogram was
modified according to Stein and Magnuson (1976). Walking, climbing,
swimming, motionless, grooming, aggression, and touching were as defined.
Because there was no substrate in the tanks, probing, digging, and burying
were not included; copulation did not occur between females; chelae
(pincerlike claws) display in response to a predator was not possible; and
feeding consisted of using the pereiopods to pick up feces and move it
toward the mouth. Aggression described the dominant individual in an
encounter; retreating indicated the subordinate.
Individual crayfish were observed at 10-sec intervals. Each of the
three crayfish in the aquarium half was observed in turn, thus each was
observed every 30 sec. A "set" consisted of 10 observations per crayfish.
This procedure was then repeated at another aquarium. Each week, five sets
of observations were made on each aquarium, two during light and three
during darkness with 25-W red light bulbs placed in the sockets. Nail
polish dots on the sides of the body and on the chelae permitted individual
recognition at any angle. The order of observation of individuals in a
tank, of tanks within a trough, and of troughs was randomized each week.
Complete randomization of the sequence of individuals to be observed might
have resulted in disturbing the crayfish by moving from tank to tank too
frequently; therefore, all crayfish in a tank were observed at once.
Activity was higher and more varied at night, so only night-time
observations were analyzed. The number of location changes observed for
each crayfish during the three observation sets on one night was summed and
a median found for all six crayfish of each treatment observed at night.
This procedure was repeated for number of actions. The small number of
events did not permit individual types of behavior.
Friedman's Randomized Blocks test was used to analyze the effects of
treatment on activity (separately for location changes and actions).
Medians were ranked regardless of treatment or date of observation and the
rank compositions of the three crayfish groups were compared. The effect of
time on behavior, regardless of treatment, was analyzed in the same way.
Termination of experiment—All crayfish were fed control food twice
during the ninth week to clear their guts of unassimilated chromium. The
Cr* crayfish were radioassayed again.
78
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Mortalities were recorded and dead crayfish frozen individually in
polyethylene bags. Mortalities among the groups were compared with the
Mann-Whitney Wilcoxon Test. Crayfish were ranked by date of death,
independent of treatment, and each pair of groups was compared. After the
last assay, remaining crayfish were frozen and stored for tissue metal
analysis as described for the caging experiment. To prevent interorgan
metal diffusion they were not allowed to thaw until dissection (Luoma
1976). Three surviving crayfish from each treatment were later dissected.
Results—
Chromium uptake in chromium-51 labeled crayfish—Crayfish accumulated
statistically significant amounts of chromium from their food during an 8-
week period (Figure 24). The uptake was most rapid during the first week of
feeding, increasing from 0 to 16 ppb, but concentration more than doubled in
the next 7 weeks. At the end of the first week of feeding, crayfish had
retained 2.75% of the chromium ingested during that week. After the eighth
week, they had retained 1.72% of the chromium ingested during the
experiment. The difference in chromium concentration after 1 week on
uncontaminated food (week 9 to week 10) is 9.6 ppb, a reduction of 23.7%.
The mean chromium concentration in Cr* crayfish was directly proportional to
cumulative chromium ingested per crayfish (b = 0.0041 ***) (Figure 25).
There was no significant difference in total food consumption between
the three groups of crayfish (Kruskal-Wallis Test, H = 0.6585 n.s., d.f. =
2). Mean total consumption for all crayfish for weeks 1 through 9 was 24.5
discs, the median was 19 discs, and the range 0 to 74 discs. Mean dry
weight per disc was 1.40 mg and there was no difference in weight between
control and chromium-contaminated discs (t-test, n = 23, t = 1.27 n.s.).
Chromium-51 labeled discs had a mean chromium concentration of 263.4 ppra
(range 118.7 to 523.0 ppm). There was a mean of 0.384 g chromium per disc.
Lethal and sublethal effects of chromium ingestion—Mortality rate did
not differ among treatments (Table 23).Treatment groups did not differ in
2
median number of actions (Figure 26) (Friedman's Randomized Blocks Test, X
2
= 1.63 n.s., d.f. = 2) or in median number of location changes (x =1.63
n.s., d.f. = 2). Activity appeared to decline over time for all treatments
(Figure 26), but differences were not significant between observation dates
2
either for number of actions (Freidman's Randomized Blocks Test, X =4.40
2
n.s., d.f. = 3) or for number of location changes (x = 5.80 n.s., d.f. =
3).
Further attempts to determine the cause of the activity decline were
not helpful. There was no significant correlation between total food
consumption and number of actions (r = 0.07 n.s.) or number of location
changes (r = 0.28 n.s.). The correlations between final chromium
concentration (for Cr* crayfish only) and activity also were not significant
(r = -0.08 n.s., for number of actions; r = -0.04 n.s., for number of
location changes).
79
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100-
50-
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(£
UJ JE
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12
12
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mean
n
I
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C.I.
=0.
0^2
begin Cr
feeding
4
Tl
1 1 1
6
kAC L~~L..\
8
f 10
"clean"
food
Figure 24. Mean chromium concentration over time in chromium-51 labeled
crayfish. Crayfish were fed food without chromium after week 9.
The regression equation for weeks 1 through 9 is: log Cr
concentration = -7.86 + 0.054 (time). The regression coefficient
is highly significant (P < 0.001). N - sample size and C.I. =
confidence interval.
80
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50
§ 40
_ __
(J O)
Z 5
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0)
30
20
LU
0
r2=0.90
oo
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10
MEAN CUMULATIVE CR INGESTED
(GxlO6)
Figure 25. Relationship between whole-body chromium concentration and total
chromium ingested by chromium-51 labeled crayfish. Each point
represents the mean for all crayfish for 1 week in the experiment.
The regression equation is: Cr concentration - 0.0041 (Cr
ingested) + 4.9. The regression coefficient is highly signifi-
cant (p < 0.001).
81
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92-98%
C.I. '
O CONTROL
median •CHROMIUM
O LABELED CHROMIUM
16 •
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DAYS
60
Figure 26. Median number of actions performed by the three groups of cray-
fish on four dates during the experiment. Confidence intervals
of 92 to 98% and sample size are given; the 95% confidence
intervals did not correspond exactly to an integral number of
actions.
82
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TABLE 23. LETHAL EFFECTS OF CHROMIUM INGESTION
Mortality
Treatment*
Cr Cr*
No. of crayfish
surviving experiment 7 96
(out of initial 12)
% survival 58 75 50
Mann Whitney Wilcoxon test
Treatment pair tested"1" Significance level* of differences
between treatments ( )
Control vs. Cr
Chromium vs. Cr*
Control vs. Cr*
92.9 - 87.5
91.7 - 86.5
< 46.5
n.s.
n.s.
n.s.
+C = control, Cr = fed chromium, Cr* = fed chromium labeled with
51Cr.
'Exact significance could not be found because of the discrete
nature of the data.
Metal uptake—Chromium concentrations were highest in the
hepatopancreas for two crayfish groups (Table 24), but there was no
significant difference in concentration between treatments (Kruskal-Wallis
Test, H = 2.489 n.s.). Since the crayfish were not differentially exposed
to any other metals, and since visual examination gave no evidence that
treatment group affected concentration of these metals (Harrell 1978),
treatments were pooled for metals other than chromium. Barium occurred in
the heptopancreas and exoskeleton. Iron and selenium appeared only in the
hepatopancreas. Zinc was most concentrated in muscle and hepatopancreas,
but also was found in the exoskeleton.
Although cadmium occurred in the leaf discs (Table 25), it was not
present at detectable levels in any crayfish tissues. The value of 200 ppm
Cr for Cr* discs obtained by the University of Wisconsin Nuclear Reactor
analysis was lower than the 263.4 ppm obtained by the chromium-51 labeling
and radioassay.This may be a result of a wide variation among samples or the
difference in analytical method. The values for all other metals in C and
Cr discs were similar, as expected.
83
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TABLE 24. METAL CONCENTRATION IN TISSUES OF CRAYFISH FED CHROMIUM IN
THE LABORATORY (MEAN ± S.E., NO. OF SAMPLES)"1"
Tissue concentration (ppm dry weight)
Crayfish
Metal group Hepatopancreas Muscle Exoskeleton
Cr
Ba
Cd
Fe
Se
Zn
C
Cr
Cr*
all
all
all
all
all
3.153
2.297
5.943
24.03
563.0
4.383
1,250.
± 0.5108
± 0.3730
± 2.333
± 4.43
± 125.9
± 0.232
0 ± 305
(3)
(3)
(3)
(9)
(9)
(9)
(9)
(9)
4.076 ± 1.698 (3)
(3)
1.886 ± 1.540 (3)
(9)
(9)
(9)
(9)
2,108 ± 278 (9)
(3)
1.018 ± 0.8315 (3)
(3)
183.3 i 27.0 (9)
(9)
£ (9)
(9)
476 ± 45 (9)
+Analysis is broken down into the three crayfish groups for Cr
concentrations only, since crayfish were differentially exposed only to
.chromium. = below detection limit.
fOnly one sample was above detection .Limit.
Discussion—
Chromium uptake—Crayfish can assimilate chromium incorporated in their
food supply, although the amount of uptake from ingestion is low. An
assimilation of less than 3% of the amount ingested agrees with the findings
reported in the literature. Rats that had not been fed absorbed 6% of the
chromium they ingested, while rats that had been fed absorbed only 3%
(Mackenzie et al. 1959). Rainbow trout (Salmo gairdneri') assimilated no
hexavalent chromium even when it was placed directly in the digestive tract.
Despite this low level of uptake, ingestion may be an extremely
important mode of uptake where chromium concentrations are high in
particulate matter and low in dissolved form. This is the situation in the
ash effluent drainage system of the Columbia Generating Station. Chromium
concentrations in suspended particulate matter from the ashpit drain were
over 1,000 ppm shortly after the generating station began operating (Helmke
et al. 1976a). In contrast, dissolved chromium in the ashpit drain ranged
from 0.006 to 0.028 mg/liter from November 1976 to April 1977 (Andren et al.
1977). Organisms collected from the ashpit drain had elevated
concentrations of chromium and barium (Schoenfield 1978) and crayfish caged
at effluent-affected sites accumulated chromium and other metals. Since
84
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TABLE 25. METAL CONCENTRATIONS
IN LEAF DISCS FED TO CONTROL (C) AND
CHROMIUM-FED (Cr) CRAYFISH +
Concentration
(ppm dry weight)
Metal C discs Cr discs
Ba
Cd
Cr
Fe
Se
Zn
46.02
3.265
513.5
1,223
30.08
200.7
328.2
1,065
+0nly one sample of each food was
analyzed. = below detection
limit.
dissolved chromium remains low, uptake from ingested chromium may be far
more important in the ash basin drainage system than laboratory experiments
predict.
The rate of chromium uptake during the experiment appeared linear.
There was no indication that chromium concentration began to level off
during the experiment. A leveling off or distinct reduction in the rate of
increase would indicate that a stable body burden of chromium had been
reached, with the organism in a state of equilibrium with its environment or
food. Maximum chromium uptake by trout was reached after 10 days in
solutions of low chromium concentration (0.0013 and 0.01 mg Cr/liter), but
in more concentrated solutions (0.05, 0.1, and 0.15 mg Cr/liter), there was
no sign of leveling off after 30 days of uptake (Fromm and Stokes 1962).
Since uptake mechanisms from food and water are different (from digestive
tract as opposed to gills and/or integument), a direct comparison with the
trout data should not be made. However, a failure to reach equilibrium
after more than 40 days of exposure to dietary chromium at concentrations >
200 ppm does not seem inconsistent, especially since the percent
assimilation was so low.
Whole-body chromium concentrations, after exposure to uncontaminated
food (week 10), were 24% less than the previous week. This may indicate
that unassimilated chromium present in the gut during radioassay in week 9,
85
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and presumably all previous weeks, was egested by week 10. It may also be
due to loss of assimilated chromium from the tissues during the previous
week. Both egestion and tissue elimination probably were operating, but the
relative contribution of each to the decline in chromium concentration can
not be ascertained from the data. Unassimilated chromium in the gut
apparently does not contribute significantly to whole-body concentrations.
Elwood et al. (1976) found that chromium concentration in Tipula did not
decrease after gut evacuation. Schoenfield (1978) presents corroborating
evidence. Gut evacuation did not appreciably reduce the whole-body chromium
concentration of Asellue racovitzai collected from sites high in suspended
particulate and soluble chromium concentrations. Apparently, chromium was
either assimilated readily from the digestive tract or egested rapidly and
little remained in the gut contents to affect the analysis. The crayfish
were radioassayed within a few hours of feeding; thus, the undigested food
probably remained in the stomach containing chromium that the animals had no
opportunity to assimilate. However, if the results obtained by Elwood et
al. (1976) and Schoenfield (1978) can be applied to crayfish, this
undigested chromium in week 9 was digested by week 10 and played no part in
the decline in chromium concentration. The assimilation ratio of less than
3% may indicate that most chromium was egested quickly or that it was
assimilated and excreted rapidly. The latter explanation is more consistent
with the results obtained by Elwood et al. (1976) and Schoenfield (1978).
For these reasons it is suggested that chromium reduction in the crayfish is
attributable primarily to tissue loss rather than to egestion of
unassimilated chromium.
Lethal and sublethal response—There is no evidence that chromium
ingestion caused crayfish deaths or behavioral differences. However, small
sample sizes and wide variability between individual crayfish might have
masked considerable behavioral differences. The high overall death rate and
the apparent decline in activity as the experiment progressed were probably
due to poor health and low food consumption. The maximum number of discs
consumed by any crayfish was 74, with a corresponding total dry weight of
0.104 g in 9 weeks or less than 5% of the wet weight of a 3-g crayfish.
This is not an adequate ration. Food consumption in most animals declined
as the experiment progressed.
The lack of natural wintertime temperatures and photoperiods during the
previous winter in the laboratory probably caused the poor health and low
appetite of the crayfish. The animals did not experience the 4°C water and
long dark period required for proper ovarian development (Aiken 1969a).
Aiken (1969b) also reports high molt mortality among crayfish kept on an
abnormal photoperiod schedule. Other physiological processes such as metal
assimilation and transport may have been impaired in the experimental
crayfish. Therefore, the tissue metal locations and concentrations may not
be the same as those expected in healthy crayfish with adequate food
consumption.
Tissue metal uptake—The lack of statistical difference between
crayfish groups in hepatopancreas chromium concentration was unexpected in
86
-------
o
view of the increased whole-body concentration in the Cr* crayfish. Small
sample size could be the sole reason for this result, but it is more likely
that the mean whole-body increase of 0.196 ppm (dry weight) provided such a
small additional amount of chromium to the hepatopancreas relative to the
pre-experimental level that no increase was detectable. It is also possible
that isotopic exchange was replacing the chromium already in the crayfish
with the labeled chromium, thus, there would be no change in total chromium
concentration.
Mean chromium concentrations obtained by neutron activation analysis
for three of the Cr* crayfish were 1.9 ppm for muscle, below detection limit
for exoskeleton, and 5.9 ppm for hepatopancreas. These differences in
tissue concentrations indicate that the hepatopancreas, and to a lesser
extent the muscle, were sites of chromium concentration within the body.
Evidence from the literature supports these results. Schiffman and
Fromm (1959) found that exposing rainbow trout to chromium in their water
resulted in small amounts of chromium in the muscle and significant amounts
in the spleen, gall bladder and bile, kidney, and liver. Of all tissues
studied (blood, spleen, liver, muscle, gut, pyloric caeca, stomach, and
kidney) in rainbow trout, only the muscle and blood failed to accumulate
chromium at concentrations higher than those in the water (Knoll and Fromm
1960). Crustacean hepatopancreas and vertebrate liver perform similar
physiologic functions, thus, it is useful to compare metal concentrations in
the two tissues. In the lobster, Homarus ameriaanus, chromium is highest in
the gills (the site of absorption from water) and lowest in the exoskeleton;
the hepatopancreas is an important storage site for chromium and other
metals (Van Olst et al. 1976). After exposing the crab, Podophthalmus
vigil, to chromium in the water, Sather (1967) detected the following
decreasing order of radioactivity in the tissues: gills > muscle > midgut
gland (hepatopancreas) > carapace > blood. Blood had low chromium levels in
all of these studies, leading several authors to conclude that blood is the
main chromium transport mechanism and that it loses its chromium rapidly to
other tissues. Thecarapace was metabolically inactive in the lobster and
crab studies, which explains the low or undetectable exoskeleton chromium
levels in the crayfish.
Conclusions—
Crayfish assimilate ingested chromium, although the percent
assimilation and total amount are small. The relative importance of the
food and water pathways was not determined, but based on its chemical
properties, dietary uptake appears to be more important for trivalent
chromium (see literature review, Appendix D). Based on studies by Schroeder
(1973), chromium in the leaf discs was probably in trivalent form, as is the
chromium in the sediments, detritus, and organisms of the ashpit drain
system.
Because neutron activation analysis of the whole body burden
was not possible no direct comparison of the two assay methods
can be made.
87
-------
The chromium uptake by laboratory crayfish was not high enough to cause
significant mortality or behavior changes. However, under more normal food
consumption patterns by healthy crayfish, the 200 ppm of chromium in the
food might have had detectable effects. Less tolerant life stages or
organisms may have been significantly affected. Although the laboratory-
exposed crayfish appeared unaffected by chromium in their food and
assimilated very little, this experiment probably underestimates the
magnitude of the effects in the ashpit drainage system where particulate
chromium concentrations are elevated over background levels by several
orders of magnitude.
Long-Term Exposures of Asellus racovitzai to Ash Effluent in the Laboratory
Introduction—
The purpose of this study was to investigate whether Asellus exposed to
water and food from the ashpit drain prior to the spring reproductive pulse
would grow more slowly and produce fewer young than Asellus given water and
food from the mint drain. Whether any observed differences in growth or
fecundity were caused by the water and food were also to be determined;
thus, Asellus were fed ashpit drain food in control water and control food
in ashpit drain water.
By hand-netting, it was discovered that Asellus were,less abundant in
the ashpit drain (sites A2, A3, and A4) than in the mint drain (site Al) and
that few of them colonized the artificial substrates in the ashpit drain.
The reason for this did not appear to be acute toxicity because the crayfish
studied earlier survived as well in the ashpit drain as in the mint drain.
Instead, the lower metabolic rates of the caged crayfish in the ashpit drain
suggested that sublethal effects on growth and reproductive success
decreased Asellus number.
Asellus rvtcovitzai begin to mate in February in central Wisconsin
(Herbst 1975). Young are released from the females starting in April and
May. Several fast-growing summer generations follow and over-wintering
individuals are born in late summer or early fall. Winter-generation adult
isopods are larger and bear more young per female (Seidenberg 1969).
^
Materials and Methods—
The experimental design was as follows: Treatments A (control food and
water) and D (ashpit drain and water) were expected to differ*the most;
treatments B (ashpit food and mint drain water) and C (ashpit water and mint
drain food) were designed to separate the effects of food and water.
Another concern was that the Asellus in treatment C might ingest the
precipitated chemical floe in the ash effluent in addition to feeding on the
uncontaminated leaf discs. Therefore, a sub-experiment was added to compare
treatment C to a fifth treatment, E, This treatment was identical to C, but
the ashpit drain water was filtered (0.45 Millipore) to remove
particulates.
The experimental apparatus was designed to hold animals in clear PVC
88
-------
tubes (8.2 cm high x 5.8 cm diameter) in 500-ml pyrex beakers. The bottom
of the tubes consisted of PVC-coated fiberglass window screen. The animals
could be photographed for growth measurements in the tubes and water could
be changed without handling the animals. The beakers were suspended in a
water bath. Temperature in the bath was controlled by circulating water in
copper tubing between a Frigid Units chiller and the water bath outside the
beakers.
Specimens of Aseilus racovitzai. were captured by hand net in the mint
drain. Water was collected in 19-liter (5 gal) polyethylene containers at
the mint drain site, Al, and ashpit drain site, A4 (Figure 1). Leaves were
soaked in ashpit drain (A4) or mint drain (Al) water for 2 weeks prior to
feeding them to Aseilus• Leaves were collected from a sugar maple, Aaer>
sacchar'imt in the fall, oven dried, and stored in plastic bags. Individual
leaves were placed in separate compartments in nylon bags; the bags were
suspended with bricks and floats in the field at sites Al and A4 for 2
weeks. Placement of bags in the field was staggered so that a set soaked
for 2 weeks was ready to be used each week. A 15-mm diameter stainless
steel cork borer was used to cut discs from the soaked leaves.
A preliminary experiment indicated that Aseilus could be transferred
directly to ashpit drain water from mint drain water without mortalities
(Table 26) and that the isopods did eat the leaf discs.
Five individuals (64 to 96 ram long, x = 80) were placed in each of 70
beakers on 17 Dec. 1977. Each beaker contained mint drain water and leaf
discs. Animals were checked and the dead replaced daily. Exposure to
treatment water and leaf discs began on 20 Dec. 1977. Fourteen beakers were
randomly selected for each treatment. Dead animals were no longer
replaced. Weekly support of the experiment proceeded as follows:
Day 1 - Collect water and leaf bags from mint drain and ashpit drain.
Place new leaf bags into mint drain and ashpit drain for 2-week
incubation. Cut leaf discs; hold water and discs in water bath.
Day 2 - Change water and food in each beaker; remove dead animals. Water
chemistry.
Day 3 - Re-randomize the beakers.
Day 4 - Photograph animals (every 2 to 3 weeks) for growth.
Day 5 - Change water and food in each beaker; remove dead animals. Water
chemistry.
Days 1 through 7 - Check temperatures, pump, etc.
The water bath was held at 3.81 ± 0.23°C and the soft-on and soft-off
light controls were set at 9.5 h light:14 h darkness for the first 6
weeks. Temperature and photoperiod were then accelerated weekly to simulate
mid-April conditions by late February, thereby encouraging earlier
reproduction.
89
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Temperature and Photoperiod Schedule
Lights
Dates °C On Off
20 December
30 January
6 February
13 February
20 February
27 February
-29 January
- 5 February
-12 February
-19 February
-26 February
- 21 March
3.81 + 0.23
5.92 + 0.88
7.75 + 0.21
10.01 + 0.07
11.82 + 0.04
13.92 + 0.04
7:00
6:45
6:24
6:06
5:48
5:30
16:30
16:54
17:18
17:42
18:00
18:24
Results and Discussion—
Survival of Asellus among the four treatments (A through D) in the main
experiment was the same (Figure 27). Shed exoskeletons (indicating growth)
were sometimes found in the first weeks of the experiment (20 December to 24
January) and some mating was observed in late January and February.
However, the entire group stopped molting in late January and many animals
began dying in all treatments in late February. None of the females showed
evidence of carrying young and the experiment was terminated;
unfortunately no growth or fecundity data was obtained. .The reason for
these mortalities in a usually hardy animal was explored and it was learned
TABLE 26. PRELIMINARY EXPOSURE OF ASELLUS RACOVITZAI
TO MIXTURES OF ASHPIT DRAIN (A3) AND CONTROL (Al) WATER
% alive
Day 0:100 25:75 50:50 100:0 +
1
2
5
8
97
97
97
97
100
100
100
100
97
97
97
97
100
100
100
100
No. of isopods 32 32 32 32
Xashpit drain: % control water.
that other researchers had experienced a similar mid to late winter loss of
isopods and amphipods in the laboratory (Herbst, personal communication and
Kitchell, personal communication) and in the field (Herbst 1975). It was
decided to wait for the hardier summer generations before continuing.
90
-------
100
90
80
70
60
50
40
DC
ff
30
20
10'
.A-l
A-4
FOOD
A-l A-4
A
C
B
D
24
36
48 60
TIME (DAY)
72
84
96
108
Figure 27.
Survival of Asellus raeovitzai exposed to leaf litter and water
from locations upstream (sampling station Al) and downstream
(station A4) from the ash effluent. Differences between treat-
ments were not significant (p < 0.001) using a Mantel-Haenszael
test that computes X2 values for observed mortalities during
each time period.
91
-------
Survival in the filtered ashpit drain water was significantly lower
than survival in the unfiltered ashpit drain water (Figure 28), possibly
because the 0.45- filter may have removed microorganisms important to
Asellus nutrition or digestion. For future studies, the total ash effluent
should be used to simulate natural conditions.
Neutron activation analysis of samples of leaf discs confirmed that
chromium and barium were higher in the ashpit drain food than in the mint
drain food (Table 27). Water chemistry parameters measured in the
laboratory experiment confirmed field observations of differences due to the
ash effluent. Mint drain water was lower than ashpit drain water in
conductivity and turbidity and higher in alkalinity, hardness, and pH (Table
28). Dissolved oxygen was lower in the mint drain water at the warmer
temperatures late in the experiment. Filtered ashpit drain water was lower
in turbidity than unfiltered ashpit drain water.
Short-Term Exposures of Adult and Young-of-the-Year GartmaruB to the Ash
Effluent
Introduction—
Time did not allow the Asellus experiment to be repeated, so some
simpler procedures to test the effects of the ash effluent on a more
sensitive benthic crustacean, Gammarus pseudolirrmaeue, were designed. Since
Gcumams is common to Rocky Run Creek, rather than the mint drain, the
results would tell us more about the downstream effects of the effluent. In
TABLE 27. TRACE ELEMENT CONCENTRATIONS (PPM + 1 S.D.)
IN LEAVES PREPARED FOR FEEDING ASELLUS MARCH 1978
Element Control (Al) Ashpit drain (A4)
Cr 7.61 ± 3.87
Zn 1,199.1 ± 210.3
Ba -+
Cd
Se
Sb
Fe 2,988.7 + 933.3
13.76 ± 2.77
1,224.0 ± 248.6
314.8 ± 79.1
—
—
—
1,510.0 + 454.0
+— = below detectable limits.
92
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100
90
80
70
60
50
40
UJ
UJ
30
20
10
FOOD
A-1
A-4
LU
A-4
'filtered
12
24 36
48 60
TIME (CAY)
72
84
96
108
Figure 28.
Survival of Asellus paeowi-tzai, exposed to filtered (sampling
station E) and unfiltered (station C) ashpit drain water.
(Differences were statistically significant, p < 0.001, Mantel-
Haenszael test (Snedacor and Cochran 1967).
93
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TABLE 28. SUMMARY OF WATER IN PHYSICAL AND CHEMICAL MEASUREMENTS OF WATER IN
THE ASELLUS LABORATORY EXPERIMENT FROM 23 DEC. 1977 TO 17 MAR. 1978.
AVERAGE, ± S.D.; (RANGE); AND NUMBER OF SAMPLES
Treatment
Dissolved oxygen
(mg/liter)
Conductivity
( mhos/cm)
at 25°C
Total Alkalinity
(ppm)
Hardness (ppm)
PH
Turbidity
Temperature Date
°C
A
8.60±1.48
(6.16-10.26)
6
474.48147.02
(425-636)
23
210.77±17.91
(164.38-243.84)
20
250. 79116. 53
(232-291.88)
20
7. 8310. 11
(7.58-7.95)
19
1.35±0.41
(0.8-2.5)
21
23 Dec-
30 Jan.
3.81±0.23
167
B
8.7811.47
(6.17-9.96)
6
471.61±37.23
(426-553)
23
210.59±16.15
(171.44-240.16)
20
249.04il6.52
(231.6-291.89)
20
7.8310.10
(7.68-8.08)
19
1.28±0.51
(0.7-3.4)
21
Week of
30 Jan.
5.92±0.88
48
C
9.62*0.68
(8.91-10.75)
6
1,167.521109.02
(932-1,294)
23
95.01115.24
(65.18-125.94)
20
133.91123.18
(99.20-173.6)
20
7.4710.31
(7.28-7.74)
18
5.6216.90
(2.2-35.0)
21
D
9.7110.60
(9.09-10.76)
6
1,154.481116.14
(943-1,290)
23
94.84116.12
(163.24-127.09)
20
133.07122.87
(97.20-162.80)
20
7.44*0.11
(7.30-7.68)
19
5.4516.90
(1.8-32.0)
21
Week of Week of Week of
6 Feb. 13 Feb. 20 Feb.
7.7510.206 10.
48
0110.068 11. 8210. 036
48 48
E
9.5410.76
(8.80-10.68)
6
1,160.651119.50
(905-1,316)
23
91.62*16.69
(62.08-125.37)
20
135.19126.86
(97.20-195.62)
20
7.4610.11
(7.3-7.66)
18
0.6010.53
(0.25-2.0)
21
28 Feb.-
21 Mar.
13.9710.042
72
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Experiment I, it was hyposethsized that the younger animals would be more
sensitive than adults. First, the toxicity of the effluent to young-of-the-
year was demonstrated, then a series of dilutions of ash effluent in Rocky
Run Creek water was selected to determine a concentration not acutely toxic
but which might later be used to study growth and fecundity.
Materials and Methods—
Experiment I—Gammar'us from Rocky Run Creek above the ash effluent were
collected on 31 May 1978 and placed in individual beakers in the same
apparatus used for Aeellue at 17°C. Aerators were added to the apparatus
for this experiment. Twenty young-of-the-year (4.6 ± 0.6 mm body length)
and 20 adult (9.8 ±2.2 mm) amphipods were used. After 24 h acclimation in
Rocky Run Creek water, the water was changed in all beakers with half the
adults and half the young receiving ashpit drain water instead of Rocky Run
Creek water. Treatments were randomly assigned to the beakers. The beakers
were checked daily for 4 days. Dead amphipods were removed and preserved in
70% alcohol.
Experiment II—Young-of-the-year Gammar>ue were collected from Rocky Run
Creek above the ash effluent on 28 June 1978 and placed individually in 60
beakers in the water bath at 17°C. After 24 h acclimation in Rocky Run
Creek water, the water in all beakers was changed with groups of 12 randomly
selected beakers receiving one of the following water types:
100% Rocky Run Water
75% Rocky Run + 25% Ashpit Drain Water
50% Rocky Run + 50% Ashpit Drain Water
25% Rocky Run + 75% Ashpit Drain Water
0% Rocky Run + 100% Ashpit Drain Water
Beakers were checked daily for dead animals for 4 days. Water was changed
once at 48 h. Since no mortalities occurred and the ash effluent was half
as concentrated as it had been in Experiment I, the experiment was continued
for 4 more days with a water change at 48 h. The experiment was then
terminated because heavy rains continued to dilute the effluent below the
toxic level determined in Experiment I.
Results and Discussion—
Young instars of Garnnarus pseudolimnaeue were more sensitive to the ash
effluent than were large individuals of the same species (Figure 29). When
the experiments were replicated on young instars using dilutions of the ash
effluent to select a non-lethal concentration, heavy rains diluted the
effluent below its toxic level (Figure 29). The threshold for acute
toxicity of the ash effluent to young-of-the-year Gammarue falls between
effluent concentrations of 1,100 and 1,900 ymhos as estimated by
conductivity.
95
-------
to
o>
UJ
>
UJ
UJ
a.
100-
80-
60-
40-
20-
ADULT
YOUNG
11000 2000
Effluent-exposed -*•
3000
CONDUCTIVITY (/tmhos / cm)
Figure 29. Percent survival of young and adult Gammcapus pseudolimneaus
exposed to the ash effluent for 96 h.
96
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Spatial and Temporal Variability in Life Histories of Heptageniidae
(Ephemeroptera)
Introduction—
Studies of the life cycles of Heptageniidae (Ephemeroptera) from
various regions of North America have shown that, for many species, certain
characteristics of the life cycle (such as timing and duration of emergence
and hatching or number of generations per year) vary spatially and
temporally. This study describes the life histories of seven species of
Heptageniidae collected from the Wisconsin River and Rocky Run Creek over 3
years and compares them to life histories reported in the literature.
These comparisons reveal patterns in the variability of life-history
strategies and give insight into factors that may influence life cycles.
Study Area—
Heptageniidae were collected (with other aquatic invertebrates) during
6 months of the ice-free season from two sites on the Wisconsin River and
two on Rocky Run Creek. Throughout the study area, the Wisconsin River has
a sandy bottom with scattered, submerged logs and fallen trees. It is
characterized by extensive seasonal floods. In 1974, water levels were high
but steady after spring floods; they fluctuated widely in 1975, and were low
but steady in 1976.
The upstream site in Rocky Run Creek had a substratum of muck,
detritus, and patches of water star grass (Heteranthera dub-ia). The
downstream station in Rocky Run Creek, near the mouth, had a sand bottom
with detritus; Potamogeton sp. and CervLtophyllum sp. were extremely
abundant. During annual spring floods, water from the Wisconsin River mixed
with water from Rocky Run Creek at the downstream sampling site.
Dissolved oxygen values ranged from a low in both streams of 7.5
mg/liter on 30 Aug. 1974, to a high of 12.8 mg/llter and 13.3 mg/liter in
the Wisconsin River and Rocky Run Creek, respectively, on 26 Oct. 1976.
Temperatures in the river ranged from 5.6°C on 26 Oct. 1976 to 26.0°C on 7
Aug. 1975. In the creek the low was 7.3°C on 26 Oct. 1976 and the high was
26.5°C on 7 July 1976. In the Wisconsin River, average midsummer values for
conductivity, total alkalinity, and hardness were 196 mhos/cm, 86, and 132
ppm, respectively. Average values of these parameters in Rocky Run were
slightly more than twice as high. The pH of both streams averaged 7.8.
Materials and Methods—
Samples and physical measurements were taken during the ice-free
seasons following the spring floods in 1974, 1975, and 1976. Four sites
were sampled in 1974 and 1975; sampling was discontinued at the upstream
Wisconsin River station in 1976.
Organisms were collected with basket-type artificial substrate samplers
(Mason et al. 1970). These consisted of 20- x 29-cm chicken barbeque
baskets (or similar wire replicas) filled with 4.5 kg of limestone gravel.
97
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The samplers were suspended from overhanging branches 5 to 10 cm above the
substrate. Three samplers were placed at each station. At monthly
intervals, organisms were removed by shaking the samplers inside an aquatic
D-frame net (1-mm mesh).
In 1976, additional samples were collected from a natural substrate at
the downstream Wisconsin River station. Samplers were constructed from
similar sections (21 x 15 cm) of a silver maple log (Acer saccharinum) from
the Wisconsin River flood plain. The logs were weighted and suspended in a
manner similar to the artificial basket samplers. There were two sets of
three replicates. Every 2 weeks alternate sets were emptied, resulting in
monthly samples from each set. The logs were drawn to the surface inside
an aquatic net. After organisms were removed with a forceps, the logs were
replaced in the water.
All samples were preserved in 70% alcohol in the field. Invertebrates
were sorted, identified, and counted in the laboratory. Head capsule widths
of heptageniid nymphs were measured to the nearest 0.25 mm with an ocular
micrometer. Monthly size-frequency histograms were used to approximate
times of emergence, hatching, and growth. Data from upstream and downstream
sites in the same stream and from artifical and log samplers on the same
dates were combined. There were no differences in life history
interpretations between these locations and substrates.
Results—
Life histories of seven species of Heptageniidae are presented below in
order of increasing number of generations per year (in the study area).
Since no adult collections were made, times of emergence were determined
approximately by the presence of large nymphs and the subsequent appearance
of small nymphs.
Stenonema fuscom (Clemens) occurred only at the upstream site in Rocky
Run Creek and had one generation per year (Figure 30a). Emergence occurred
in late May to early June and young nymphs hatched about 1 month later.
Growth was rapid in summer and by late fall the nymphs were almost full
size.
The low numbers of S. integrum (McDunnough) from the Wisconsin River
made interpretation and comparison of life-history patterns difficult. The
smallest nymphs were found only in June, indicating that hatching occurred
in spring. Also at that time, there was usually a range of size classes
from small to medium (0.5 to 2.5 mm head-capsule width). After July,
occasional individuals were caught through September. S. integrum probably
has one generation per year.
S. exiguion (Traver) had two generations each year in the Wisconsin
River (Figure 30b). One grew quickly during the summer and emerged in
August and September. A longer winter generation followed, growing during
the fall and emerging in the spring. Because times of emergence and
hatching are not known exactly, the two generations could be either multiple
cohorts or truly bivoltine (eggs of one generation are laid by adults of the
98
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1*20%
* • log sampler only
a. Stenonema
fuscom
b. Stenonema
exiguum
(2558)
c. Stenonema
terminafum
(3699)
d. Heptagenia
flavescens
(1047)
e. Stenacron
interpunctatum
MJJASOJJASOOJJASSO
1974 1975 1976
Figure 30. Percentage of nymphs collected at monthly (1974 and 1975) or
twice-monthly (1976) intervals from the Wisconsin River and
Rocky Run Creek. NumSers in parentheses indicate the total
number of nymphs collected during the year.
99
-------
previous generation). Possibly, both strategies are followed.
Size frequency histograms for S. termination (Walsh) from the Wisconsin
River (Figure 30c) suggest two generations per year. As did S. exiguum, 5.
termination usually had a fast summer generation emerging in late summer, and
a longer winter generation with a flight period in spring. In 1975,
emergence appeared at low levels throughout the sampling season, rather than
during more discrete periods as observed in 1974 and 1976.
Heptagenia didbasia (Burks) nymphs were collected only during June in
low numbers from the downstream Wisconsin River station; therefore, their
life history was difficult to interpret. Size classes present in June
ranged from very small to very large nymphs. This suggests that emergence
occurred in late June or early July, with some eggs hatching in spring. The
number of generations present each year in the Wisconsin River could not be
determined.
Data for H. flavescens (Walsh) from the Wisconsin River indicate spring
and fall emergence periods in 1974 and 1975, with nearly continuous summer
recruitment of young (Figure 30d). This suggests at least two successful
generations. In 1976, increased abundance and samples taken twice each
month from log substrates indicate continuous emergence and hatching. The
discrepancy between samplers apparently is an effect of sampling interval
rather than substrate. Data on log substrates, taken from only those dates
when artificial substrates were sampled (1-month intervals), also suggest
two separate emergence periods. Because of low summer populations in 1974
and 1975, multivoltinism could have occurred in all 3 years. The frequent
sampling interval did not affect the life-history interpretations of other
species.
Size-frequency histograms for Stenacron interpunatatwn (Say) from Rocky
Run Creek (Figure 30e) indicate that hatching usually began in late June,
often continuing into the summer at a low level. The presence of a few,
large animals throughout the summer months suggests that emergence was
continuous, with a strong pulse in spring and another weaker pulse in late
August or early September. Corresponding periods of growth appear, one
beginning in late June or early July and extending into the fall, and
possibly another preceeding the August-September emergence. In the study
area, S. interpunctatum was multivoltine with most of the population
hatching in summer, growing over winter, and emerging in spring.
Discussion—
Habitat and Distribution of Heptageniidae—Habitat preferences of the
species found in this study were described by Flowers (1975). Heptagenia
didbasia, Stenonema integrumf and 5. termination are all typically found in
medium to large, deep rivers, especially those with sandy substrates. S.
exigiam is abundant in this habitat, but also thrives in smaller, rocky
streams. S> fuseom and Stenacron interpunctatum are common in a wide range
of lotic habitats; the latter is silt tolerant and is occasionally found on
the rocky shores of lakes.
100
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Lewis (1974) summarized the known distributions for Stenonema
species. Most are widespread in the eastern and central forested regions of
the United States, while S. fuscom is restricted to the Great Lakes area.
Stenacron interpunetatian inhabits the entire eastern half of the United
States and southern Canada. Distributions of some Stenonema species
coincide, while others have little overlap. Distributions of Heptagenia
didbasia and ff. flavescens have not been studied extensively.
Temporal Variability (Annual)— The life-history phenology of species
changes from year to year. Some species (e.g., Stenonema fueoom) vary in
time of hatching by only a few weeks, probably in response to prevailing
environmental conditions such as temperature and photoperiod (Nebeker
1971). Other species are more variable in the timing of life-history
events. For example, 5. termination had a more extended, but low larval
emergence in 1975 than in other years. Perhaps the extreme fluctuations in
water level during 1975 influenced development and emergence of this
species.
5. exi,guum also varies considerably in the amount of time taken for
development following oviposition. Young nymphs sometimes are caught well
before the previous generation reaches full size and sometimes only after
large nymphs have disappeared from the samplers. The possibility that part
of an egg batch is delayed in hatching might explain how some nymphs could
follow a life-history pattern typical of multiple cohorts, while others
could exhibit true bivoltinism. Another possibility would be that delayed
growth follows hatching. As noted, it could not be established definitely
which of the life-history patterns, or both, occurs.
Few reports in the literature examine year-to-year variation in life-
history patterns. McClure and Stewart (1976) suggested that changes could
be made in the life cycle of the mayfly, Chloroterpes mexioanue, in response
to changing environmental conditions, especially where brood overlap
existed. Other authors documenting year-to-year variation in mayfly life
cycles were llinshall (1967) and Brittain (1972). In these studies, a single
generation per year was found and the time and duration of life-history
events did not vary by more than 2 to 4 weeks each year.
Clifford (1970) suggested a mechanism for year-to-year variation when
he found that Leptophlebia sp. in subarctic Canadian streams could emerge
any time after reaching a certain size, though they often continued growing
beyond that size. The effect was to accumulate nymphs capable of emerging
until conditions for emergence were suitable.
Spatial variability (local and geographic)—Life-history data from this
study were compared with those of other studies in Wisconsin (Table 29) and
elsewhere in North America. Shaffer (1975) collected specimens at
approximately monthly intervals from artificial substrate basket samplers in
the Kickapoo River, a small stream in southwestern Wisconsin with a mud
substrate and frequent riffles. Flowers (1975) collected Heptageniidae
monthly, using a hand net, from various streams throughout the state. Some
species collected in these studies differed only slightly in the timing and
duration of life-history events. Others showed different numbers of
101
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TABLE 29. COMPARISON OF LIFE HISTORIES OF HEPTAGENIIDAE FROM WISCONSIN
Species*
Stenonema fueaom
5. integmtm
S> exiguum
K™*
o
NJ
S. termination
Heptagenia didbaaia
H. flaveeaens
Stenacron interpunctatm
Schwarzmeier;
Wisconsin River
S&'cfEfotft
1 univoltine
winter brood
1 univoltine
winter brood
2 generations
(multiple cohorts/bivoltine)
2 generations
(bivoltine)
Spring emergence^
+ hatching;
number of generations unknown
Bivol tine/ multivol tine
Multivoltine
Flowers (1975):
Wiscons^s^reams
1 univoltine
winter brood
1 univoltine
winter brood
1 univoltine
winter brood
1 univoltine
2 generations
(multiple cohorts)
2 generations
Bivoltine
Shaffer (1975)$
Kickapoo River, Wisconsin
Not present
Not present
Bivoltine
Bivoltine
Continuous summer
emergence + hatching;
number of generations unknown
Not present
Multivoltine
+Species are arranged in order of increasing numbers of generations per year.
^Classifications designated by Schwarzmeier for data collected by Scnaffer (1975).
§Year(s) of collection.
^Incomplete data; or limited interpretation possible.
-------
generations per year. For example, life histories of Stenonema exiguum and
5. integrum from Rocky Run Creek and the Wisconsin River are consistent with
results obtained by Flowers (1975). Data for Stenonema exiguum, S.
termination, and Stenacron interpunatatum correspond to Shaffer's (1975)
results, whereas Flowers (1975) found fewer generations per year for these
species. A lack of complete data for Heptagenia diabasia and H. flaveeaens
precludes comparisons.
Even though all these streams are exposed to similar climatic
conditions, local differences could explain the variation in life
histories. For example, streams sampled by Flowers usually had lower
maximum (summer) temperatures than did the Wisconsin River, Rocky Run Creek,
or the Kickapoo River. This suggests that, for some species, warmer
temperature regimes allow faster development and more generations per year.
A comparison of Wisconsin data to heptageniid life histories reported
from southern locations show that two species had more generations per year
than did Wisconsin species. Stenonema integrum has two generations in the
Ohio River basin (Lewis 1974), while S. exiguum is multivoltine in Florida
(Pescador and Peters 1974) and other parts of the southern U.S. (Lewis
1974). Stenaeron interpunatatum is multivoltine in all parts of its range
where its life history has been studied (Peters and Warren 1966, Lewis 1974,
Pescador and Peters 1974). Heptagenia flavescens had one spring emergence
in Arkansas (Peters and Warren 1966), as compared to two or more generations
in Wisconsin.
Other researchers have found that the time and duration of mayfly
emergence varies from stream to stream. For many species, two general
patterns exist. In warmer streams (in southern regions or at low altitudes)
earlier emergence, and sometimes later hatching, may occur with no change in
the number of generations per year (Macan 1957, Pleskot 1961, Maitland 1965,
Minshall 1967), or the emergence period may be extended over a longer time
period compared to emergence in colder streams (Lemkuhl 1968, Clifford
1969). These patterns suggest a possible mechanism for achieving more
generations per year in warmer climates, as is found in many
Heptageniidae. If emergence occurred early enough, additional time would be
available for another generation to hatch and grow. If emergence of several
generations extended over a long enough time period, overlap in generations
could occur. Eventually, in the warmest climates, emergence and hatching
could become nearly continuous (Corbet 1964).
The comparisons made from different streams and in different years in
this study reveal patterns in the variability of life histories of
Ifeptageniidae. Temporal variability occurs to some extent in most
species. Some vary in the time of hatching and emergence by only a few
weeks; others show greatly extended emergence in some years compared to
others. Spatial variability is evident in species having more generations
per year in warmer streams (either locally or in the southern part of their
range). This seems to be further evidence that temperature is an
influential factor regulating heptageniid life cycles. Possibly, these
species also emerge earlier in warmer streams, although no adult collections
were made to test this. Variability in life histories is an advantage to
103
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mayflies because it allows them to survive in a variety of conditions,
whether within a single stream over time or over different local or regional
climatic regimes.
104
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(Ephemeroptera : Heptageniidae). Am. Midland Natur. 78:369-388, 1967.
Nebeker, A.V. Effect of High Winter Water Temperatures on Adult Emergence
of Aquatic Insects. Water Res., 5:777-783, 1971.
Needham, P.R., and R.L. Usinger. Variability in Macrofauna of a Single
Riffle in Prosser Creek, California, as Indicated by the Surber
Sampler. Hilgardia, 24:383-409, 1956.
Neill, W. Notes on the Role of Crayfishes in the Ecology of Reptiles,
Amphibians, and Fishes. Ecology, 32:764-766, 1951.
Nelson, S.G., D.A. Armstrong, A.W. Knight, and H.W. Li. The Effects of
Temperature and Salinity on the Metabolic Rate of Juvenile Macrobraehium
foeenber^fii (Crustacea : Palaemonidae). Comp. Biochem. Physiol. Comp.
Physiol., 56(4):533-538, 1977.
Newell, R.C. Factors Affecting the Respiration of Intertidal Invertebrates.
Am. Zool. 13(2):513-528, 1973.
Odum, E.P. Fundamentals of Ecology. W.B. Saunders Co., Hiiladelphia,
Pennsylvania, 1971. 574 pp.
Pescador, M.L., and W.L. Peters. The Life History and Ecology of Baetisoa
rogerei Berner (Ephemeroptera : Baetiscidae). Bull. Florida State
Museum Biol. Sci., 17(3):151-209, 1974.
Peters, W.L., and L.O. Warren. Seasonal Distribution of Adult Ephemeroptera
in Northwestern Arkansas. J. Kansas Entomol. Soc., 39:386-401, 1966.
Pielow, B.C. 1966. The Measurement of Diversity in Different Types of
Biological Collections. J. Theoret. Biol., 13:131, 1966.
Pleskot, G. Die Periodizitat der Ephemeropteren-Fauna Einiger
Osterreichischer Fliessgewasser. Ver. Int. Ver. Theor. Angew Limnol.,
14:410-416, 1961.
Priegel, G.R., and D.C. Krohn. Characteristics of a Northern Pike Spawning
Population. Tech. Bull. No. 86. Wisconsin Dept. of Natural Resources,
Madison, Wisconsin, 1975. 30 pp.
Prosser, C.L. (ed.). Comparative Animal Physiology. W.B. Saunders Co.,
Philadelphia, Pennsylvania, 1973. 966 pp.
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Rice, R.R., and K.B. Armitage. The Effect of Riotoperiod on Oxygen
Consumption of the Crayfish Orooneatee nais (Faxon). Comp. Biochem.
Physiol., 47(1A):261-270, 1974.
Richardson, R.E. The Bottom Fauna of the Middle Illinois River, 1913-
1925. Bull. Illinois Nat. Hist. Survey, 17:387-475, 1928.
Sather, B.T. 1967. Chromium Absorption and Metabolism by the Crab,
Podophthalnrue virgil. In: Proceedings of the International
Symposium on Radioecological Concentration Processes, Stockholm,
Sweden, pp. 943-976. 1967.
Schiffman, R.H., and P.O. Fromm. Chromium Induced Changes in the Blood
of Rainbow Trout, Salmo gairdneri. Sew. Ind. Wastes, 31:205-211, 1959.
Schoenfield, M.B. Trace Elements in Aquatic Organisms from the Environment
of a Coal Burning Generating Station. M.S. Thesis, University of
Wisconsin, Madison, Wisconsin, 1978. 226 pp.
Schroeder, D.C. Transformations of Chromium in Natural Waters. M.S.
Thesis, University of Wisconsin, Madison, Wisconsin, 1973. 86 pp.
Seidenberg, A.J. Studies on the Biology of Four Species of Fresh-Water
Isopoda (Crustacea, Isopoda, Asellidae) in East-Central Illinois. Ph.D.
Thesis, University of Illinois^ Urbana, Illinois, 1969. 148 pp.
Shaffer, W.S. Potential Effects of a Proposed Impoundment on the
Macroinvertebrate Community of the Kickapoo River with Special Reference
to the Influence of Temperature of the Feeding Rate and Emergence of
Atherix varnegata (Diptera-Rhagionidae). M.S. Thesis, University of
Wisconsin-Madison, Madison, Wisconsin, 1975. 135 pp.
Shannon, C.E., and W. Weaver. 1963. The Mathematical Theory of
Communication. University of Illinois Press, Urbana, Illinois, 1949.
117 pp.
Siegel, S. 1956. Nonparametric Statistics. McGraw-Hill, New York, 1956.
312 pp.
Snedecor, G.W., and W.G. Cochran. Statistican Methods. 6th Ed. Iowa
State University Press, Ames, Iowa, 1967. 593 pp.
Stearns, C.R., B. Bowan, and L. Dzamba. Meteorology. In: Documentation
of Environmental Changes Related to the Columbia Electric
Generating Station. Tenth Semi-Annual Report. Report 82. Inst. for
Environmental Studies, University of Wisconsin-Madison, Madison,
Wisconsin, 1977. pp. 184-194.
Stein, R. A., and J.J. Magnuson. Behavioral Reponse of Crayfish to a Fish
Predator. Ecology, 57:751-761, 1976.
Ill
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Stephenson, D.A., and C.B. Andrews. Hydrogeology. In: Documentation of
Environmental Change Related to the Columbia Electric Generating
Station. Seventh Semi-Annual Report. Inst. for Environmental Studies,
University of Wisconsin-Madison, Madison, Wisconsin, 1976. pp. 66-76.
Swanson Environmental, Inc. Cooling Lake Make-Up Water Intake Monitoring
Program. March 1976-June 1977. (Wisconsin Rawer and Light Co.,
Columbia Energy Center, Portage, Wisconsin) Swanson Environmental, Inc.,
Southfield, Michigan, 1977. 93 pp.
Taylor, E.W., P.J. Butler, and A. Al-Wassia. The Effect of a Decrease in
Salinity on Respiration, Osmoregulation and Activity in the Shore Crab,
Carainue maenas (L.) at Different Acclimation Temperatures. J. Comp.
Physiol. 119(2):155-170, 1977.
Van Olst, J. C., R.F. Ford, J.M. Carlberg, and W.R. Dorband. Use of Thermal
Effluent in Culturing the American Lobster. In: Power Plant Waste Heat
Utilization in Aquaculture—Workshop I, Trenton, N.J. PSE and G Co.,
Newark, New Jersey, pp. 71-97.
Vernberg, F.J., and W.B. Vernberg. Thermal Influence on Invertebrate
Respiration. Chesapeake Science, 10:234-240, 1969.
Vernberg, W.B., P. De Coursey, and W.J. Padgett. Synergistic Effects of
Environmental Variables on Larvae of Uoa pugi.lator> (Bosc.). Mar. Biol.
22:307-312, 1973.
Ward, J.V. Bottom Fauna-Substrate Relationships in a Northern Colorado
Trout Stream: 1945 and 1974. Ecology, 56:1429-1434, 1975.
Water Quality Criteria. A Report of the Committee on Water Quality
Criteria. Environmental Studies Board, National Academy of Sciences and
National Academy of Engineering. EPA-R3-73-033. U.S. Environmental
Protection Agency, Ecological Research Series, Cincinnati, Ohio, 1973.
594 pp.
Weber, C.I. (ed.) Biological Field and Laboratory Methods for Measuring the
Quality of Surface Water and Effluents. EPA 670/4-73-001. U.S.
Environmental Protection Agency, Environmental Research Center,
Cincinnati, Ohio, 1973. 186 pp.
Wiens, A.W., and K. B. Armitage. The Oxygen Consumption of the Crayfish
QTfQoneetes inmunis and Oraoneotes na-is in Response to Temperature and to
Oxygen Saturation. Physiol. Zool., 34:39-54, 1961.
Wilhm, J.L. Range of Diversity Index in Benthic Macroinvertebrate
Populations. J. Water Pollut. Control Fed., 42:221-224, 1970.
Wilhm, J.L. Biological Indicators of Pollution. In: River Ecology,
B.A. Whitton, (ed.). University of California Press, Berkeley and Los
Angeles, California, 1975. 725 pp.
112
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Willard, D.E., B.L. Bedford, W.W. Jones, M.J. Jaeger, and J. Benforado.
Wetlands Ecology. In: Documentation of Environmental Change Related
to the Columbia Electric Generating Station. Eleventh Semi-Annual
Report. Report 92. Inst. for Environmental Studies, University of
Wisconsin-Madison, Madison, Wisconsin, 1977. pp. 94-104.
Wiser, C.W., and D.J. Nelson. Uptake and Elimination of Cobalt-60 by
Crayfish. Am. Midland Natur., 72:181-202, 1964.
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APPENDIX A
REVIEW OF LITERATURE ON ENTRAINMENT FROM COOLING LAKE INTAKE STRUCTURES
The topics of concern in Appendices A, B, and C are: (1) Entrainment
from cooling lake intake structures; (2) acid rain; (3) alternative disposal
of fly ash. Appendix A discusses the possible entrainment damage to fish
and invertebrate populations at the Columbia site; possible damage appears
minimal. In addition, a 1978 study of fish entrainment at the site by
Swanson Environmental Inc. (1977) has revealed that fish loss due to the
current water intake systems is minor. Acid rainfall, the topic of Appendix
B, is not considered a potential problem for aquatic ecosystems at Columbia
because of the high hydrogen ion buffering capacity that results from the
calcareous nature of the drainage basin. The results of Appendix C indicate
that the high pH of the ash expected from Unit II at Columbia will
substantially reduce the pollution potential from a landfill. However, the
landfill site must be chosen carefully to avoid direct connection with the
ground water.
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 via 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; (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) 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-GRl 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 Ilarcy (1973, 1976), Carpenter et al. (1974), Ginn et al. (1974),
King (1974), Davies and Jensen (1975), and Polgar (1975). Several reviews
114
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such as those of Coutant (1970, 1971) and Hillegas (1977) have been
published. Although survival of damaged organisms is often quite low, it
does not appear that the number of organisms lost seriously effects the
aquatic systems.
Biocides such as chlorine are usually used at such low concentrations
that they pose no threat to entrained organisms or the receiving body of
water (Marcy 1971, Bass and Heath 1975, Basch and Truchan 1976, Brungs 1976,
Seegert and Brooks 1977). However, thermal shock, combined with small
amounts of chlorine, has a greater effect than increased temperatures or
chlorine levels alone (Ginn et al. 1974, Eiler and Delfino 1974).
Cooling system designers now use predictive tools to minimize impact.
Curves and models can predict the amount of mechanical damage (Polgar 1975)
and the extent of lethal and sublethal thermal effects (Coutant 1971)
expected for a given intake design. Models have been developed by Goodyear
(1977), Christensen et al. (1977), and others that forecast the effects of
removal on given fish populations.
Potential Effects of Cooling Water Intake at the Columbia Site
The effects of entrainment of aquatic organisms from the Wisconsin
River by the Columbia Generating Station differ from the 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 more applicable than
in "once-through" cooling situations. When 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 as 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, no significant loss of invertebrates from
the Wisconsin River is expected. 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 was available and generating stations diverted up to 30% of
the river flow.
Adult and Juvenile Fish—
A 1-year study of fish entrainment at the Columbia site (Swanson
Environmental Inc. 1978) reported the number, species, length, and
reproductive condition of fish impinged on the temporary screen box unit and
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on the traveling screen unit currently in use. Sampling was conducted for a
continuous 24-h 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 per year (mean ± 90% confidence
limits). The number of adult and juvenile fish impinged at Columbia are low
and even if all impinged fish die, there should be no effect on the river
system.
Fish Eggs and Larvae—
The Swanson Environmental, Inc. (1978) study also included sampling for
fish eggs and larvae. Submersible pumps mounted behind the traveling screen
unit pumped the sample water into 423-y nets. Pump rates were sufficient to
prevent fish from avoiding the sampler. Estimated annual entrainment of
larval fish was 126,659±93,994 larvae per year (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 of 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 to 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 IS^day period (Priegel 1970). By sampling
once every 7 days, the period of walleye larval entrainment could have been
missed. Walleye larvae also may avoid entrainment by staying in the main
currents and bypassing the intake as they enter the Wisconsin River. 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, no measureable effects of entrainment on the
river system are expected. An exception might occur when organism
distribution is patchy near the intake and a significant portion of one
year-class (i.e., 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. Toxicity of Chlorinated Condenser Cooling
Waters to Fish. EPA-600/3-76-009, U.S. Environmental Protection Agency,
Environmental Research Laboratory, Duluth, Minnesota, 1976.
Bass, M.L., and A.G. Heath. Toxicity of Intermittent Chlorine Exposure to
Bluegill Sunfish, Leopomis maerochirus: Interaction with Temperature.
Assoc. Southeast. Biol. Bull., 22:40, 1975.
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Brungs, W.A. Effects of Wastewater and Cooling Water Chlorination on
Aquatic Life. EPA-600/3-76-098, U.S. Environmental Protection Agency,
Environmental Research Laboratory, Duluth, Minnesota, 1976.
Carpenter, E.J., B.B. Peck, and S.J. Anderson. Survival of Copepods Passing
through a Nuclear Power Station on Northeastern Long Island Sound, USA.
Mar. Biol., 24:49-55, 1974.
Christenson, S.W., D.L. DeAngelis, and A.G. Clark. Development of a Stock
Progeny Model for Assessing Power Plant Effects on Fish Populations.
In: Proceedings of the Conference on Assessing the Effects of Power
Plant-induced Mortality on Fish Populations. Pergamon Press, Inc., New
York, 1977.
Coutant, C.C.. Biological Aspects of Thermal Pollution. I. Entrainment
and Discharge Canal Effects. CRC Critical Rev. Environ. Control,
l(3):341-348, 1970.
Coutant, C.C. Effects on Organisms of Entrainment in Cooling Water: Steps
toward Predictability. Nuclear Safety, 12:600-607, 1971.
Davies, R.M., and L.D. Jensen. Zooplankton Entrainment at Three Mid-
Atlantic Power Plants. Water Pollut. Control Fed., 47(8):2130-2142,
1975.
Eiler, H.O., and J.J. Delfino. 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, 1974.
Franklin, D.R., and L.L. Smith, Jr. Early Life History of the Northern
Pike, Esox luai-us L., with Special Reference to the Factors Influencing
the Numerical Strength of Year-Classes. Trans. Am. Fish. Soc.,
92(2):92-110, 1963.
Freeman, R.F., and R.K. Sharma. Survey of Fish Impingement at Power Plants
in the United States. Vol. II. Inland Waters. ANL/ES-56, Argonne
National Laboratory, Argonne, Illinois, 1978. 328 pp.
Ginn, T.C., W.T. Waller, and G.L. Laver. The Effects of Power Plant
Condenser Cooling Water Entrainment of the Amphipod Garmarus sp. Water
Res. 8(ll):973-45, 1974.
Goodyear, C.P. Assessing the Impact of Power Plant Mortality on the
Compensatory Reserve of Fish Populations. In: Proceedings of the
Conference on Assessing the Effects of Power Plant-induced Mortality on
Fish Populations. Pergamon Press, Inc., New York, 1977.
Hillegas, J.M., Jr. Phytoplankton and Zooplankton Entrainrnent. A
Summary of Studies at Power Plants in the United States. Paper
Presented at Savannah River Ecological Laboratory Symposium. Augusta,
Georgia, 1977.
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King, J.R. A Study of Power Plant Entrainment Effects on the Drifting
Macroinvertebrates of the Wabash River. M.S. Thesis, DePauw University,
Greencastle, Indiana, 1974.
Marcy, B.C. Survival of Young Fish in the Discharge Canal of a Nuclear
Power Plant. J. Fish. Res. Board Canada, 28:1057-1060, 1971.
Marcy, B.C. Vulnerability and Survival of Young Connecticut River Fish
Entrained at a Nuclear Power Plant. J. Fish Res. Board Canada,
30(8):1195-1203, 1973.
Marcy, B.C. Planktonic Fish Eggs and Larvae of the Lower Connecticut
River and the Effects of the Connecticut Yankee Plant. In: The Impact
of a Nuclear Power Plant, D. Merriman and L. Thorpe, eds. The
Connecticut River Ecological Study, Monograph No. 1. Am. Fish. Soc.,
Bethesda, Maryland, 1976.
Polgar, T.T. Assessment of Near Field Manifestations of Power Plants. In:
Induced Effects on Zooplankton. Proceedings of the Second Thermal
Ecology Symposium, Augusta, Georgia, 1975.
Priegel, G.R. Reproduction and Early Life History of the Walleye in the
Lake Winnebago Region. Wisconsin Department of Natural Resources Tech.
Bull. 45. Wisconsin Department of Natural Resources, Madison,
Wisconsin, 1978.
Seegert, G.L., and A.S. Brooks. 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, Wisconsin, 1977.
Swanson Environmental, Inc. Cooling Lake Make-Up Water Intake Monitoring
Program. March 1976 - June 1977. Wisconsin Power and light Co.,
Columbia Energy Center, Portage, Wisconsin. Swanson Environmental,
Inc., Southfield, Michigan, 1977.
Whitton, B.A. River Ecology. Studies in Ecology, Vol. 2. University of
California Press, Berkeley and Los Angeles, California, 1975.
Wisconsin Department of Natural Resources. Final Environmental Impact
Statement for the Columbia Generating Station of the Wisconsin Power and
Light Company. Wisconsin Department of Natural Resources, Madison,
Wisconsin, 1973.
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APPENDIX B
REVIEW OF LITERATURE ON ACID RAIN
Recent studies in North America and Europe have documented the
occurrence of rains with a pH ranging from 2.1 to 5.0 (Likens and Bormann
1974, Beamish 1974, Dickson 1975, Beamish 1976, Schofield 1976). Rainwater
is normally slightly acidic (pH 5.7), a result of the equilibrium reaction
between atmospheric carbon dioxide and water that forms carbonic acid
(H^CO^). However, both natural and anthropogenic processes can add three
strong mineral acids—sulfuric (^SO^), nitric (HNO-j), and hydrochloric
(HC1)—to atmospheric water with a resulting sharp decrease in pH (Gorham
1976). The most predominant of these acids is ^SO^ which can be formed in
substantial amounts from the sulfur dioxide (802) 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
HNOo and HC1, which also are 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 ice and snow runoff.
The work of Cogbill and Likens (1974) illustrates that acid rain 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 1956-66.
The initial effects of acid input into lakes and streams depends
largely on edaphic characteristics that determine buffering capacity. All
waters affected by acid rain are in areas that are geologically highly
resistant to chemical weathering and usually have a low concentration of
major ions—particularly biocarbonate (HCOo)—resulting in a specific
conductance less than 50 mhos/cm (Wright and Gjessing 1976). Acid rainfall
into weakly buffered systems causes the bicarbonate ion to be lost and then
replaced by sulfate. Hence, sulfate is the major anion in acid soft water,
whereas bicarbonate predominates in non-acid soft water. Acid lakes
frequently contain elevated aluminum and manganese concentrations that are
attributed to dissolution from surrounding soils. Elevated levels of other
heavy metals (Pb, Zn, Cu, and Ni) 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 population, since the loss of an exploitable fish
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population is the most noticeable and economically important consequence of
acid rain. Fish loss is reported to be a gradual process, resulting not
from acutely lethal pH changes, but from the failure to recruit new year
classes into the population (Beamish 1974). At pH values above the lower
lethal level, laboratory and field studies have demonstrated interference
with spawning (Mount 1973, Beamish 1976). The presumed mechanism
responsible for 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 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, (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).
Selective breeding of acid-tolerant fish strains has been proposed as a
means of stocking waters that have lost their natural populations. However,
the observed rates of population extinction indicate that acidification has
been too rapid for natural selection processes to effectively maintain fish
populations under natural conditions.
The effects of acid rain on aquatic organisms such as microdecomposers,
primary producers, zooplankton, and zoobenthos are less conspicuous, but are
equally as serious as damage to fish. 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. 1976).
Rooted macrophytes, zooplankton, and benthic invertebrates also are stressed
by acidification of waters (Hendrey et al. 1976). Table B-l summarizes some
of the effects of pH on aquatic organisms.
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TABLE B-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.
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 are
restricted;
Some caddisflies and dragonflies 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 non-existent;
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.
European Inland
Fisheries Advisory
Committee (1969);
Lackey (1938)
Hendrey et al.
1976a
U.S. Environmental
Protection Agency
1973
U.S. Environmental
Protection Agency
1973
U.S. Environmental
Protection Agency
1973
Hendrey et al. 1976a
Beamish 1974, 1975
Beamish 1975
Continued
121
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Table B-l. Continued
pH
Effect
Reference
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 vutelus) fail
to hatch.
6.0-6.5 Unlikely to be harmful to fish unless
free C02 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.JD is near the lower limit for
Gammarus reproduction;
Microorganisms and plants are normal;
however, toxicity of other substances
may be affected by pH shifts within
this range.
U.S. Environmental
Protection Agency
1973
Mount 1973
Beamish 1976
Milbrink et al. 1975
U.S. Environmental
Protection Agency
1973
U.S. Environmental
Protection Agency
1973
122
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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., and conductivities of 178 to 273 mhos/cm; (2) winds are
predominately from the west and south (Stearns et al. 1977), therefore power
plant emissions should miss most of the nearby aquatic systems that are
mainly west and south of the plant; (3) the current pH 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 operating in 1975.
The effect of additional sulfur emissions when Columbia II begins
operation should be considered. Also to be accounted for are 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. Loss of Fish Populations from Unexploited Remote Lakes in
Ontario, Canada as a Consequence of Atmospheric Fallout of Acid. Water
Res., 8:85-95, 1974.
Beamish, R.J. Long Term Acidification of a Lake and Resulting Effects on
Fishes. Ambio, 4(2):98-102, 1975.
Beamish, R.J. Acidification of Lakes in Canada by Acid Precipitation and
the Resulting Effect on Fishes. Water Air Soil Pollut. 6:501-514,
1976.
Cogbill, C.V., and G.E. Likens. Acid Precipitation in the Northeastern
United States. Water Resour. Res., 10(6):1133-1137, 1974.
Dickson, W. The Acidification of Swedish Lakes. Report No. 54. Inst.
of Freshwater Research, Drottningholm, Sweden, 1975. pp. 8-20.
European Inland Fisheries Advisory Committee Working Party on Water Quality.
Water Quality Criteria for European Freshwater Fish: Extreme pH Values
and Inland Fisheries. Water Res., 3:593-611, 1969.
Gjedrem, T. Genetic Variation in Tolerance of Brown Trout to Acid Water.
SNSF-Project FR5/76, Norway, 1976. 11 pp.
Gorham, E. Acid Precipitation and Its Influence upon Aquatic Ecosystems:
An Overview. Water Air Soil Pollut., 6:457-481, 1976.
Hendrey, G.R., K. Baalsrud, T.S. Traaen, M. Laake, and G. Raddum. Acid
Precipitation: Some Hydrobiological Changes. Ambio, 5(5-6):224-227,
1976a.
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Hendrey, G.R., R. Borgstrom, and G. Raddum. 1976b. Acid Precipitation in
Norway: Effects on Benthic Faunal Communities. Presented at the 39th
Annual Meeting, Am. Soc. Limnology and Oceanography, Savannah, Georgia,
1976b.
Lackey, J.B. The Flora and Fauna of Surface Waters Polluted by Acid Mine
Drainage. Public Health Rep. 53:1499-1507, 1938.
Likens, G.E., and F.H. Bormann. Acid Rain: A Serious Regional
Environmental Problem. Science, 184:1176-1179, 1974.
Milbrink, G. , and N. Johansson. Some Effects of Acidification on Roe of
Roach, Rut-ilue ruti-lus L., and Perch, Pemz fiuoiatilis L., with
Special Reference to the Avad System in Eastern Sweden. Report No.
54. Inst. of Freshwater Research, Drottningholm, Sweden, 1975.
Mount, D.I. 1973. Chronic Effect of Low pH on Fathead Minnow Survival,
Growth and Reproduction. Water Res., 7:987-993, 1973.
Robinson, G.D. , W.A. Dunson, J.E. Wright, and G.E. Mamolito. Differences in
Low pH Tolerance among Strains of Brook Trout (Salvelinus
fontinalis). J. Fish Biol., 8:5-17, 1976.
Schofield, C.L. Acid Precipitation: Effects on Fish. Ambio, 5(5-6):228-
230, 1976.
Stearns, C.R., B. Bowen, and L. Dzamba. Meteorology. In: Documentation of
Environmental Change Related to the Columbia Electric Generating
Station. Report 82, Tenth Semi-Annual Progress Report. Inst. for
Environmental Studies, University of Wisconsin -Madison, Madison,
Wisconsin, 1977. pp. 171-183.
U.S. Environmental Protection Agency. 1973. Acidity, Alkalinity, and pH.
Water Quality Criteria, Ecological Research Series, R3-73-033. U.S.
Environmental Protection Agency, 1973. pp. 140-141.
Wright, R.F., and E.T. Gjessing. Acid Precipitation: Changes in the
Chemical Composition of Lakes. Ambio, 5(5-6) :219-223, 1976.
Van Loon, J.C., and R.J. Beamish. Heavy Metal Contamination by Atmospheric
Fallout of Several Flin Flon Area lakes, and the Relation to Fish
Populations. J. Fish Res. Board Can., 34:899-906, 1977.
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APPENDIX C
REVIEW OF LITERATURE ON ALTERNATIVE DISPOSAL OF FLY ASH
The increased national emphasis on the use of coal to meet energy
requirements may double 1975 coal ash production levels by the year 1995
(PEDCo-Environmental, Inc. 1976). Annual coal ash production is currently
estimated to be 61.9 x 106 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 such as cement, asphalt and concrete, fertilizer, fire control,
road bed stabilizer, soil aeration, and sanitary landfill cover PEDCo-
Environmental 1976, Theis 1976a, Harriger 1977). There is continuing
research 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 the
precipitation of phosphorus from natural waters and the ash seals the
nutrient in the sediment; however, the side effects of such treatment may be
severe (Theis and DePinto 1976). Fly ash concentrations of 10 to 20 g/liter
were toxic to Stone Lake (Michigan) 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 1976); however, the high conductivities of fly
ash-water solutions may increase salt concentrations to injurious levels for
many sensitive crops (Olsen and Warren 1976). Theis (1976a) suggests the
extraction of quantities of rare metals from ash for industrial re-use; for
example, a generating station producing 260 tons of ash/day could provide
approximately 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.
Presently, more fly ash is produced than is demanded by commercial
users (Theis 1976a). The average rate of ash production is 0.5 kg/kWh
(PedCo-Environmental, Inc. 1976) and increased coal use will increase the
amount of ash to be disposed of. The New Source Performance Standards
(NSPS) applicable to new power plants prohibit discharges from ash settling
ponds to enter natural waters (Dvorak and Pentecost 1977). To comply with
these regulations, ash from Unit II of the Columbia Generating Station is
being held in a segregated portion of the ash basin until a site for
permanent land disposal is found.
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 the different nature of the
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contaminants is considered, 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 the disposal method is
recent, there is little knowledge of the long-term effects of such
disposal. Studies are needed for the creation of standards for land
disposal of toxic substances, which is virtually unregulated at the federal
level (Fields and Lindsey 1975).
The potential for ground-water contamination by leachate produced when
water percolates through a coal ash landfill draws the most widespread
concern. High salt concentrations in leachate may be a significant problem,
especially if the leachate reaches ground-water supplies that are already
high in salt. Increased pH due to ash leachate may be a localized problem
(Olson and Warren 1976); however, pH is important because it affects metal
solubilities and adsorption. 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 that many desirable
ions replace undesirable ions in an ion exchange process. He concluded that
most ground-water contamination is limited to the immediate vicinity of the
landfill because of the slow movement of the groundwater. 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 sanitary landfill results to a landfill
designed for fly ash. Theis (1976b) and Tneis and Marley (1976) discuss the
potential for ground-water contamination from land disposal of fly ash.
They determined that the important characteristics of ash are 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 are released from leachate in insignificant amounts or are
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rapidly sorbed onto soil particles; however, As, Ni, and Se 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 the combustion of
eastern U.S. 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 were 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. Riblic Health
Service drinking water recommendations. Landfill wells also exhibited
higher concentrations of Zn, Ca, Cr, Cu, Mg, and K than the off-site wells;
however, Ca, Cr, and Cu were fairly low 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. Calcium, Cd,
Cu, Fe, Mg, Na, Se, Zn, and SO^ levels 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. Metal concentrations were higher 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 low concentrations of Cr, Cu, and Zn in the ground water
evidenced attenuation by clay and restricted metal mobility in ground
water. This indicates the need to contain surface runoff 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 increases at lower pHs (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 pHs of 6, 9, and
12 (Theis and Wirth 1977). Iron and Mn precipitate at pH > 7.5 (Harriger
1977). Fields and Lindsey (1975) conclude that low pH affects ion exchange
and that 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
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to 8.0. Frost and Griffin (1977) found, however, that 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 statiow. The greatest environmental concern with low pH ashes
is the large amount of surface leachable Fe (Theis and Wirth 1977). Theis
(1976a) states that a greater amount of metal probably will be released from
ash into ground water than into surface water. This is because of the lower
pH and high C0? 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 the threat of ground and surface water contamination as much as
possible. 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 (Zaporozec 1974, Fields and Lindsey 1975); however, 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 may have the
capacity to seal themselves against leachate loss. As soluble CaO moves
into the soil and forms CaCOo, the permeability of the soil may be
significantly reduced (Olsen and Warren 1976). Fly ash is often
deliberately applied to sanitary landfills because of its moisture absorbing
characteristics (PEDCo-Environmental, Inc. 1976).
Another suggestion to reduce the potential of contamination is
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; use of sites whose soils or intervening materials have high exchange
and adsorption capacities; use of leachate lagoons to prevent surface-water
contamination, use of sites where the water table is much below the bottom
of the waste; avoiding areas subject to flooding (Waldrip and Rune 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
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concerning site location, investigation, monitoring, and management for
sanitary landfills. Much of the information in both reports can be applied
to coal ash landfills.
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-year (or more frequent) flood-frequency
class. A study for the U.S. Environmental Protection Agency Battelle
Memorial Institute (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 expected from Columbia II will
substantially reduce the pollution potential from a landfill. However, the
landfill site must be chosen carefully 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
probably will continue to exceed demand (PEDCo-Environmental 1976, Theis
1976a, 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, Harriger 1976, Theis 1976b, Theis and
Marley 1976).
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5. Because of these mechanisms, and due to 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).
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. Metal solubilities are reduced and clay properties are
improved under these conditions (Fields and Lindsey 1975, Theis 1976a,
Harriger 1977). Fortunately, the western U.S. 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 1976, Dvorak and
Pentecost 1977). Fly ash appears to have some capacity to form a seal
itself (Olsen 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
Battelle Memorial Institute. Program for the Management of Hazardous
Wastes. Final Report for the U.S. Environmental Protection Agency.
Office of Solid Waste Management Programs, Richland, Washington, 1973.
385 pp.
California State Water Resources Control Board. Disposal Site Design and
Operation Information. Sacramento, California, 1975. pp. 19-21.
Davis, J.E., and J.H. Faber. Annual Report: National Ash Association.
National Ash Association, Washington, D.C., 1977.
Dvorak, A.J., and E.D. Pentecost. Assessment of the Health and
Environmental Effects of Power Generation in the Midwest. Vol. II.
Ecological Effects. Draft. Argonne National Laboratory, Argonne,
Illinois, 1977. 169 pp. (Permission obtained.)
Fields, T., and A.W. Lindsey. Landfill Disposal of Hazardous Wastes: A
Review of Literature and Known Approaches. EPA/530/SW-165, U.S.
Environmental Protection Agency, Cincinnati, Ohio, 1975. 36 pp.
Frost, R.R., and R.A. Griffin. Effect of pH on Adsorption of Arsenic and
Selenium from Landfill Leachate by Clay Minerals. Soil Sci. Soc. Am.
J., 41:53-57, 1977.
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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. Pakkala, and D.J. Lisk. National
Survey of Element and Radioactivity in Fly Ashes. Environ. Sci.
Technol. 11:1194-1201, 1977.
Griffin, R.A., K. Cartwright, N.F. Shimi, J.D. Steele, R.R. Ruch, W.A.
White, G.M. Hughe, and R.H. Gilkeson. Attenuation of Pollutants in
Municipal Landfill Leachate by Clay Minerals. Part 1: Column Leaching
and Field Verification. Environmental Geology Notes, No. 78, November
1976. Illinois State Geological Survey, Urbana, Illinois, 1976. 34 pp.
Harriger, T.L. Impact on Water Quality by a Coal Ash Landfill in North
Central Chautaqua County, New York. Ph.D. Thesis, State Univerity
College, Fredonia, New York, 1977. 192 pp.
Harriger, T.L. , W.M. Benard, D.R. Corbin, and D.A. Watroba. 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, 1977. (Abtracts).
Heidman, J.A., and D.R. Brunner. Solid Waste and Water Quality. J. Water
Pollut. Control Assoc., 48:1299, 1976.
Olsen, R.A., and G. Warren. Aquatic Pollution Potential of Fly Ash
Particles. In: Toxic Effects on the Biota from Coal and Oil Shale
Development. Natural Resources Ecology Laboratory, Colorado State
University, Internal Project Report No. 7, Ft. Collins, Colorado, 1976.
pp. 91-112.
PEDCo-Environmental, Inc. Residual Waste Best Management Practices: A
Water Planner's Guide to Land Disposal. EPA/440/9-76/022, U.S.
Environmental Protection Agency, Cincinnati, Ohio, 1976.
Suarez, D.L. Heavy Metals in Waters and Soils Associated with Several
Pennsylvania Landfills. Ph.D. Thesis. Pennsylvania State University,
University Park, Pennsylvania, 1974. 222 pp.
Theis, T.L. Potential Trace Metal Contamination of Water Resources through
Disposal of Fly Ash. Notre Dame University, CONF-750530-3, South Bend,
Indiana, 1976a. 21 pp.
Theis, T.L. Contamination of Ground Water by Heavy Metals from the Land
Disposal of Fly Ash. Technical Progress Report. 1 June 1976 to 31
August 1976. Prepared for U.S. Energy Research and Development
Administration. Notre Dame University, South Bend, Indiana, 1976b. 44
pp.
Theis, T.L., and J.V. DePinto. Studies on the Reclamation of Stone Lake,
Michigan. EPA-600/3-76-106, U.S. Environmental Protection Agency,
Ecological Research Series, Cincinnati, Ohio, 1976. 84 pp.
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Theis, T.L. , and and J.J. Marley. Contamination of Ground Water by Heavy
Metals from the Land Disposal of Fly Ash. Technical Progress Report. J
June 1976 to 29 February 1976. Prepared for U.S. Energy Research and
Development Administration, Notre Dame University, South Bend, Indiana,
1976. 21 pp.
Theis, T.L., and J.L. Wirth. Sorptive Behavior of Trace Metals on Fly Ash
in Aqueous Systems. Environ. Sci. Techno1., 11:1096-1100, 1977.
Waldrip, D.B. 1975. The Effect of Sanitary Landfills on Water Quality in
Southern Indiana. Ph.D. Thesis, Indiana University, Bloomington,
Indiana, 1975. 160 pp.
Waldrip, D.B., and R.V. Ruhe. Solid Waste Disposal by Land Burial in
Southern Indiana. Water Resources Research Center, Technical Report No.
45. Purdue University, West Lafayette, Indiana, 1974. 110 pp.
Zaporozec, A. Hydrogeologic Evaluation of Solid Waste Disposal in South
Central Wisconsin. Wisconsin Department of Natural Resources, Tech.
Bull. No. 78, Madison, Wisconsin, 1974. 31 pp.
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APPENDIX D
LITERATURE REVIEW: THE DYNAMICS AND EFFECTS OF
CHROMIUM AND OTHER METALS IN ORGANISMS
Many metals are essential components of living organisms in trace
amounts. Metals such as ZN, Fe, and Cu are necessary constituents of many
enzymes and pigments (Prosser 1973). There is recent evidence that annimals
require chromium in their diets for normal glucose metabolism and that
dietary Se may offer protection against chemical carcinogens, methylated
mercury, and Ca (Allaway 1975). It is when concentrations exceed the
necessary or beneficial levels that organisms may be adversely affected by
metals; this may be occurring in the stream system receiving coal ash
effluent with its elevated metal concentrations.
Recently, there has been increased concern over the effects of elevated
environmental metal levels on individual organisms and, particularly, over
the ramifications of increased metals in food chain relationships and
bioaccumulation by higher trophic levels. Bioaccumulation is well known in
some metals, such as Hg and Cd. Both metal-susceptible and metal-tolerant
organisms may be hazards to their consumers in higher trophic levels. Kania
and O'Hara (1974) found that mosquitofish (Gambusia) exposed to sublethal
concentrations of Hg were less able to escape predation by bass. The
raosquitofish with the highest metal concentrations would be the most heavily
consumed, leading to increased metals in the food chain. Highly resistant
organisms could accumulate very high metal concentrations before being
preyed upon or dying and being consumed by detritivores; crayfish may be
such a hazard. Gillespie et al. (1977) determined that Orconeetes
propinquus is highly resistant to Cd and could contribute significant
amounts to the next trophic level. Crayfish, fed on by many species of
fish, amphibians, reptiles, birds, and mammals are important in food webs
(Neill 1951). Davis and Foster (1958) report that food chains tend to
select for essential elements. This is of limited usefulness, however,
since the majority of common metals are essential in small amounts and toxic
in larger concentrations. The threat of bioaccumulation may not be as great
for chromium as for other metals, however, since Sather (1966) found that
the lower trophic levels (algae, sponges, and snails) concentrate more
chromium than fish and crayfish.
Chromium toxicity and dynamics are highly dependent on chemical form
and oxidation state. These parameters govern the behavior of chromium in
19
the Columbia ash drainage system. Chromium has four oxidation states: Gr-
and Cr are rare in nature; Cr and Cr are most common. Hexavalent
chromium salts are very water soluble (up to several g/liter) while most
trivalent salts (including the very common hydrous oxides) are insoluble and
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thus exist in very low concentrations in the pH of natural waters (Foster
1963, Schroeder 1973). Heat, organic matter, and chemical reducing agents
can reduce Cr+6 to Cr+3 (Foster 1963). The hexavalent form is widely
recognized to be highly toxic, with the trivalent form moderately toxic to
not toxic (Foster 1963, Mathis and Cummings 1973, Allaway 1975). Allaway
(1975) found no reports of toxicity for dietary Cr . Chromium entering
natural waters is usually in hexavalent form, but is rapidly reduced to Cr
precipitated, and sorbed by the sediments. This reduction is usually due to
organic matter or Fe+2 (with the Cr+3 then sorbed by the Fe(OH)3
precipitate) (Schroeder 1973). These reactions are occurring in the ash pit
drain because the coal ash effluent has a high Fe content. Thus, most
chromium moves to the sediment in the trivalent form; whatever remains
dissolved in the water is most likely Cr+ . Hexavalent chromium is probably
reduced to Cr+^ in living organisms, but there is no way to test this
hypothesis since chromium cannot be extracted without affecting its
oxidation state (Schroeder 1973). Huffman and Allaway (1973) indicate that
Cr+6 is changed to Cr+3 in the stomachs of rats, and it is not easily
absorbed by the intestine at neutral pHs.
There is disagreement over the relative importance of food and water in
the uptake of metals by aquatic organisms. However, the mechanisms of
uptake, transport, elimination, and regulation are becoming lucid. Davis
and Foster (1958) report that although absorption and adsorption from water
are important in the bioaccumulation of radioisotopes, the food chain is the
most important factor. For animals that accumulate substances by ingestion,
the concentration in the body will fluctuate with metabolic rate. Bentreath
(1973) determined that the direct accumulation, of zinc-65 and manganese-54
from water was small in comparison with dietary uptake by the plaice
(Pleu-poneates plateesa)* Freshwater animals appear relatively impermeable
to Zn, thus all Zn is normally obtained from food and eliminated in the
feces with the hepatopancreas regulating uptake, elimination, and transport
(Bryan 1966, 1967). Bryan also found that this results in high Zn
concentrations in crayfish hepatopancreas and stomach fluids.
Concentrations remained constant in muscle, even with high amounts in the
water, indicating that the absorbed Zn is probably returned to the blood.
Food is also more important than water as a route for zinc uptake in marine
crabs (Bryan 1966).
Bryan and Hummerstone (1971, 1973) found that Nereis accumulates Cu,
zinc, and cadmium from water and ingested sediment; water is the primary
route for Zn and Cd. Copper accumulation is probably unregulated since body
concentrations are related to the concentration in the sediment. Tissue Zn
concentration remains constant despite environmental concentration,
indicating some degree of regulation. Odum (1961) reported that arthropods
acquire Zn from water or food and that there are two pools in the body: 1)
Unassimilated and rapidly lost with concentration dependent on the aqueous
environment; 2) assimilated and excreted slowly at a rate proportional to
134
-------
metabolic rate. Fowler et al. (1970) tested Zn uptake in euphausiids and
determined that accumulation occurred in similar tissues regardless of mode
of uptake, except that none appeared in the exoskeleton after ingestion.
The areas of localization were therefore independent of mode of uptake, but
the quantities were different.
The crustacean exoskeleton is relatively impervious to many ions in
solution (Wiser and Nelson 1964). Cobalt accumulation was greater in small
crayfish than in larger crayfish per gram of body weight because they have a
relatively greater proportion of surface area on which adsorption can
occur. The integument had the highest cobalt concentrations, followed by
the hepatopancreas. Metal elimination occurred at a slower rate than
uptake. Metal-tolerant isopods (Asellus meridianus) accumulated Cu and Pb
from food and water (Brown 1977), but there was no evidence that non-
tolerant Aseltus accumulated the metals from food. The non-tolerant animals
also did not survive the exposure and had much smaller proportions of Cu in
the hepatopancreas than did Cu-tolerant animals. On this basis, Brown
proposes two possible tolerance mechanisms: Improved metal storage;
improved metal detoxification.
Little work has been done to determine the importance of ingestion as a
mechanism of chromium uptake. Uptake of Cr+" did not occur even when it was
placed directly in the stomachs of rainbow trout (Salmo gairdneri) (Knoll
and Fromm 1960). The gills were the primary site of uptake from water due
to differences in concentrations across the membranes. Blood maintained a
concentration similar to that of the water, while all tissues except muscle
exceeded environmental concentrations. In uncontaminated water, chromium
elimination was rapid from all tissues except spleen.
10
Rats absorbed Cr poorly in the intestine because of its low
solubility at neutral pHs (Huffman and Allaway 1973). Fasted rats absorbed
6% of the Cr they ingested. Acid conditions in their stomachs caused Cr
i O
to be changed to Cr . Tissue uptake was greatest in the liver, kidney, and
blood (MacKenzie et al. 1959).
A marine polychaete (Hermione hystrix) placed in sea water containing
Cr Cl- exhibited tissue accumulation only on the body surface and in the
digestive tract. When exposed to Cr 0^, however, there was a small amount
of passive tissue accumulation, which depended on water concentration
(Chipman 1966). Which uptake route was most important was not evidenced.
Chromium uptake from water apparently occurs through the gills and is
transported by the blood. This is reported for lobsters (Van Olst et al.
1976), which apparently regulate uptake of essential and non-essential
metals, for crabs (Sather 1966), where the gills regulate chromium
absorption dependent on oxidative phosphorylation and carbonic anhydrase
action, and for largemouth bass, Mier>opter>us sailmoi-des (Fromm and Schiffman
1958). Elimination occurs via the gills in lobsters (Van Olst et al. 1976)
and partially via the liver in fish (Fromm and Schiffman 1958).
135
-------
The effects of many heavy metals on organisms are well documented
(Becker and Thatcher 1973, Eisler 1973, Eisler and Wapner 1975.) Most
laboratory documentation has been concerned with lethal effects and acute
toxic limits (Table D-l summarizes these for chromium), but there are some
studies of sublethal effects (Table D-2 for chromium). Sublethal effects
are varied and depend on the specific metals and organisms involved. Copper
retards growth and development and damages tissue in crayfish at levels as
low as 0.06 mg/liter (Hubschman 1967a, 1967b). Mercury affects the
metabolic and swimming rates of larval crabs (DeCoursey and Vernverg
1972). Chromium irritates and causes pathological changes in the digestive
tract a well as reducing oxygen consumption and possibly acting as a protein
coagulant (Fromm and Schiffman 1958, Cheremisinoff and Habib 1972). An in
vitro study (Buhler et al. 1977) indicated that trout enzymes are fairly
insensitive to Cr+6 inhibition, but they, as well as Kuhnert et al. (1976),
found significant enzyme reductions in several rainbow trout (Salmo
gaerdneri) tissues upon in vivo exposure. Chromium appears to be different
from most metals because it does not bind to the gill epithelia and
mechanically interfere with respiration (Fromm and Schiffman 1958, Buhler et
al. 1977).
Many factors affect the degree of toxicity and some of these factors
pertain to the organism itself. Raymont and Shields (1963) suggest that
resistance is probably attributed to permeability of the gut and body wall,
composition of body tissue, rates of excretion, and size. Adaptation to Zn
by Nereis is probably a result of reduced body surface permeability and an
increased ability to excrete Zn, while Cu tolerance appears to result from a
complexing system which detoxifies and stores Cu in the epidermis and
nephridia (Bryan and Hummerstone 1971, 1973). Juvenile organisms are
usually more susceptible than mature individuals (Hubschman 1967b, Doyle et
al. 1976, Van Olst et al. 1976). The relative tolerance of fish and
invertebrates to metals is controversial. Ma this and Cummings (1973) state
that fish are less affected by metals, while Warnick and Bell (1969)
conclude that fish are more susceptible. Many environmental factors affect
toxicity—water hardness, temperature, and osmotic concentration of the
medium (Bryan and Hummerstone 1971, 1973, Zitko and Carson 1976). Bryan
(1971) presents the following table to illustrate the variety of factors
affecting the toxicity of metals to aquatic organisms:
136
-------
Form of metal
in water
Presence of
other metals
or poisons
Soluble
Particulate
Ion
Complex
Chelate
Compound
[Precipitate
Adsorbed
Antagonistic effects
Additive effects
Synergistic effects
Factors influencing physiology
of organism and perhaps form
of metal in water
Salinity
Temperature
Dissolved oxygen
pH
Light?
Condition of
the organism
Stage of life history
Changes in life cycle (e.g., molt)
Size
Activity
Acclimation to metals
All of these factors may be operating in the ash effluent disposal system of
the Columbia Generating Station, affecting not only toxicities but sublethal
responses of organisms as well.
137
-------
TABLE D-l. SUMMARY OF WORK DONE TO DETERMINE CONCENTRATIONS
OF CHROMIUM THAT ARE LETHAL TO ORGANISMS
Form
Cr+3: CrCl3 6H20
Cr • Na CrO
Cr+6: K2Cr207
Cr+3 K-chromic
Cr+6
Cr+6: K2Cr207
Cr+6
Cr+3
Concentration
2.0 mg/liter
0.21 mg/liter
0.7 mg/liter
42 mg/liter
1.0 mg/liter
0.6-0.7
mg/liter
280 mg/liter
3.5 mg/liter
Various
12.1, 9.3
Effect
3-week LC-50
100 h TLm
2-day toxic
threshold
2 -day toxic
threshold
3-week toxic
toxic thresh-
old for longer
tests
48-h TLm
Various
24 h Tlm>
96 h TLm
(mg/liter)
Organism
Daphnia magna
Daphnia magna
Daphnia magna
Nereis
Hydropsyche
larvae
Stenonema
rubrim larvae
Homarus
amensanue
Kais
Other
Hardness: 45
rag/liter
Adding several
Na compounds
prolonged
survival
Marine
Soft water
Larvae more
sensitive
than juveniles
and adults
Reference
Biesinger and
Christensen
(1972)
Dowden and
Bennett
(1965)
Bringmann
and Kuhn
(1959)
Raymont and
Shields (1963)
Roback (1965)
Van Olst et al
(1976)
Rehwoldt et al
(1973)
6.4, 3.2
58, 50
46, 43.1
16.5, 11.0
15.2, 12.4
10.2, 8.4
Gaxmaru.6
caddisfly
damselfly
Chironomus sp.
Armiaola sp. eggs
Arm-iaola sp. adults
Continued
138
-------
TABLE D-l. Continued
Form
Cr+6
Cr"1"6 Na2Cr20?
Cr+6
Cr ?
Cr"1"6: K2Cr,07
K2Cr64
CrKSOA
Cr"1"6: K2Cr207
Concentration
5.0 mg/liter
10-12.5
mg/liter
0.08 mg/liter
195 mg/liter
£ 20 mg/liter
1.0 mg/liter
113 mg/liter
170 mg/liter
5.07 mg/liter
67.4 mg. liter
7.46 mg/liter
71.9 mg/liter
4.10 mg/liter
3.33 mg/liter
17.6 mg/liter
27.3 mg/liter
118.0 mg/liter
133.0 mg/liter
Effect
40% kill-
15 days
80% kill-
15 days
Significant
mortality
48 h TLm
Not lethal
Acute toxic
limit
96 h TI^
96 h TV
Soft water
Hard water
Soft water
Hard water
Soft water
Soft water
Soft water
Hard water
Soft water
Hard water
Organism Other
Salmo gairdner*i
Chinook salmon Hardness:
and Salmo 70 mg/liter
gairdneri
Microptemts
salmoidee
Gasteroeteue
aauleatus
Lepomis
machroehimiB
Pimephales promelas
Lepomis maorochims
Caraesiue aumtue
Lebistes reticulatus
Pimephalee promelas
Lepomis maerochirus
Reference
Fromm and
Stokes (1962)
Olson and
Foster (1956)
Fromm and
Schiffman
(1958)
Hawksley
(1967)
Trama and
Benoit (1960)
Pickering and
Henderson
(1965)
139
-------
TABLE D-2. SUMMARY OF WORK DONE TO DETERMINE THE SUBLETHAL EFFECTS OF CHROMIUM ON ORGANISMS
Form
Cr ?
Cr+6: Cr04
Cr+6
Concentration
0.32-1.6 mg/liter
6.A-16.0 mg/liter
1.39 mg/liter
0.139 mg. liter
5.0 mg/liter
Effect
56 days, inhibits
algal growth
56 days, inhibits
algal growth
Drastic reduction
in production
Slight (but sig-
nificant) decrease
in production
50% reduction in
photosyntheses
(A days)
Organism
Lepooinelie
eteinii
Chlorella
variegatue
Algae
Maarooyetia
pyrifera
Other Reference
Hervey (19A9)
Freshwater Carton (1972)
Salt water Clendenning
and North
(1960)
Cr ?
Cr
+6
Cr
.+3
10 mg/liter
100 mg/liter
12.5 mg/liter
17.5 mg/liter
100 mg/liter
Decreased respir-
ation and activi-
ty due to reduced
microblal repro-
duction
10% reduction in
BOD in 1 day
50-90% BOD
reduction
10% BOD reduction
50% BOD reduction
90% BOD reduction
Sewage sludge
Sewage sludge
Ingols and
Fetner (1961)
Fieukelekian and
Gillman (1955)
Continued
-------
TABLE D-2. Continued
Form
Cr«:
CrCl36H20
Cr+6
Cr+6:
Cr+6
Cr+6:
Cr+6: Cr04
Cr+6
Concentration
0.6 mg/liter
0.33 mg/liter
0.0125 mg/liter
0.2-0.4 mg/liter
2-4 mg/liter
X).016 mg/liter
XJ.013 mg/Uter
31.0 mg/liter
sublethal
Effect
50% reproductive
impairment
16% reproductive
impairment (maximum
safe concentration)
48% fewer off-
spring
Maximum safe
concentration
Changes in blood,
internal, or
intracellular
effects
Retarded growth
rates
Retarded growth
rates
No kill in 96 h
Reduced 02 con-
sumption; path-
Organism Other
Daphnia magna Hardness :
45 mg/
liter
Neanthee
arenaoeodentata
Salvelinue Hardness:
fontinalie 45 mg/
liter
Salmo gairdneri
Chinook salmon Hardness:
70 mg/
liter
Salmo gairdneri
Salmo gairdneri
Micropterue
ealmoidee
Reference
Biesinger and
Christensen (1972)
Southern California
Coastal Water
Research Project
(1976)
Benoit (personal
communication)
Beisinger and
Christensen (1972)
Schiffman and
Fromm (1959)
Olson and Foster
(1956)
Carton (1972)
Fromm and Schiffman
(1958)
ological changes
in gut
-------
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146
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APPENDIX E
WET WEIGHT-DRY WEIGHT RELATIONSHIP
To determine the dry weight of each dissected crayfish, the dry weights
for all dissected tissues and the carcass had to be determined separately
then summed. The error in these values is greater than that of intact body
dry weights due to possible loss of body fluids and tissue during dissection
and to statistical compounding of the error. For this reason, wet weight
was used whenever a whole-body weight was needed in analysis. This occurred
when total chromium body burdens in the laboratory feeding experiment was
measured (expressed as ppb wet weight of chromium). Wet weights were also
used in the determination of weight-independent metabolic rates. Dry
weights of crayfish tissues and leaf material were easily and more
accurately determined, therefore all metal concentrations for these samples
are expressed as ppm dry weight.
Samples using the different weight expressions can be compared if wet
weight:dry weight ratios are used to convert to common units. The
relationship for the crayfish in both experiments is shown in Table E-l.
There is little difference in the regression equations for males and females
in the field experiment, thus a pooled regression for both sexes is
sufficient. The equation for the crayfish in the laboratory experiment is
quite different, however, so those data should be treated separately.
By substituting the measured wet weight value into the appropriate
equation, dry weight can be obtained. These dry weights may then be used to
obtain dry weight metal concentrations or metabolic rates. However, since
treatments were pooled to obtain the regression equations, there is some
risk in comparing the values for different treatments. If the regression
equations are not the same for all treatments, comparisons would be
distorted.
147
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TABLE E-l. RECESSION EQUATIONS, COEFFICIENTS OF DETERMINATION (r2),
AND SAMPLES SIZES FOR THE RELATIONSHIPS BETWEEN WET WEIGHT
AND DRY WEIGHT OF CRAYFISH USED IN EXPERIMENTS"1"
Crayfish group
Metabolic rate
experiment
Males
Females
Males + females
Crayfish dissected,
chromium-feeding
experiment
2
Regression equation r n
Y = 0.271X + 0.126 0.872 23
Y - 0.287X + 0.077 0.949 22
Y - 0.277X + 0.109 0.931 45
Y - 0. 204X + 0. 045 0. 823 8
Y - dry weight; X = wet weight,
148
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APPENDIX F
ACCURACY OF UNIVERSITY OF WISCONSIN NUCLEAR REACTOR DATA
Table F-l presents the metal concentrations obtained for the "blind"
standards inserted with the samples analyzed by the University of Wisconsin-
Madison Nuclear Reactor, along with the actual concentrations in the
substance. It should be noted that only muscle and exoskeleton were
analyzed by the reactor. The hepatopancreas was analyzed using facilities
of the Soil Science Department. The discrepancies are rather large, but
there is no apparent trend and most are within the same order of
magnitude. All analyses compared crayfish groups for each tissue;
comparisons were never made between tissues analyzed by different methods.
Consequently, it was valid to make comparisons using relative differences
between crayfish groups, even though the accuracy of the reactor standards
was questionable.
Because each value is based on only one sample rather than being a mean
of several, no attempt was made to correct the crayfish tissue values.
Statistical comparisons between groups would not change even if corrections
were made.
149
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Ui
o
TABLE F-l. COMPARISON OF METAL CONCENTRATIONS IN STANDARDS ANALYZED BY THE UNIVERSITY OF
WISCONSIN-MADISON NUCLEAR REACTOR WITH THE REPORTED ACTUAL CONCENTRATIONS OF THE
STANDARDS. STANDARD DEVIATIONS ARE REPORTED FOR THE VALUES OBTAINED BY THE
REACTOR LABORATORY AND FOR THE ACTUAL CONCENTRATIONS IS KNOWN.
Sample
SO-4
standard
Orchard
leaves
standard
Liquid
standard"
Source
Reactor
Koons and
Helmke (1978)f
Reactor
NBS§
Reactor
actual^
Metal Concentration(ppm)
Ba Cr Fe Se Zn
875.5+48.3 46.71+0.98 16,740+182 + 317.3+11.8
722+1.9 75+5.2 23,500+1.3 + 93+2.7
+ 3.008+0.346 + < 0.03 +
+ 2.3 + 0.08 +
49.34±2.58 27.84±0.19 + 0.363*0.009 +
60 40 + 5.0 +
+No value reported for actual value of standard.
tR.D. Koons and P.A. Helmke. Neutron Activation Analysis of Standard Soils.
Am. J., 42(2):237-240, 1978.
§U.S. National Bureau of Standards.
^Prepared from solutions of known concentrations.
Soil Sci. Soc.
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TECHNICAL REPORT DATA
(Please read Instructions on the reverse before completing/
REPORT NO.
EPA-600/3-80-081
3. RECIPIENT'S ACCESSION NO.
4. TITLE AND SUBTITLE
5. REPORT DATE „„„„
August 1980 Issuing date.
Responses of Stream Invertebrates to an Ashpit Effluent
isconsin Power Plant Impact Study
6. PERFORMING ORGANIZATION CODE
. AUTHOR(S)
John J. Magnuson, Anne M. Forbes, Dorothy M. Harrell,
and Judy D. Schwarzmeier
8. PERFORMING ORGANIZATION REPORT NO.
9. PERFORMING ORGANIZATION NAME AND ADDRESS
Department of Limnology
University of Wisconsin-Madison
Madison, WI 53706
10. PROGRAM ELEMENT NO.
1BA820
11. CONTRACT/GRANT NO.
R803971
12. SPONSORING AGENCY NAME AND ADDRESS
Environmental Research Laboratory-Duluth
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
Fly ash from the 527-MW coal-fired Columbia Generating Station Unit I (Columbia
Co., Wisconsin) is discharged as a slurry into an adjacent ashpit. Water from the
ashpit is pumped to a ditch that joins the ashpit drain and Rocky Run Creek before
they reach the Wisconsin River. Habitat alterations have been noted as relatively
minor changes in water quality parameters (e.g., alkalinity, hardness, pH, and
turbidity), as increased amounts of some dissolved trace elements (Cr, Ba, Al, Cd,
and Cu), and as the precipitation of trace elements (Al, Ba, and Cr) into a floe that
coats the stream bottoms. The ashpit drain became an unsuitable habitat for aquatic
invertebrates after Columbia I began operating.
The conductivity of the effluent increased in January 1977 when sodium bicar-
bonate was first used to increase the efficiency of the electrostatic precipitators.
Since then conductivity measurements have indicated effluent concentration at
distances downstream from the generating station.
Rocky Run Creek is still a suitable habitat for many aquatic invertebrates, but
evidence of sublethal stresses and habitat avoidance exists. The major effect of
Columbia I on aquatic invertebrates is hypothesized to be continued habitat alteration
and, in particular, reduced substrate quality and avoidance of unpreferred habitat.
The susceptibility of early life stages of crustaceans to the ash effluent may also be
important. Acute toxicity to adult forms is unimportant.
17.
KEY WORDS AND DOCUMENT ANALYSIS
DESCRIPTORS
b.IDENTIFIERS/OPEN ENDED TERMS
c. COS AT I Field/Group
Thermal pollution
Ashpit effluents
Aquatic invertebrates
Sublethal effects
Wisconsin power plant
study
Aquatic invertebrate
habitats
06/F
07/B
07/C
18. DISTRIBUTION STATEMENT
RELEASE TO PUBLIC
19. SECURITY CLASS (This Report)
UNCLASSIFIED
21. NO. OF PAGES
165
20. SECURITY CLASS (Thispage)
UNCLASSIFIED
22. PRICE
EPA Form 2220-1 (Rev. 4-77) PREVIOUS EDITION is OBSOLETE
•fr U.S. GOVERNMENT PRINTING OFFICE: 1980--657-165/0075
151
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