vvEPA
United States
Environmental Protection
Agency
Invirormerrta)
Latxjf attx y
Dulirrb MN 56804
Research and Development
Element Flow in
Aquatic Systems
Surrounding
Coal-Fired Power
Plants
Wisconsin Power
Plant Impact Study
T i ••'•• r A V v
:j| i .*. v/\ , y-
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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 nine series are
1 Environmental Health Effects Research
2 Environmental Protection Technology
3 Ecological Research
4 Environmental Monitoring
5 Socioeconomic Environmental Studies
6 Scientific and Technical Assessment Reports (STAR)
7 Interagency Energy-Environment Research and Development
8 "Special" Reports
9 Miscellaneous Reports
This report has been assigned to the ECOLOGICAL RESEARCH series This series
describes research on the effects of pollution on humans, plant and animal spe-
cies, and materials Problems are assessed for their long- and short-term influ-
ences Investigations include formation, transport, and pathway studies to deter-
mine the fate of pollutants and their effects This work provides the technical basis
for setting standards to minimize undesirable changes in living organisms in the
aquatic, terrestrial, and atmospheric environments
This document is available to the public through the National Technical Informa-
tion Service, Springfield, Virginia 22161
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EPA-600/3-80-076
July 1980
ELEMENT FLOW IN AQUATIC SYSTEMS SURROUNDING COAL-FIRED POWER PLANTS
Wisconsin Power Plant Impact Study
by
Anders Andren
Marc Anderson
Nicholas Loux
Robert Talbot
Institute for Environmental Studies
University of Wisconsin-Madison
Madison, Wisconsin 53706
Grant 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
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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 recommen-
dation for use.
ii
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FOREWORD
The U.S. Environmental Protection Agency (EPA) was created because
of increasing public and governmental concern about the dangers of pollution
to the health and welfare of the American people. Iblluted air, water, and
land are tragic testimony to the deterioration of our natural environment.
The complexity of that environment and the interplay between its components
require a concentrated attack on the problem. Research and development, the
necessary first steps, involve definition of the problem, measurements of
its impact, and the search for solutions. The EPA, in addition to its own
laboratory and field studies, supports environmental research projects at
other institutions. These projects are designed to assess and predict the
effects of pollutants on ecosystems. One such 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, involving investigators and
experiments from many academic departments at the University of Wisconsin,
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.
Norbert A. Jaworski
Director
Environmental Research Laboratory
Duluth, Minnesota
iii
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ABSTRACT
Water quality parameters of a 192-ha (480-acre) cooling pond adjacent
to the Columbia Generating Station, Portage, Wisconsin, has been investi-
gated. Analyses were made for major and minor elements, nutrients, pH,
alkalinity, 0^, chloroogranics, phenols, and polyaromatic hydrocarbons.
Similar parameters were also measured in the nearby fly ash discharge
basin and its associated drainage stream. Laboratory dissolution and
precipitation studies of fly ash were performed in an effort to understand
the chemistry of the discharged ash water and its potential effects on
receiving waters. Mass balance calculations were made and are presented
to ascertain whether the cooling pond acts as an efficent sink for in-
organic and organic compounds, and if so, what the fate of these compounds
is. Data presented in this report are also discussed in terms of plant
operating characteristics. Remedial procedures are presented which could
alleviate present and anticipated problems.
iv
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CONTENTS
Foreword
Abstract iv
Figures vi
Tables viii
1. Introduction 1
2. Conclusions and Recommendations 4
3. Methods 7
Experimental procedures used in the monitoring study 7
Experimental procedures used in the laboratory leaching study 9
4. Chemical Characteristics of the Columbia Cooling Lake 12
Literature Review 12
Results and discussion 13
5. Organic Contaminants in Cooling Pond Sediments 25
6. Chemical Characteristics of the Columbia Fly Ash Basin 27
Literature review 27
Laboratory leaching experiments 28
Results 29
Monitoring study of the Columbia ash basin 49
7. Environmental Impact on Natural Water Systems 69
Chemical mass balances of the cooling pond and ash basin.... 69
The effect of fly ash discharge and storage
on receiving waters 74
References 77
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FIGURES
Number Page
1 Location of sampling stations at Columbia Generating Station 8
2 Temperature profile at various stations in Wisconsin River and
cooling lake waters 14
3 Average monthly major element concentrations in
Wisconsin River and cooling lake waters 15
4 Average monthly nutrient concentrations in
Wisconsin River and cooling lake waters... 16
5 Calcium carbonate saturation in cooling lake as a function of
dissolved calcium concentration at various temperatures........ 22
6 Change in dissolved elemental concentrations as a function
of leaching time 31
7 Change in pH with leaching time in system open to atmosphere 33
8 Change in pH with leaching time in system closed to atmosphere... 34
9 Dissolved aluminum and silica as a function of the square root
of leaching time 36
10 Dissolved elemental concentrations as a function of pH 38
11 Silicate stability diagram with data points for K+/H+ 41
12 The change in dissolved potassium with pH 42
13 Dissolved iron as a function of the square root of leaching
time 43
14 Solubilities of Ca+2 and Mg+2 carbonates and hydroxides at 25°C.. 45
15 Electrophoretic mobilities of Columbia fly ash suspensions as
a function of pH. 47
16 Monthly comparison of Wisconsin River, cooling lake, and
ash basin analyses 49
VI
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17 Comparison of average Na, K, Mg, and Ca concentrations
in Wisconsin River, cooling lake, and ash basin 50
18 Comparison of cooling lake and Wisconsin River dissolved
metal concentrations with those observed in ash basin 53
19 Soluble element concentrations at various stations at the
Columbia plant for September 13, 1977 57
20 Soluble element concentrations at various stations at the
Columbia plant for September 27, 1977 58
21 Scanning electron microscope picture of windblown precipitate
from the shore at second basin of Columbia ashpit 60
22 Scanning electron microscope picture of the precipitate formed
on sediment trap float in the second basin of the
Columbia ashpit 60
23 X-ray diffraction pattern for windblown precipitate in the
second basin 61
24 X-ray diffraction pattern for precipitates on a float
in the second basin 62
25 X-ray diffraction pattern for surface-suspended sediments
in the first basin 63
26 X-ray diffraction pattern for sediments from the first basin 64
27 X-ray diffraction pattern for sediments from the first
station in the second basin 65
28 X-ray diffraction pattern for sediments from the second basin.... 66
vii
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TABLES
Number Page
1 Comparison of Element Mobilizations by Weathering and
Fly Ash Disposal 2
2 Average Daily Water Flow at Columbia Generating Station 10
3 Analytical Procedures Used in this Study 10
4 Concentrations of Major Elements in Columbia Cooling Pond and
Lake Mendota 17
5 Trace Elements in Cooling Pond and Wisconsin River 18
6 MINEQL Computer Program Results for Average Concentrations
Observed in Cooling Lake 20
7 Equations and Formation Constants Used in Constructing Figure 5.. 23
8 Elemental Ratios for Selected Elements in Suspended Solids 24
9 Aromatic Hydrocarbon Compounds Identified in the Columbia
Cooling Basin 26
10 Total Elemental Concentrations of Columbia Fly Ash
Expressed as Percentage of Dry Weight. 29
11 Equilibrium Constants Describing Ferric Hydroxide Solid
Hiase Equilibria with Major Aqueous Ferric Complexes 44
12 Log Equilibrium Constants at 25°C and Log pC02 = -3.52 46
13 MINEQL Computer Program Results for Soluble Concentrations
Observed During September 1977 Sampling Period 54
14 Comparison of Suggested Water Quality Criteria with
Values Observed in Ash Basin 56
15 Chemical Concentrations of Selected Dissolved Components
in Groundwater Samples from Ashpit Dike and Sedge Meadow 68
16 Mass Balance for Selected Elements in Cooling Pond 70
viii
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17 Neutron Activation Results for Suspended Sediments
Collected in Wisconsin River and Cooling Bond 71
18 Mass Balance for Ash Basin 72
19 Possible Elemental Concentration Changes in Wisconsin River
Due to Ash Basin Discharge 73
ix
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SECTION 1
INTRODUCTION
Water utilization may, in many instances, place severe constraints on
site selection and subsequent operational aspects of coal-fired steam
plants. Various federal and state regulations require that the discharged
water, whether it comes from cooling uses or fly ash disposal activities, be
maintained at essentially the same quality as that of the receiving water.
Maintenance of good water quality is not only desirable from an
environmental viewpoint (that is, minimizing the release of hazardous
substances), but is also important in the optimization of plant operations
(see, for example, Sigma Research Report 1975). The latter point is
especially germane where cooling ponds are present (Sams et al. 1978). It
therefore becomes important to understand the various environmental factors
that are involved in determining the water quality of these aquatic systems.
Similarly, the production of fly ash is, and will continue to be, a
tremendous disposal problem. The annual discharge of individual elements in
fly ash from coal-fired steam plants in the U.S. to landfills, fly ash
basins, and other receiving waters has been compared to the mobilization of
elements by natural processes, such as weathering (Klein et al. 1975).
Table 1 indicates that elemental mobilization by fly ash ranges from 0.2 to
82% of the natural weathering products carried by U.S. rivers. The
percentage will undoubtedly increase with the projected increase in coal
utilization.
Since at one time or another the fly ash will come into contact with
water, through disposal in either solid landfill, ash basin, or as soil
amendment, study of the aqueous behavior of this material becomes important
(Holland et al. 1975).
Few studies on the aquatic chemistry of cooling ponds exist in the
literature, but an increasing amount of literature on the chemical
composition of fly ash is now emerging. However, much more information is
needed on the aqueous behavior of various fly ash materials, especially how
this behavior relates to disposal questions.
The research presented in this report describes approximately 2 yr of
detailed investigations of the aqueous chemistry of a 192-ha cooling pond
and a 24-ha ash basin at the Columbia Generating Station near Portage,
Wisconsin. The specific objectives were: (1) to compare water quality in
the cooling pond to that of source waters and other Wisconsin lakes, (2) to
estimate the amount of chemicals that annually reach the cooling pond and to
relate this input to observed water concentrations, (3) to conduct
-1-
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TABLE 1. COMPARISON OF ELEMENT MOBILIZATIONS BY
WEATHERING AND FLY ASH DISPOSAL3
Weathering mobilizations Fly ash disposal
Element (x ICr tons/yr) (percent of weathering)
Al
As
Ba
Br
Ca
Co
Cr
Cs
Fe
Hg
K
Mg
Mn
Mo
Na
Pb
Sb
Se
Si
Th
U
V
Zn
73,000
15
550
34
42,000
17
80
4
43,000
0.5
27,000
19,000
770
3.2
18,000
22
4.3
0.8
75,000
11
3.4
120
120
3
15
3
0.2
2
5
7
7
9
1
2
2
1
-
1
7
4
82
8
5
18
6
12
aFrom Klein et al. 1975.
laboratory experiments on fly ash dissolution, and (4) to compare laboratory
experimental results with observed chemical parameters in a fly ash basin.
The first section deals with field measurements taken in the Wisconsin
River and at selected sampling stations in the cooling pond. Concentrations
of major, minor, and nutrient elements are discussed in terms of spatial and
temporal variations. These data are then compared to similar data sets
taken from Lake Mendota, Wisconsin, one of the best studied lakes in the
world. The mechanisms responsible for maintaining the observed levels of
chemicals in the cooling pond are then discussed in terms of a chemical
equilibrium model. A model is also presented which predicts calcium
carbonate precipitation (percent saturation) as a function of temperature
and calcium concentration. This model was developed in an effort to provide
plant operation guidelines since scaling problems in the cooling system
sometimes present a severe problem.
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The second major section deals with the chemistry of fly ash water
interactions. Controlled laboratory studies were designed to evaluate the
stoichiometry of liquid-solid phase reactions. Data from these measurements
are then discussed in terms of various chemical models, where emphasis was
placed on explaining the variables that control the dissolution and
formation of major mineral phases. These results are then compared to field
measurements from the ash basin.
Results from field and laboratory measurements are finally discussed in
terms of environmental input on natural water systems. While much of this
discussion deals with the Columbia Generating Station, it is nevertheless
felt that the methodology used in this evaluation is transferable to other
sites.
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SECTION 2
CONCLUSIONS AND RECOMMENDATIONS
Except for copper concentrations, the overall water quality of the
cooling basin at the Columbia Generating Station is similar to that of many
southern Wisconsin eutrophic lakes. Copper is elevated in the basin as a
result of copper piping which is used for cooling purposes within the
plant. Copper concentrations, however, are generally below those levels
thought to be toxic to aquatic life. Because of evaporation, the
conservative element concentrations (Na, K, SO^, and Cl) in the cooling pond
water generally are higher than those in source waters (the Wisconsin
River). Although variable, this difference is usually 20 to 30%. Nutrient
levels in the cooling pond are generally lower than those in the Wisconsin
River (exceptions discussed below).
The cooling basin acts as a repository for a large fraction of the
incoming non-conservative elements. That is, many of the nutrients and
trace metals (whether in dissolved or particulate form) are removed from the
water column by adsorption and settling.
Cooling basin bottom waters may turn anoxic during late summer, causing
nutrients from bottom sediments to be released to the overlying waters. Low
oxygen concentrations are common in both source and cooling pond waters
during this period. Since nutrient release from bottom sediments to the
water column may cause severe biological fouling problems, it is important
to maintain oxygenated cooling pond waters at all times.
Because of elevated temperatures and relatively high alkalinity values,
the cooling pond water is usually supersaturated with calcium carbonate.
This condition is most severe in late summer and can cause severe scaling
problems within the plant's cooling system. The most effective remedy
appears to be frequent blowdowns during this period.
Cooling pond water contained non-detectable levels (less than 7 to 10
ng/liter) of chlororganics and polyaromatic hydrocarbons. Bottom sediments
contained measurable quantities of chlorinated phenols, phthalate esters,
and polyaromatic hydrocarbons. The source of the latter is thought to be
windblown coal dust from the nearby coal storage area. While polyaromatic
hydrocarbon and chlororganic analyses of fish caught in the cooling basins
are not yet available, PCB concentrations were similar to those found in
fish collected from non-contaminated lakes.
-4-
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Results from laboratory studies on dissolution and precipitation
reactions of fly ash during equilibrium conditions can not be directly
extrapolated to the ash basin at the Columbia Generating Station. The short
residence time of water and intermittent fly ash input places kinetic
constraints on the system. However, results from laboratory studies and
chemical equilibrium modeling were extremely valuable in our attempt to
understand i>n situ control mechanisms.
Major element chemistry determines the reaction mechanism of fly ash
dissolution by establishing solid-solution metastable equilibria. Dissolved
calcium and magnesium concentrations are controlled by their carbonate and
hydroxide solid phases. Aluminum and iron hydroxide phases control their
dissolved concentrations in appropriate pH regions. Silica appears to exist
mainly as an amorphous phase but also may be incorporated in an alumino-
silicate phase at mid-pH values. Hydrolytic dissolution of Ca, Mg, Na, and
K are responsible for high pH values observed both in laboratory experiments
and in the ash basin waters. However, atmospheric C02 entering the solution
will lower the pH at steady-state conditions. The isoelectric pH (the pH
where the particles in aqueous suspension contain a net zero charge—pH-rvp)
of Columbia fly ash is approximately 7.55, which also indicates that Fe and
Al dominate the solid phase reactions at mid-pH values.
Dissolved Cd concentrations at high pH values suggest that this element
precipitates as hydroxide and carbonate phases. Below pH 9.0 adsorption
reactions most likely control its dissolved concentrations. Dissolved P
concentrations are stongly influenced by adsorption reactions between pH 4.5
and 8.5. Phosphorous most likely precipitates as a hydroxyapatite phase at
higher pH values. Co-precipitation of P with Ca and Mg probably also occurs
above pH 10. Other trace cations and anions behave in an analogous manner.
Elevated concentrations of Al, B, Cd, and Cu are present in the ash
basin at levels deemed toxic to aquatic life. In addition, pH values are
always in the range of 9 to 12, most of the time less than 11.0.
Elemental mass-balance calculations in the ash basin indicate that the
major fraction of elements discharged into the basin remain in the system
(except Na and SO^). Although ions such as Al, Cr, B, As, and Se might be
hazardous to biota, elemental discharge from the ash basin has a negligible
effect on the water quality of the Wisconsin River, with the possible
exception of Na, B, and Cr. The concentration changes in the river,
however, cannot be measured In situ because they are less than the standard
deviation of analytical precision. Thus, any biological effect that results
in the discharge of chemicals from the ash basin should be confined to the
drainage ditch.
Diversion of ash basin effluent into the cooling pond may lead to
several possibilities. Several elements are present in concentrations above
recommended water quality criteria. However, less than 40% of the water in
the cooling pond would come from this source. The only elements of
potential concern would then be SO^, B, and Al. Aluminum would rapidly
precipitate and possibly aid in removing part of the suspended sediments in
-5-
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the cooling pond. Further information is required concerning average ash
basin B concentrations. However, even after dilution, this element could be
present in significant concentrations. Sulfate would not directly be a
problem in the cooling pond after dilution. However, potential anaerobic
conditions in the cooling pond could lead to ^S generation. After Na^CO.,
had been added to improve the efficiency of the electrostatic precipitator,
Ca and COo concentrations in neutralized ash basin discharge waters were
often below concentrations present in the cooling pond. Thus, diversion may
partially alleviate the scaling problems in the plant. Some evidence
indicates that Al precipitation is enhanced once the ash basin water is
neutralized. Use of one of the current ash basins or construction of an
additional settling basin for receiving the neutralized effluent may
tremendously decrease total Al and adsorbed species discharges.
Although no environmentally significant concentrations of ash basin
contaminants have been observed in the surrounding groundwater at this time,
we believe that further study is necessary. Currently, modeling is
impractical because of the need for additional hydrogeological
information. This is particularly needed with respect to the amount of
leachate infiltrating directly through the ash basin dike and remaining on
the surface of the sedge meadow. Additionally, current wells may not be
directly intersecting ash basin infiltrate and new plastic wells may be
required before subsurface water movements can be adequately described.
Monitoring well water and ash basin dissolved element concentrations during
an extended period of steady-state flow would aid in developing a
representative model. Information concerning soil composition (cation
exchange capacity, organic matter, clay, Fe(OH)-, Mn(OH),, and Si02 content)
would also be required if adsorption modeling is attempted.
The periodic monitoring of existing wells surrounding the ash basin
should continue. In addition, samples of stagnant water bodies adjacent to
the ash basin should also be included in the sampling regime. If ground-
water concentrations exceed accepted water quality criteria, maximum ash
basin pumping (with consequent increased infiltration) could be delayed
until heavy rainfall and melting snow dilute surface and subsurface water
contaminants.
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SECTION 3
METHODS
EXPERIMENTAL PROCEDURES USED IN THE MONITORING STUDY
The Columbia Generating Station, located on the Wisconsin River in
south-central Wisconsin, consists of two nearly identical 527-MW units.
Columbia I went into operation in May 1975 and Columbia II in April 1978.
During the period of the monitoring study, Columbia I was burning high ash,
low sulfur coal from Montana.
Figure 1 shows the locations of the sampling stations where water
samples were collected at 1-month intervals from June 1976 to April 1977.
Wisconsin River water is pumped to the 192-ha (480-acre) cooling pond
through the ditch and underground pipes located west of the pond. Although
cooling water is not directly discharged from the cooling pond into the
Wisconsin River, during periods of blowdown water flows through a spillway
into a sedge meadow surrounding the plant. Fly ash is flushed into the
settling basins with cooling pond water obtained at the plant intake. Fly
ash leachate eventually reaches the Wisconsin River by way of the drainage
ditch depicted on the right and bottom of Figure 1.
Table 2 lists a water balance for average daily flow through the
site. Pumping records, temperature data, and well levels outside the ponds
were used by the hydrogeology subproject of the Columbia impact study to
derive these estimates of water flow. The cooling pond circulation time
averages 5 days and the hydraulic residence time approximately 80 days. The
depths of the cooling and ash ponds vary. However, the maximum depth at any
station is limited to 3 m. Further details on water flow around the power
plant, including flow characteristics and geomorphology, can be found in
Anderson and Andrews (1976).
A Keranerer water sampler was used to collect all water samples and 2-
liter polyethylene bottles were used for storage. Nucleopore (0.4 ym)
membrane filters were used for filtration. Trace metal samples were
preserved by adding 2 ml of redistilled nitric acid (16 ) per liter of
sample. The September 1976 suspended solid samples were analyzed by neutron
activation analysis on preweighed 0.45 pm Millipore filters dried at 80°C.
All polyethylene and glass containers used during collection and
analyses were washed with hot 50% hydrochloric acid (10% nitric acid for
trace element storage bottles) and rinsed several times with distilled water
and sample solution.
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Secondary
Sett Wig
101
10
ASH BASIN
(70 acres)
it-MJMPING
K STATION
Primary
Sett 1 in
1
GENERATING
STATION
flTSCHARGt!'?
i J.
'WISCONSIN
RIVER ,
:OAL HANDLING]!
COAL STORAGE
(1,200,000
tons)
INTAKE CHANNEL
BURIED PIPES
(480 acres) ;
LIB CROSS
ISLAND
Chi.,Mil., S.P.
5 Pac. Railroad
COLUMBIA GENERATING STATION SITE.
Figure 1. Location of sampling stations at the Columbia Generating Station,
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Table 3 lists the analytical procedures used in the study. In most
cases, determinations were performed utilizing techniques detailed in
Standard Methods for the Examination of Water and Waste Water (American
Public Health Association 1971, 1974). Calcium and magnesium analyses were
performed by adding of lanthanum to prevent refractory compound formation
during flame atomic adsorption. All flameless and most flame atomic
adsorption analyses were performed on a Perkin Elmer model 603 Atomic
Adsorption Spectrophotometer equipped with an HGA 2100 graphite
furnace.Dissolved oxygen and temperature analyses were performed on site
with a Yellow Springs Instrument dissolved oxygen meter. The instrument was
precalibrated several times using the Winkler dissolved oxygen technique and
laboratory thermometer. The carmine spectrophotometric procedure for boron
was performed using a standard additions procedure. X-ray diffraction
analyses of ash basin sediment samples were performed using a mounting
procedure described by Gibbs (1965).
The three stations located at the intake channel and in the preliminary
settling basin (Figure 1) were used to evaluate the chemical characteristics
of incoming Wisconsin River water. Single samples were collected at the
intake and discharge stations depicted on the map. At all stations inside
the cooling and ash basins, water samples were obtained at surface, mid-
depth, and bottom.
From June to October 1976, 28 monthly water samples were collected,
filtered, and analyzed for 15 parameters. In November, three stations (110,
113, and 114) were deleted from the monitoring schedule since variations in
dissolved element concentrations in the cooling pond were minimal. Sulfate
and nitrogen analyses were discontinued at that time, and total dissolved
phosphorus, aluminum, cadmium, chromium, copper, and iron were added to the
list of measured elements. In September 1977, a short-term intensive
sampling of the ash basin was instituted with an overall design of
evaluating sedimentation in the ash basin in addition to estimating
dissolved concentrations of arsenic, lead, zinc, and boron.
The results presented in this report are reduced to average values and
ranges in elemental concentrations observed at the sampling stations.
Appendix A provides specific details regarding precision of analyses and raw
data obtained during the period of study.
EXPERIMENTAL PROCEDURES USED IN THE LABORATORY LEACHING STUDY
Composite samples of fly ash used in the laboratory experiments
include: (1) fly ash collected underwater off the end of the ash delta
(submerged), (2) freshly deposited fly ash on the ash delta (wet delta), (3)
fly ash deposited on the ash delta but dewatered (dry delta), and (4) fly
ash from stages II and III of the electrostatic precipitator hopper bins (5
September 1975). Samples 1, 2, and 3 were collected from the ash delta at
locations indicative of various stages of leaching by pond waters or
rainfall. After collection, each composite sample of fly ash was air dried
at room temperature (20°C) and subsequently sieved through a U.S. Standard
325 sieve. The fly ash fraction collected from sieving consisted of
-9-
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TABLE 2. AVERAGE DAILY WATER FLOW AT THE COLUMBIA GENERATING STATION3
o
Daily flow (m /day)
Source
Destination
5.4 x 10*
1.1 x 104
2.1 x ID4 b
1.2 x 104
1.0 x 104 b
Wisconsin River
Cooling pond
Cooling pond
Cooling pond
Cooling pond
Cooling pond
Spillway
Groundwater
Ash basin
Atmosphere (evaporation)
aAnderson and Andrews (1976).
Estimates, Anderson and Andrews (1976).
TABLE 3. ANALYTICAL PROCEDURES USED IN THIS STUDY
Element
Procedure
Reference
Alkalinity
Aluminum
Boron
Cadmium
Calcium
Chloride
Chromium
Copper
Iron
Lead
Magnesium
Nitrate
Nitrite
Oxygen
PH
Phosphorus
Potassium
Silica
Sodium
Sulfate
Temperature
Zinc
Titrimetric
Flameless atomic absorption
Carmine-Spectrophotometric
Flameless atomic absorption
Flame atomic absorption
AgN03-Cr04
Flameless atomic absorption
Flameless atomic absorption
Flameless atomic absorption
Flameless atomic absorption
Flame atomic absorption
Cadmium reduction
Sulfanilamide-Spectrophotometric
Yellow Springs Instrument probe
Glass electrode
Phosphomolybdate-Spectrophotometric
Flame atomic absorption
Molybdosilicate-Spectrophotometric
Flame atomic absorption
Turbidimetric
Yellow Springs Instrument probe
Flame atomic absorption
Standard Methods (1971)a
Perkin Elmer (1975)b
Standard Methods (1971)
Perkin Elmer (1975)
Standard Methods (1974)°
Standard Methods (1971)
Perkin Elmer (1975)
Perkin Elmer (1975)
Perkin Elmer (1975)
Perkin Elmer (1975)
Standard Methods (1974)
Standard Methods (1974)
Standard Methods (1974)
Standard Methods (1974)
Standard Methods (1974)
Eisenreich et al. (1975)
Standard Methods (1974)
Strickland and Parsons
(1968)
Standard Methods (1974)
Standard Methods (1974)
Perkin Elmer (1975)
aAmerican Public Health Association (1971).
bPerkin (1975).
cAmerican Public Health Association (1974).
-10-
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particles less than 44 ym in diameter. Drying of the sieved fly ash was
completed in a dessicator where the samples were stored until use.
Duplicate equilibrium leaching experiments were conducted in large
polyethylene containers. A solid-solution ratio of 1 g ash/liter of doubly
distilled water was utilized for each fly ash sample. These slurries were
stirred for several months at a constant rate using electric stirrers
equipped with polyethylene blades. At selected time intervals, an aliquot
was removed from each slurry and the pH was measured with a low junction
potential electrode (Sargent Welch Model S-30072-25). Each aliquot was then
filtered through a prewashed 0.4 um Nuclepore filter, and the filtrate was
divided for elemental analysis. Dissolved aluminum (Okura et al. 1962),
dissolved reactive phosphorus (Murphy and Riley 1962), and dissolved
reactive silica (Strickland and Parsons 1968) were determined
colorimetrically from one fraction. The remaining filtrate was acidified to
0.5% nitric acid for later analysis by atomic absorption spectroscopy
(lerkin Elmer Model 603) for Ca, Cd, Fe, K, Mg, and Na. Solids collected on
the Nuclepore filters were subjected to analysis by X-ray diffraction.
During electrophoretic mobility measurements large volumes of fly ash
slurries were prepared using the same four fly ash samples. A constant
solids concentration of 200 mg/liter was used to facilitate zeta potential
measurements. The slurries were stirred for 1 week to obtain pH
stability. At this time, 12 aliquots from each slurry were dispensed into
250 ml linear polyethylene bottles (LPE). The pH was adjusted from 1.0 to
12.0 with perchloric acid or potassium hydroxide. These slurries were then
shaken at a constant temperature (20°C) for 1 additional week. The final pH
was recorded and the electrophoretic mobility was determined with a lazer-
Zee Meter (Pen-Ken, Inc.). The remaining slurry was filtered and elemental
analyses were conducted on the filtrate as previously described. All field
samples were filtered in the same manner so that dissolved concentrations
could be compared directly with those determined in these laboratory
experiments.
Each fly ash sample was subjected to wet digestion to obtain values for
the total concentration of major and minor elements. A satisfactory method
was developed for digesting 25 mg of fly ash in a LPE bottle below the
boiling point of hydrofluoric acid (112°C). This technique minimized silica
loss by volatilization as fluoride compounds, a major problem associated
with the decomposition of inorganic siliceous materials (Langmyhr and Paus
1968). The LPE bottles containing the fly ash and acids (HN03, HC1, and HF)
were placed on a hot sand bath at 100°C + 3°C for digestion. After cooling,
the solution was diluted and elemental analysis was performed by colorimetry
or flame atomic absorption spectroscopy. Duplicate sample digestions were
conducted five times. Calculations of individual elemental recoveries are
based on analyses of NBS Standard Fly Ash #1633 using the digestion
technique described above.
-11-
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SECTION 4
CHEMICAL CHARACTERISTICS OF THE COLUMBIA COOLING LAKE
The initial investigation at the site of the Columbia Generating
Station examined the water quality in the Columbia cooling lake. The
primary objective of this study was to determine the major processes
influencing water quality in the Columbia cooling basin. Potential hazards
and suggestions for minimizing possible problems were evaluated by comparing
cooling lake water with data on water quality characteristics from the
Wisconsin River and selected southern Wisconsin lakes. In addition, a
computer program was used to predict elemental speciation and saturation in
the cooling lake.
LITERATURE REVIEW
The early literature concerning water quality of power plants deals
primarily with water treatment practices required for plant maintenance.
These practices included addition of toxic compounds (e.g., Hg, Zn, Cr, B,
and Cu salts) and complexing agents (e.g., phosphates and organic compounds
such as EDTA) for prevention and minimization of fouling and scaling
problems. A recent concern about water quality degradation in discharge
waters caused by these additives was reported by Chamberlain and Anderson
(1971). These authors suggested the use of ion exchange resins for removing
zinc-organic inhibitors from discharge waters. Stratton and Lee (1975)
found elevated concentrations of nitrate, phosphate, sulfate, zinc, iron,
copper, chromium, and mercury in cooling tower blowdown water.
Concentrations of manganese, nickel, and cadmium were not sufficiently high
to be of environmental concern. The authors attributed most of the elevated
concentrations to chemicals used in water treatment.
In recent years several reports have addressed the environmental impact
of cooling systems on aquatic environments. A report by Sigma Research
(1975) recommended the development of models for the realistic evaluation of
power plant operation on biota. In particular, the authors were concerned
about the effects of sublethal chlorination. Anderson and Smith (1977)
reported increased mercury concentrations in cooling basin sediments after
power plant operation, although similar elemental increases were not
observed in fish. Sams et al. (1978) used a holistic approach to determine
the relationship between the chemistry of a cooling reservoir and its source
water. Using background data and information from samples obtained during
the study period, the authors observed a progressively increasing
concentration of major elements in the reservoir (presumably through
evaporative concentration). At the same time, levels of nutrients decreased
and concentrations of dissolved reactive phosphorus, total phosphorus,
-12-
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dissolved reactive silica, and iron typically were lower than concentrations
observed in the source water. Dissolved oxygen ranged from 5 to 11
mg/liter, except for sporadically low values observed during the summer.
Concentrations of potentially toxic trace elements (chromium, copper, lead,
and zinc) were near or below the analytical detection limit. The authors
concluded that: (1) dissolved solids increased by a factor of 3 to 7; (2)
phosphorus was precipitating with iron and calcium; and (3) the environment
was becoming less suitable for autotroph production.
One may infer from these reports that drastic alterations in the water
quality of cooling basins are not to be expected in the absence of toxic
additives. Evaporative concentration may exert a significant effect on the
biotic communities only when exceptionally high concentrations of dissolved
solids occur. The elevated water temperature present in cooling basins may
accelerate biological processes and may limit the dissolved oxygen
concentrations during the summer months.
RESULTS AND DISCUSSION
A comparison of parameters at various stations in the cooling pond at
the Columbia Generating Station indicated that the water circulation period
was sufficiently short to preclude horizontal and vertical gradients for any
of the elements determined. Figure 2 illustrates a surface temperature
gradient in the cooling pond for September 1976. This profile is typical
for periods when the plant is operational. Surface and bottom elemental
concentrations demonstrate that the cooling pond is ordinarily well mixed.
Except for a period in August 1976, when bottom oxygen levels were low and
dissolved reactive phosphorus concentrations were high, all surface, mid-
depth, and bottom samples contained roughly equivalent elemental
concentrations.
Average concentrations of calcium, magnesium, sodium, potassium,
chloride, sulfate, and alkalinity invariably remained higher in the cooling
pond than in the Wisconsin River (Figure 3). In addition, a gradual trend
toward higher concentrations was observed at the end of the sampling
period. These factors indicate evaporative concentration. Magnesium,
sodium, and chloride display reasonably similar time profiles and may be
regarded as conservative elements. The anomalous relationship between the
river and the cooling pond for sodium and chloride in January 1977 may
possibly be explained as a short-term phenomenon caused by road salt
application during the winter months.
When compared with major elements, nutrient behavior is much more
variable and concentrations in the cooling basin normally are lower than
those in Wisconsin River water (Figure 4). The dissimilarities in the
behavior of major elements and nutrients indicate that predictive modeling
for the nutrients is difficult. For major elements predictive modeling is
relatively straightforward if evaporation rates are known. The most
significant nutrient variations occur during August when bottom water oxygen
levels approached the detection limit. At this time, total and dissolved
reactive phosphorus concentractions increased by a factor of 100. This
suggests that phosphorus is retained in the sediments with iron. The
-13-
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O 30
LU
QC
D
<
DC
LU
Q_
2
LU
H
20
10
A A
RIVES COOLING LAKE
Figure 2. Temperature profile at various stations in Wisconsin River and
cooling lake waters (obtained September 1976).
-14-
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A WISCONSIN RIVER
O COOLING LAKE
z
o
1-
DC
i-
LLJ
Z
0
o
DC
^
O
^
o
o
-3
Ca
-4
-3
Mg
-4
Na -4
-4
K
-5
-3
Cl
-4
S04 -4
Alk. -3
ooS8s@88^"
A A
f^
/""\ f^ >¥ >V AA i^r fc'i )0<
-
@ e o $ s o
"8 s Q ° 8
@ © & 8 8
BBS
Q Q Q O @ O
. a a o . o o o o o . .
JJASONDJFMA
1976
1977
Figure 3. Average monthly nutrient concentrations in Wisconsin River and
cooling lake waters. (Dotted lines depict detertion limits.)
-15-
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A WISCONSIN RIVER
O COOLING LAKE
Z
O
tr
h-
z
LU
O
Z
o
o
cc
-5
NO3-6
-7
-5
NO2-6
-7
-5
Total "6
P04 _7
-8
-5
DR "6
4 -7
-8
-4
Si -5
-6
-3
O
-4
A .
o A A o o
° n
-
-
O O O O---O
A
a e o o ° °
-
A
o
A
A A A A -
— o— -o o— o— o— o— O---O— o — °--
A*£AAA*AAA
.00o§§0°oo
-
-
AA A A^^ftO
LJ A Q ^ ^^ w
1 1 1 1 1 1 1 1 1 1 1
JJASONDJFMA
1976
1977
Figure 4. Average monthly major element concentrations in Wisconsin River
and cooling lake waters.
-16-
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release of phosphorus during Fe(OH)o reduction has been demonstrated by
numerous investigators (for example, Syers et al. 1973) and will not be
discussed here. Other indications of possible reducing conditions can be
observed in the nitrate and nitrite profiles in Figure 4. Bacterial
reduction may explain the lower nitrogen concentrations associated with the
elevated phosphorus levels. The maximum in dissolved reactive silica in the
fall of 1976 may also be explained by biological activity since diatom
growth exerts significant effects on soluble silica in other Wisconsin lakes
(Vigon 1976).
Major element concentrations in the cooling pond were compared in Table
4 with concentrations observed in Lake Mendota, one of the most thoroughly
studied lakes in the world. In almost all cases, the ranges in
concentrations are remarkably similar. The most significant difference
appears to be the temperature maximum observed in the cooling pond.
However, closer inspection of the nutrient concentrations yields additional
insight into processes occurring within the cooling basin. The dissolved
reactive phosphorus concentrations in the cooling pond, considerably lower
than in Lake Mendota, are often near the detection limit of 0.003
mg/liter. This may indicate a limiting nutrient status. Total phosphorus
concentrations observed in both bodies of water lead to their classification
TABLE 4. CONCENTRATIONS OF MAJOR ELEMENTS IN THE
COLUMBIA COOLING POND AND LAKE MENDOTA
Element
Concentrations observed
in cooling pond
(mg/liter)
Concentrations observed
in Lake Mendota3
(mg/liter)
°2
Ca
Mg
Na
K
N03
N02
Total P
Ortho P
Ortho Si
Cl
SO-
Alkalinity15
PH
Temperature (°C)
0-15
28-36
10-20
7-12
1.8-2.0
<0. 01-0. 10
<0. 01-0. 02
0.03-M.7
<0.003-1.7
0.1-1.6
9-19
12-19
117-142
6.6-8.2
0-37
0-15
26-30
23-28
4.5-8.0
3.5-4.0
<0.01-0.7
0.0025-0.02
0.05-0.65
0.02-0.40
0.1-1.5
6.2-9.6
18-30
140-193
6.5-9.2
0-27
Unpublished data, Water Chemistry Program, Univ. of Wisconsin-Madison
(1965).
'Expressed as mg/liter CaCO-j.
-17-
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as eutrophic lakes (Vollenweider 1968). A comparison of seasonal
distributions of total phosphorus in the cooling pond with those observed in
Lake Mendota (Sonzogni 1974) illustrates a late fall maximum for epilimnetic
water in Lake Mendota in contrast to the summer maximum observed in the
cooling pond. Since the cooling pond is normally well mixed, its chemical
and physical characteristics approach those of epilimnetic water in a
stratified lake. The presence of a hypolimnion isolates the epilimnion from
sediment interactions until fall overturn. In contrast, the water in the
cooling pond is immediately influenced by sedimentary processes. Since
anoxic conditions in the sediments are more likely to occur in the summer
when the temperature is high and wind action low, a maximum amount of
phosphorus will be released to the water during this season. Additionally,
since the sediment-water interactions are normally confined to the upper few
inches of the sediment, then the speed and magnitude of a response will show
an inverse dependence to volume of water in the overlying column. Ptwoni
(1974) examined the nutrient concentrations in 10 shallow manmade
impoundments in south-central Wisconsin. The author observed higher and
more variable median summer dissolved phosphorus concentrations in these
lakes when compared to a larger body of water. In summary, the cooling pond
and other shallow basins will experience an immediate response to sediment
conditions, and the magnitude of the responses will be inversely dependent
on the depth of the basin.
The range and average values for trace elements in the cooling pond
were also compared with these values in the Wisconsin River waters (Table
5). For all elements except copper, cooling pond concentrations are equal
to or lower than those found in the source water. Comparison of copper
concentrations at plant intake and discharge demonstrates that the plant is
a significant source of this element in the system. Copper concentrations
are often twice as large in discharge waters. The most plausible source for
TABLE 5. TRACE ELEMENTS IN THE COOLING POND AND THE WISCONSIN RIVER
Cooling Pond
Wisconsin River
Element
Al
Cd
Cr
Cu
Fe
Pb
Zn
B
Range
(mg/liter)
0.003
<0.0001
<0.001
0.002
<0.01
<0.002
<0.005
<0.1
- 0.080
- 0.0003
- 0.011
- 0.12
Average
(mg/liter)
0.027
<0.0001
0.008
0.10
Range Average
(mg/liter) (mg/liter)
0.027 -
<0.0001 -
<0.001 -
<0.0003 -
0.22 -
<0.002
<0.005 -
<0.1
0.102
0.0001
0.001
0.002
0.52
0.010
0.056
<0.0001
<0.001
0.001
0.35
0.010
-18-
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increased copper concentrations is from within the plant because soluble
copper values decrease when the plant is not operational.
Data for soluble aluminum and iron indicate that both elements are
precipitating in the cooling basin. The concentration differences between
the cooling pond and source water suggests that the majority of these
elements remain in the sediments of the cooling pond. The concentrations of
all trace elements are well below water quality recommendations, although
copper concentrations approached 0.020 mg/liter on several occasions, an
upper limit suggested by Van Hook and Shuttts (1976).
A computer program was used as an aid to determine whether selected
dissolved elements exceed their solubility product. Table 6 presents the
results obtained from the MINEQL computer program designed to solve
simultaneous equilibrium equations (Westall et al. 1976). Using average
_0
values of cooling pond concentrations, an ionic strength of 6x10 , a
constant pH of 8.0, and neglecting redox and adsorption reaction (assuming
oxygenated water), the program generated a prediction of thermodynamic
equilibrium concentrations for a variety of elements. The second column in
Table 6 represents the initial molar concentration of dissolved elements and
succeeding columns display the predicted final equilibrium percent
distribution of soluble species and precipitates. Three species, Fe(OH)o,
AltOH)-}, and CaCOg are predicted to precipitate in the basin. Although
copper concentrations decline when the plant ceases operation, no insoluble
compound is predicted. Presumably, the rate of plant input is balanced by
the rate of biological assimilation or inorganic adsorption of copper onto
sedimenting particles in the basin (when the plant is operational). The
majority of the trace elements (assuming no adsorption) have not exceeded
solubility constraints. Therefore, for accurate modeling, a subroutine must
be added to the program to account for adsorption reactions.
Problems with scaling have been reported at the Columbia plant (WPL
personal communication 1977), particularly in late summer. Figure 5
illustrates the saturation of calcium carbonate with temperature assuming
constant pH (8) and alkalinity (120 mg/liter). This figure was derived from
the equations and temperature-dependent formation constants listed in
Table 7. An increase in temperature from 20 to 35°C or a calcium
concentration of 20 mg/liter results in potential precipitation of one-half
the soluble calcium. Obviously, the solubility of calcium carbonate is
sensitive to temperature as well as pH and alkalinity. Scaling problems may
be minimized only by decreasing one of these three variables. An average
cooling pond calcium concentration of 30 mg/liter demonstrates that calcium
carbonate is supersaturated at any of the temperatures listed in Figure 5,
which is consistent with MINEQL calculations based on formation constants at
25°C. Predictions of the formation of this compound in the cooling pond, in
addition to within the plant, are thus possible. Although these
calculations have not been corrected for ionic strength, this effect would
be minimal. Figure 5 can thus be utilized as a reasonably accurate
predictive tool for saturation of calcium carbonate in the cooling pond.
-19-
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TABLE 6. MINEQL COMPUTER PROGRAM RESULTS FOR AVERAGE
CONCENTRATIONS OBSERVED IN THE COOLING POND (25°C)
Initial
Element concentration3
Ca 9.06xlO~4
Mg 7.49xlO~4
K 4.30xlO~5
Na 3.41xlO~4
Cu 1.42xlO~7
Cd 8.90xlO~10
S04 1.84xlO~4
Cl 3.81xlO~4
Soluble
species
Ca-"
+CaHCO+
CaS04
Mg*4
MgHC03+
MgS04
MgC03
K+
Na+
CuC03
Cu(C03)2~2
Cu(OH)+
Cu44"
Cd"1"1"
Cd(OH)+
CaCl+
CdS04
CdC03
so4
MgS04
CaS04
Cl~
% Precipitate %
60.1 CaC03 36.1
2.2
1.0
92.2
3.4
2.0
1.4
99.9
99.9
93.2
2.6
2.2
2.0
87.4
7.4
2.4
1.5
1.3
87.0
8.0
4.9
100.0
(continued)
-20-
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TABLE 6 (continued)
Element
Si04
TO4
Fe
Al
co3
Initial
concentration3
1.42xlO~5
4.16xlO~8
13.3
1.61xlO~6
9.27xlO~7
2.46xlO~3
Soluble
species
H4Si04
Al H3Si04
HP07
4
H2ro4
MgHP04
CaHP04
A1(OH)~
HCO~
H2C03
MgHC03
% Precipitate
97.1
2.9
73.4
10.6
2.6
Fe(OH)3
2.5 A1(OH)3
82.1 CaC03
1.4
1.0
%
100.0
97.5
13.3
aTotal molar dissolved cooling pond concentrations.
-21-
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3OO
c
0
••
+*
<0
k
3
+*
(0
(A
2OO
1OO
2O C
pH = 8.O
Alk. =12O mg/L as CaCO3
2O
25 3O
Ca (mg/L)
35
4O
Figure 5. Calcium carbonate saturation in the cooling lake as a function of
dissolved calcium concentration at various temperatures.
-22-
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TABLE 7. EQUATIONS AND FORMATION CONSTANTS USED IN CONSTRUCTING FIGURE 5
Equations
r . , _ Measured alkalinity
11 J 507000
[CO,] =
[TA] K0
[IT] + 2 K
-H-. _ Ca concentration (mg/liter)
J ~ 40,080
K
sp
% saturation
[CO']
tCa"""] [C03] x
100
K
sp
Formation constants
Temperature (°C)
[TA]
[H+]+2K,
|Ksp
100
20
25
30
40
4.17X10"11
4.68x10 U
5.13X10"11
6.02xlO~U
5.25xlO~9
4.57xlO~9
3.98xlO~9
3.02xlO~9
aThe effects of conjugate bases other than carbonate are minimal on the
total alkalinity measurements.
bGarrels and Christ 1965.
Suspended solids in both cooling pond and the Wisconsin River waters
ranged from 8 to 15 mg/liter for the samples obtained during the study.
Table 8 displays elemental concentration ratios for several sparingly
soluble elements present in suspended sediments obtained during September
1976. Since total quantities of suspended sediments in the cooling pond and
Wisconsin River were equivalent, ratios ranging from 2 to 3 imply a
fractionation of inorganic sediments once the source water enters the
cooling pond. The majority of the sediments entering the system appear to
deposit in the bottom of the cooling basin. Biogenic material generated
within the cooling pond then replaces the fraction lost from sedimentation.
-23-
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TABLE 8. ELEMENTAL RATIOS FOR SELECTED ELEMENTS IN SUSPENDED SOLIDS
Wisconsin River to Cooling Pond
Element elemental ratios for suspended solids
Co 2.0
Hf 2.7
La 2.8
Lu 2.8
Rb 1.8
Sb 2.3
Sc 3.1
Sm 3.0
Tb 2.7
-24-
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SECTION 5
ORGANIC CONTAMINANTS IN COOLING POND SEDIMENTS
Welch (1979) has recently received our present knowledge on coal burning
related organic chemicals. He concluded that aromatic hydrocarbons,
particularly the polyaromatic hydrocarbons (PAH), should receive particular
attention. John and Nickless (1977) found several PAH in river sediments
downstream from a coal mining area (Table 9). Waditer and Blackwood (1978)
also found that PAH could leach from coal storage areas (Table 9). These
observations prompted Welch (1979) to analyze sediments for both chlorinated
and nonchlorinated PAH in cooling pond sediments at the Columbia Power
Generating Station. The coal storage area adjacent to the cooling pond has
a capacity of about 1,200,000 tons. It was postulated that part of the coal
becomes windborne and will subsequently settle into the cooling basin. It
was also postulated that these PAH, if present in the sediments, would
become chlorinated because of rather efficient chlorination of discharged
cooling water. Sediments were thus collected approximately 25 m from the
cooling water discharge using a power dredge. The sediment was stored in
pre-cleaned glass bottles with Teflon-lined caps.
Details of the analytical procedure has been published by Welch
(1979). Briefly, dry sediment was soxhlet extracted with methylene chloride
for 5 days. The extract and solvent was then dried (anhydrous sodium
sulfate), evaporated, and filtered (Whatman No. 2). The filtered solution
was finally evaporated to dryness, weighed, and dissolved in circa 15 ml
benzene. Sulfur was removed using freshly precipitated copper. The
solution was again evaporated to dryness and subsequently taken up in 0.5 ml
of a 1:1 methanol-benzene mixture. This fraction was then cleaned using gel
permeation chromatography, evaporated to dryness, and finally taken up in
0.5 ml acetonitrile. This sample was then fractionated using HPLC and
finally analyzed with a Finnigan 4023A GC/MS/DS system. No recovery studies
were performed. The results must therefore only be considered in a
qualitative sense.
The sediment (155 g, dry weight) yielded approximately 33 mg of
extractable organic material. The specific compounds are identified in
Table 9. No chlorinated PAH were found. As a matter of fact, very few PAH
were found. While several other PAH undoubtedly are present, they are below
the limit of detection. The same technique was used to identify aromatic
hydrocarbons in the Duluth-Superior Harbor. Over 60 aromatic compounds were
identified including about 40 PAH and their alkylated homologs. Several
chlorinated biphyenyls were also identified. Most aromatic compounds, if
present in the Columbia cooling lake, are thus quite low in concentration.
-25-
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TABLE 9. AROMATIC HYDROCARBON COMPOUNDS IDENTIFIED IN THE
COLUMBIA COOLING BASIN3
Compound M+ Other ions
Methylethylbenzeneb 120 105, 91, 77
Dimethylnapthaleneb 156 141, 128
Dichlorothiazolopyrimidinec 205 207, 209, 170, 172
Hienanthrene 178 152, 89
Tetrmethylphenanthrene 234 219, 204
aFrom Welch (1979).
Exact structure not determined.
^Tentatively identified.
dGC/MS conditions: 2 mm x 1.8 m glass column with 3% OV-1 on 60/80 Gas
Chrom Q; temperature program 100-225°C/4°/min, initial hold 1 min.;
Finnigan 4023A scanned 50-500 amu at 2.0 sec/scan, started at time of
injection.
-26-
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SECTION 6
CHEMICAL CHARACTERISTICS OF THE COLUMBIA FLY ASH BASIN
The second major area of inquiry at the Columbia Generating Station was
concerned with the identification of the potential hazards associated with
fly ash disposal in an impoundment pond. Two approaches were developed for
meeting this objective: (1) Controlled laboratory leaching studies designed
to evaluate the stoichiometry of liquid-solid phase reactions and (2) on-
site measurement of major and selected trace element concentrations. The
mobilization of potentially toxic elements is limited by adsorption and/or
precipitation reactions on solid phases present in the ash basin.
Therefore, before a theoretical model for elemental mobilization can be
generated, these solid phases must be characterized in terms of their
mineral composition and their surface charge properties. The mineral
composition must be assessed because of the possibility of co-precipitation
of environmentally significant trace elements, and the surface charge
properties are significant because of their profound influence on adsorption
reactions. After it was determined that soluble element concentrations for
major and minor elements were controlled by solid phase reactions, both the
laboratory leaching and the on-site monitoring studies focused on solid
phase identification.
Identification of potentially toxic concentrations of various elements
was achieved in the on-site measurement study. However, a theoretical
prediction of behavior (especially for trace constituents) cannot be derived
because all thermodynamically stable solid phase reactions have not been
definitively identified. Kinetic studies will be required once the
appropriate processes have been determined.
LITERATURE REVIEW
The previous literature on fly ash leachate may be divided into five
categories. The first area, typified by toxicity studies by Reese and
Sidrok (1956) and Holliday et al. (1958), demonstrated that elevated
concentrations of Al, Mn, and B in fly ash leachate may be toxic to aquatic
life. These studies were concerned primarily with observing elemental
concentrations and identifying the parameters limiting optinum biotic
development. A second area of inquiry concerned the pozzolanic
characteristic of fly ash and the suitability of using the material as a
filler concrete. Brink and Halstead (1956) examined the matrix composition
of several types of fly ash and determined the potential extractable
components under varying leaching conditions. A third area of study
concerns laboratory leachate studies investigating major and trace element
mobilization and its possible relationship to initial matrix composition.
-27-
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Jones and Lewis (1960), Shannon and Fine (1974), Natusch (1975, 1976), Theis
and Wirth (1977), and Eggett and Thorpe (1978) demonstrated elevated
concentrations of S04, Ca, Mg Na, K, Si, Al, B, Fe, Cd, Cu, Hg, Cr, Zn, H,
and OH in fly ash leachate and attempted to determine correlations between
elemental mobilization and the physical and chemical characteristics of the
fly ash. Studies by Tenny and Echelberger (1970) and Higgins et al. (1976)
illustrate the fourth area of inquiry which concerns the suitability of
using fly ash for treatment of eutrophic lakes. Both studies characterized
short-term elemental fluctuations in terms of the fly ash leachate and
adsorption of phosphorus and organic material from water derived fron
eutrophic lakes. The final area of study is exemplified by more recent
studies concerning elemental mobilization with attendant environmental
degradation at several power station fly ash disposal sites located in the
United States. Cherry et al. (1976) demonstrated biological magnification
of several potentially toxic elements in aquatic organisms at a power
station in South Carolina. Coutant et al. (1978) described solid and
soluble phase chemical compositions at several sampling sites in a drainage
ditch adjacent to a power plant in Tennessee. Theis and Richter (1978)
documented the impact of fly ash leachate on the groundwater at a power
station near Lake Michigan.
Previous site-specific studies on the interaction of fly ash with water
have focused mainly on the chemical composition of the leachate. Although
such studies are useful it is also important to keep in mind that the solid
phase determines subsequent aqueous interactions. The present literature
does not contain enough information that specifically relates the chemical
composition of the total phase (that is, mineral phase composition) to that
of expected elemental concentrations in the leachate. As a first
approximation, the potential environmental impact of fly ash disposal is
limited by the characteristics of the parent coal. Extreme variations in pH
(from acidic to basic) and soluble major and minor element compositions in
fly ash leachate have been observed in the literature. Such variations make
it very difficult to predict the environmental impact of fly ash and
indicates that a better understanding of aqueous interactions can only be
made by combining detailed field and laboratory studies.
LABORATORY LEACHING EXPERIMENTS
Equilibrium leaching experiments of fly ash in water were included as
part of the aquatic chemistry study since adequate information describing
those processes that control the chemical composition of ash pond waters is
not available. These systems provide an environment in which certain trace
elements may be leached from impounded ash then enter groundwaters by
infiltration, receiving waters through direct ash pond discharge, and the
surrounding land area via runoff.
The results obtained by previous investigators do not provide a basis
for defining the geochemical processes ultimately responsible for
controlling the distribution of major, minor, and trace elements between
dissolved and particulate phases. This investigation was directed toward
identifying and describing important geochemical processes that determine
the metastable equilibrium conditions established in this complex
-28-
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heterogeneous system. Consequently, considerable emphasis was placed on
delineating the geochemistry of major (Al, Fe, Si) and minor (Ca, Mg, Na, K)
constituents of fly ash. These elements were chosen for study since they
represent major components of Columbia fly ash (Helmke et al. 1976, Talbot
1977). Furthermore, these elements are known to exert extremely important
controls on the trace element distribution of natural water systems
(Schindler 1967, Garrels and Mackenzie 1977, Stumm and Morgan 1970). A more
complete understanding of the heterogeneous processes involving these
elements should facilitate the interpretation of the trace element
distribution and behavior. Such information is required before the fate of
elements entrained in this system can be determined. An estimate of the
amount available for transport to uncontaminated natural ecosystems would
then be possible. This approach to studying the ash pond chemistry provides
a basis for interpreting similar data collected at other ash pond sites.
Although environmental conditions and ash composition vary, similar major
solid phases are expected to be formed. The relative proportions, and in
some cases the overall importance, of these various phases, are probably
influenced most by the parent ash composition.
RESULTS
The fly ash used in this investigation can be described as a collection
of heterogeneous particles that are principally spherical in shape. Each
ash particle is composed of a suite of elements contained in a glassy
amorphous matrix. Table 10 presents the total elemental concentration of
each Columbia fly ash sample. An empirical formula:
Si0.769Ca0.294A10.233M80.156Fe0.059K0.012Na0.009
was calculated using values obtained from total digestion of the hopper fly
ash. These fly ash samples thus appear to have an aluminosilicate
TABLE 10. TOTAL ELEMENTAL CONCENTRATIONS OF COLUMBIA FLY ASH
EXPRESSED AS PERCENTAGE OF DRY WEIGHT
Element Submerged Wet delta Dry delta Hoppers
Al
Ca
Cd
Fe
K
Mg
Na
P
Si
Zn
6.3
11.4
<0.001
3.0
0.46
3.4
0.26
0.12
23.0
0.0020
5.7
11.7
<0.001
2.7
0.40
3.4
0.34
0.13
20.6
0.0019
6.5
11.2
<0.001
3.2
0.43
3.5
0.28
0.12
22.4
0.00021
6.3
11.8
<0.001
3.3
0.45
3.8
0.21
0.11
21.6
0.00017
-29-
-------
structure. However, the presence of a well-defined mineral phase is
unlikely because the composition of the samples is complex. Other
investigaters have postulated that many trace elements are contained as
condensed species surrounding the ash particle (Davison et al. 1974, Linton
et al. 1976). The exact mechanism of attachment to the ash particle is
still unknown as is the chemical form(s) of the trace elements.
Once the fly ash particles contact an aqueous environment, solid phase
dissolution reactions occur immediately. Figure 6(a-i) illustrates the
dissolution as a function of leaching time for several major and minor
elements in fly ash. Sharp increases in dissolved aluminum, iron, and
silica concentrations are clearly indicated. In addition, the largest
amount of Al, Ca, Fe, Mg, Na, and Si is leached from the hopper ash which
had not been exposed previously to water. The dissolution of these elements
is probably enhanced by the rise in solution pH (Figure 7). Aluminum, iron,
and silica are particularly soluble at elevated pH values. The release of
these elements to the dissolved phase (Figure 6a-c) clearly concurs with the
dramatic increase in pH shown in Figure 7. Hydrolysis of oxide forms of Ca,
K, Mg, and Na are suspected to cause the initial pH increase. Shannon and
Fine (1974) attribute the high pH of water extracts collected during fly ash
leaching to the hydrolytic reactions of these elements; for example,
CaO + H20 » Ca"1"2 + 20H~ and (1)
Na20 + H20 = 2Na+ + 20H~ (2)
Analogous reactions may also be written describing the hydrolysis of
carbonate forms of these elements. However, magnesium oxides are
essentially inert (Cotton and Wilkinson 1972) and therefore are not expected
to be appreciably hydrolyzed. Magnesium entering the bulk solution at this
time is most likely associated with soluble calcium or magnesium compounds
other than the oxide.
All four fly ash slurries stabilized at nearly the same pH value after
1 week of leaching. This indicates that the processes controlling pH were
the same in all cases regardless of any previous leaching the ashes may have
undergone. Figure 8 shows that the same pH trend with time was observed
when the experiment was conducted under a nitrogen atmosphere to exclude
atmospheric C02 from the system. Several weeks elapsed before the pH
declined to lower values. The pH decrease in the open systems after 1 or
2 h apparently resulted from diffusion of atmospheric C02 into this highly
alkaline medium. Only 9x10 moles of C02 per hour would typically be
required in each instance to lower the pH from its higher value to its level
at 48 h into the experiment. This diffusion rate, however, only considers
neutralization of OH~ species. Since silica (H-SiO.) and other anionic
species (for example, Al(OH), may also accept H as the pH decreases, a
somewhat higher diffusion rate than calculated here might be required.
-30-
-------
SUBMERGED OFF DELTA
WET DELTA
DRY DELTA
HOPPERS
(a)
-LOGAI. [M]
t.o
-LOG Si. [M]
5.0
(O
-LOG Fe. [M]
-LOG Cd, [M]
-LOG P, [M]
I I
Ihr. Chr. 12hr. 24hr. 46hl 72he. 1wk. 2wk«. 1mo. «mo«. 6n»t.
TIME
Figure 6(a-e).
Change in dissolved elemental concentrations as a function of
leaching time. (continued)
-31-
-------
2.0
(f)
-LOG Ca, [M]
3.0
(9)
-LOG Wg,
-LOG K , [M]
(I)
-LOG Na, [M]
I I
Ihr. • hr: J2hr. 24ht 48hr. 72hr.
TIME
2wk«. Imo. 4mo«.
Figure 6(f-i).
Change in dissolved elemental concentraitons as a function
leaching time.
-32-
-------
S H > a.
D u cc o
CO $ O X
«<*>•»
I
CM
UJ
c E
•-
PL,
ca
CD
C
(1)
CO
CO
c
•H
a
to
cu
CU
P.
QJ
60
a
1^
cu
60
-33-
-------
-------
Although the actual amount of C02 consumed cannot be calculated, the
atmosphere can easily supply the necessary amount based on the above
estimation.
Aluminum and Silica —
The heterogeneous reactions involving aluminum and silica provide a
basis for understanding the metastable equilibrium conditions that determine
the stability of fly ash in an aqueous system. The release of aluminum and
1/2
silica to the bulk solution as a function of the square root of time (t ' )
is depicted in Figure 9. The change in dissolved concentrations of aluminum
and silica with time have been divided into three stages to aid the
interpretation of their heterogeneous reactions. The rate constants (K) for
dissolution of aluminum and silica from the fly ash during stage 1 can be
calculated from their respective slopes. The value of K was found to be
_c _i _ —
3.70 x 10 moles 1 i hour x/ for aluminum and 1.83 x 10 moles 1
hour ' for silica. Hence, aluminum enters the bulk solution approximately
twice as fast as silica. This ratio suggests that additional dissolution
reactions (for example, M20o + 5H20 = 2A1(OH), + 2H+) may be occurring for
aluminum in addition to those corresponding to dissolution of an
aluminosilicate phase.
The amount of aluminum compared to the amount of silica that
precipitates during stage 2 implies that an aluminum-rich phase must form.
The solubility of amorphorus aluminum hydroxide is exceeded 2.6 times as
stage 2 commences, which strongly suggests that aluminum is precipitating as
hydrous aluminum oxide or gibbsite. Co-precipitation and adsorption
processes most likely account for the concurrent decline in dissolved
silica. In addition, formation of an aluminosilicate mineral phase may be
occurring to some extent.
The dissolved silica concentration at t ' = 12.96 h or 3 days (log KSQ
= -4.93) is below that necessary (log K = -4.70) to maintain stability of an
aluminosilicate phase such as kaolinite with respect to gibbsite.
Therefore, the gradual decrease in dissolved aluminum during stage 3 is
attributed to formation of an aluminum oxide phase. This phase is probably
formed by precipitation from the bulk solution as the pH decreases, and
partly by the incongruent reaction:
Al2Si205(OH)4(s) + 5H20 = A1203.3H20(S) + H4Si04(aq) (3)
The gradual increase in dissolved silica agrees with this reaction. These
solid phase rearrangements are still apparently occurring after 6 months of
equilibration caused by the slow reaction kinetics of aluminosilicate
minerals and a downward pH trend favoring gibbsite stability. Hence, these
phases are expected to form in the ash pond if the pH is allowed to decrease
by decreasing or stopping the continuous input of fly ash.
-35-
-------
2
o
00
c
•H
,fi
O
t«
(U
O
O
3
o-
co
cu
a
O
O
§
n)
co
tfl
O
•H
•a
n)
I
•H
cd
T)
o
CO
to
S(H x Jajn /
oo
•H
P4
-36-
-------
The phases forming during stage 3 are also predicted by the plots of
dissolved aluminum and silica versus pH. These plots are shown in Figure
lOa and b. The abrupt decrease in dissolved aluminum between pH 4.5 and 8.5
clearly coincides with the pH region where hydrous aluminum oxide is most
stable (Parks 1972). In several samples from this pH region, micro-
crystalline gibbsite was identified by X-ray diffraction analysis. However,
the principle phase is probably amorphous aluminum oxide. The slopes of the
lines between pH 4.5 to 6.5 and 6.5 to 8.5 for aluminum in Figure lOa (that
is, -3.3 and +2.0, respectively) are much steeper than would be expected
from equilibrium relationships of hydrous aluminum oxide with corresponding
aqueous aluminum species. This result implies that Figure lOa demonstrates
only the presence of an aluminum phase in the pH region where aluminum is
virtually insoluble in an aqueous system. If equilibration times at each pH
value had been longer the pH dependence of aluminum would most likely be
more consistent with that of a hydrous aluminum oxide system.
The decreasing dissolved silica concentrations between pH 12.3 and 4.3
(Figure lOb) may reflect the specific adsorption of silica onto the hydrous
aluminum oxide phase, since this behavior is not expected by reaction (Eq.
3). Thus, this apparent adsorption may suppress the dissolution of silica
if the ash particles encounter natural waters. Silica in the bulk solution
below pH 4.3 concurs with the disappearance of the aluminum phase. Silica
that enters the dissolved phase below pH 4.3 is apparently derived solely
from the aluminosilicate structure of the fly ash. The nearly constant
molar ratio of dissolved aluminum and silica, below pH 4.3 (Al)/(Si) = 1.10,
is further evidence that only the aluminosilicate structure of the fly ash
is dissolving.
The existence of an aluminosilicate phase over the entire pH range was
confirmed by X-ray diffraction analysis. Unfortunately, these diffraction
patterns are not distinct enough to allow deduction of an exact composition
or structure. Since the Al/Si ratio in the original fly ash samples most
likely includes aluminum and silica associated with other phases (for
example, amorphous silica and aluminum oxides), the dissolved aluminum and
silica concentration at pH 4.3 was used to estimate their ratio in the
parent aluminosilicate phase. This pH value appears to represent a pH
region where incongruent dissolution or adsorption reactions do not alter
the aluminosilicate solubility. The calculated Al/Si molar ratio of 1.0 is
similar to the ratio that would result from the congruent dissolution of a
mineral phase similar in composition to kaolinite. If this aluminosilicate
phase is represented by kaolinite, its solubility can then be calculated
according to the reaction
Al2Si205(OH)4(s) = 6H+ = 2A1+3 + 2H4Si04(aq) + H20 (4)
The solubility relationship of this phase at equilibrium can be expressed
as :
(Al+3)2 (H.SiO,)2
KSO -- +*6 -
-37-
-------
SUBMERGED OFF DELTA
WET DELTA
DRY DELTA
HOPPERS
(a)
-LOG Al , [M]
-LOG Si , [M]
-LOG Fe, [M]
-LOG cd, [M]
10 1« 4.0 SD U 7.0 10 U TOO 11.0 12.0
PH
Figure 10(a-d).
Dissolved elemental concentrations as a function of pH.
(continued)
-38-
-------
2.0
(e)
-LOG Ca,
3.0
(0
-LOG Mg, [M]
(9)
-LOG Na, [M]
s.o
6.0
-LOG P, [M]
—«••-
•
•\,
2.0 3.0 4.0 5.0 6.0 7.0 8.0 9.0 10.0 11.0 12.0
PH
Figure 10(e-h). Dissolved elemental concentrations as a function of pH.
-39-
-------
where log KSO = 11.28 (Hem et al. 1973). Using dissolved aluminum and
silica concentration at pH 4.3, the free energy of formation for this
aluminosilicate phase was calculated to be -902.4 kcal/mole (25°C). Robie
and Waldbaum (1968) report the standard free energy formation for kaolinite
as -902.9 kcal/mole (25°). The agreement of these free energy values is
certainly not conclusive proof that this aluminosilicate phase is kaolinite,
but it does suggest that the phase is compositionally similar to kaolinite.
Figure 11 demonstrates that this system is not in equilibrium with
respect to K - aluminosilicate phases. Although not shown here, the same is
true for Na - aluminosilicate phases (Talbot 1977). The low abundance of K
and Na compared to Al, Si, and Ca makes the interpretation of these elements
difficult. The heterogeneous processes of K and Na are overshadowed by
reactions involving more prevalent phases. However, the pH dependence of
potassium and sodium is consistent with that exhibited by feldspar or mica-
type phases (Figures lOg, 11, and 12). Figure 11 indicates that the
solution proceeds toward equilibrium with either feldspar or mica-type
phases. The reaction path followed by potassium indicates that equilibrium
between a feldspar and mica phase would have been established with a
logarithmic K /H+ molar ratio of 7-8. The exact nature of these phases
remains uncertain, since X-ray diffraction analysis did not reveal their
composition.
Dominant heterogeneous processes involving aluminum and silica probably
causes the K /H ratio to decrease shortly after the leaching began (Figure
11). The movement of the K+/H+ ratio into the gibbsite stability field
coincides with the apparent incongruent dissolution of the aluminosilicate
phases described previously (Figures 7 and 10, a and b). Eventually the
bulk solution begins to re-establish metastable equilibrium with the
feldspar or mica-type phases. The attainment of metastable equilibrium is
indicated by the increasing K+/H+ ratio, but the reaction rate and
subsequent drop in pH are extremely slow owing to the apparent control by
incongruent dissolution processes.
Iron—
Iron is also considered a major component of this hetergeneous
system. Figure 13 shows the dissolved iron concentration as a function of
the square root of leaching time. Iron may be present initially on the fly
ash as complex oxides (Brimblecombe and Spedding 1975) and possibly as
carbonates, sulfates, or carbides (for example, Fe^C^). These various iron
compounds are rapidly dissolved, probably through hydrolysis reactions.
Subsequently, iron precipitates since the solubility of hydrous ferric oxide
1 /?
is exceeded 2.6 times. The precipitation rate then decreases as t '
increases. The slope of dissolved iron versus pH, plotted in Figure lOc,
also indicates that hydrous ferric oxide exists over a wide pH range. X-ray
diffraction patterns indicated that microcrystalline hematite was present,
especially in the mid to high pH range.
A computer program was used to calculate the equilibrium constants
-40-
-------
AMORPHOUS
SILICA SAT.
LOG H4Si04, (M)
Figure 11. Silicate stability diagram with data points for K+/H+(wet delta
ash) from the leaching experiments superimposed. Numbers are
in consecutive order with respect to sampling intervals (short
to long equilibration times) during the leaching experiments.
-41-
-------
3
CM
O
0
O
0>
s
Q.
2
O
in
en
en
cfl •
4-J 4J
o ti
& cu
^ B
HI Cfl
O tl
w cfl
CO
•H pc!
a »-4
•H o
M-J
cu
60 CU
o o
(4
CU T3
CN
3
DOT -
60
•H
-42-
-------
o
to
o
IA
oo
.5
-------
corresponding to the hydrolytic reactions of hydrous ferric oxide with
pertinent aqueous iron species. The program estimated the equilibrium
constants (K^) from a non-linear least-squares fit of experimental data
points for dissolved iron versus pH. The calculated K values are presented
in Table 11. They are referenced to 25°C and zero ironic strength (uobs. =
0.007). The free energy of formation (AG°) for this hydrous ferric oxide
TABLE 11. EQUILIBRIUM CONSTANTS DESCRIBING FERRIC HYDROXIDE SOLID
PHASE EQUILIBRIA WITH MAJOR AQUEOUS FERRIC COMPLEXES
Species
Fe(OH)o ,„)
-J \ ** /
Fe(OH)+2
2
Fe3(aq)
Fe(OH)4
Log I^3
39.50
13.30
22.70
31.00
33.50
Estimated log K^
39.50
13.37
22.70
32.66
34.56
aThe reaction corresponding to K^ is Fe+3 + nOH = Fe(OH)3-tl,
From Novozamsky et al. (1976).
phase was calculated to be -169.3 kcal/mole (25°C) from the estimated KQQ
value. The standard free energy of formation for hydrous ferric oxide at
25° is -166.0 kcal/mole and for hematite -177.1 kcal/mole (Carrels and
Christ 1965). Thus, the iron phase formed in this system appears to be
partly microcrystalline in nature. At pH values above 3.0 the ferric oxide
phase most likely controls the dissolved iron concentration.
Calcium and Magnesium—
The proposed relationships between calcium and magnesium in this system
are shown in Figure 14. The distribution of the various phases was derived
from basic equilibrium expressions listed in Table 12. The data points for
calcium and magnesium fall directly on the phase boundary between calcite
and dolomite. Hence, dissolved calcium and magnesium concentrations seem to
be controlled exclusively by carbonate phases even at pH values where these
solids are thermodyamically unstable.
At pH values greater than 10.5 magnesium precipitates, most likely as
brucite (Figures lOd and 14). Consequently, calcium is released to the
solution by the mixed carbonate phase due to mass balance constraints. The
bulk solution then contains calcium in excess, and calcite (or argonite)
apparently forms on the ash surface as described by Bricker and Garrels
-44-
-------
* SUBMERGED
* WET DELTA
• DRY DELTA
* HOPPERS
20 •
15
03
o
Q.
a
CM10
Dissolved
Ca2 and Md*2
(6.4)
(2.0)
10
15
20
2pH-pMg
Figure 14. Solubilities of Ca+2 and Mg+2 carbonates and hydroxides at 25°C
Numbers in parenthesis correspond to pH values observed during
the electrophoretic mobility experiments
-45-
-------
(1967). The presence of calcite at high pH was verified by X-ray
diffraction analysis.
Phosphorous and Cadmium—
Hiosphorus and cadmium were used to model the behavior of trace
constituents typical of fly ash. Hiosphorus was chosen owing to its
importance in natural water eutrophication processes and cadmium because of
its acute toxicity to aquatic organisms. Results of these experiments
indicate that the trace element distribution between dissolved and
TABLE 12. LOG EQUILIBRIUM CONSTANTS AT 25°C AND LOG C02 = -3.52
Reaction
Log K
Source
CaCo3 = Ca
C0
~
MgC03 = Mg+2 + C0~2
- 8.35 Langmuir (1968)
- 8.00 Garrels and Mackenzie (1967)
Robie and Walbaum (1968)
CaMg(C03)2 = CA"1"^ + tig"1"-* + 2C
03 -16.70
Mg4(C03)3(OH)2 = 4Mg+2 + 3CO~2 20H~ -34.90
Ca(OH)
Mg(OH)
H20 =
H20 +
HCO~
2 = Ca+2 + 20H~
2 = Mg+2 + 20H~
H+ + OH~
CO 2 = HCO~ + H+
= H+ + C0~2
- 5.43
-11.15
-14.00
- 7.82
-10.33
Stumm and Morgan
(1970)
Garrels and Mackenzie (1967)
Stumm and Morgan
(1970)
Garrels and Mackenzie (1967)
Schindler (1967)
Schindler (1967)
Schindler (1967)
particulate phases is strongly influenced by the major solid-phase
components of this system. Adsorption and precipitation reactions appear to
be controlling mechanisms in the appropriate pH regions.
The most important isoelectric point (pHjgp) of the Columbia fly ash,
with respect to natural water systems, occurs in the mid pH region near pH
7.55 (Figure 15). Essentially 100% of the aluminum and iron are associated
with various phases on the ash particle in this pH region (Talbot et al.
1978). This suggests that the surface characteristics of the ash are
controlled by these incipient phases (e.g., A1(OH)3 and Fe(OH)3). Figure
lOh demonstrates that almost all phosphorus (99.8%) is removed from solution
as the pH decreases from 8.5 to 4.0. Huang (1975) observed a similar pH
dependence for phosphorus adsorption on amorphous aluminum hydroxide. The
-46-
-------
1 1 1 1 1
,<§> * •
.
^
*
• ^
• *
<* »
.^
* *
.
uj S 2
gr ul Ijj S
U O ^J ui
|Q •" ^ AB
3 HI CC O
~ CO S> ^ ^C
*•*
-
• *
#
iii iii
III
-
.
-
A
*
•
B *
* *
** *
<&
.g.
•
-------
reason for the slight variation in dissolved cadmium concentrations in the
same pH region may be caused by the heterogeneus nature of the fly ash.
The appearance of both cadmium and phosphorus in the bulk solution at
low pH results from ash particle dissolution. Only 5% of the phosphorus
remains on the ash particle at pH 1.0. In contrast, approximately 90% of
the cadmium is still associated with the ash particle. Cadmium is
apparently strongly adsorbed to the ash surface at low pH values. Trace
anions, however, seem to be almost entirely released to the dissolved
phase. Hence, adsorption reactions appear to completely control the
dissolved trace element concentrations at low pH. Further investigation of
the release of potentially toxic anions such as arsenic and boron is
probably warranted.
At the high pH values observed in the ash pond (Andren et al. 1976,
1977) calcium and magnesium carbonate and hydroxide phases may control the
surface characteristics of the fly ash. In addition, hydrous ferric oxide
may also partly determine the identy of the ash surface at elevated pH
values. The carbonate phases of calcium and magnesium do not have large
adsorption capacities or strengths for trace metals (larekh et al. 1977).
Nevertheless, the respective carbonate and hydroxide solubilities of the
trace metals should keep their dissolved concentrations low. For instance,
at pH 11.0 the cadmium concentration is about 10~9 . Since Kg = (Cd+2)
(OH~)2, we have (10~9) (10~3)2 = 10~15. The reported Kgp for Cd(OH)2 is
4x10 (Sturam and Morgan 1970). Similarly, assuming C02 saturated
conditions, K for CdC03 is 5xlO~12. These calculations suggest that the
carbonate and hydroxide phases are controlling mechanisms for dissolved
cadmium at high pH. The same is probably true for other chemically similar
transition metals.
The simultaneous decrease of dissolved calcium and phosphorus at pH
10.0 (Figures 10, e and h) implies that phosphorus precipitates as a
hydroxyapatite-type phase or is removed by occlusion via co-precipitating
magnesium calcites. Using hydroxyapatite as an example, we can calculate
the corresponding ion activity product according to the reaction:
5 Ca+2 + 3P04~3 + OH~ = Ca5 (P04)3 OH^ (6)
Then, Kgp - (Ca+2)5 (OH~) (P04~3)3 where Ksp = 10~55'6 (Stumm and Morgan
1970). Substituting values obtained from plots of dissolved concentration
_• O
and estimating the dissolved PO/ concentration by:
_3 7.8xlO~13[HPop
* r TT ' i
-48-
-------
_O _Q
the concentration of PO^ is 1.2 x 10 . llie ion product for
hydroxyapatite can be computed by:
(10~3*6)5 (Iff3) (10~7*9)3 = 10~45 at PH 11 .
This apparent oversaturation may reflect the influence of the chemical
composition of the aqueous phase (Corsaro and Sutherland 1967), kinetic
factors (Jenkins et al. 1971), and hindrance by magnesium on nucleation and
growth (Berner 1975). Although specific adsorption of phosphorus can
substantially decrease its dissolved concentration (Malotky 1978), it
appears to have minimal influence here. Further investigation of the
controls on the trace element distribution is needed before a rigorous
explanation of these processes can be attempted.
MONITORING STUDY OF THE COLUMBIA ASH BASIN
In addition to the laboratory leaching experiments, a monitoring study
of the ash basin at the Columbia Generating Station was conducted from June
1976 to April 1977. The study had two objectives: (1) To identify
potential environmental hazards associated with fly ash leachates and to
compare these to the results of the laboratory studies, and (2) to
characterize the surfaces of suspended sediments in different parts of the
ash basin.
The ash basin is normally well mixed with regard to chemical and
thermal stratification. Except for ice cover during the winter of 1976-77,
surface and bottom water samples exhibited nearly identical chemical
compositions.
Variations in dissolved elemental concentrations typically demonstrated
horizontal gradients (from the first ash basin to the second) when the plant
was operational. Dissolved chemical concentrations were a function of flow
rate and therefore presumably were controlled by kinetic constraints since
many of the elements were often supersaturated.
Figures 16 and 17 summarize the data obtained for the monitoring period
for average major element concentrations in the ash basins and source water
(the cooling pond and the Wisconsin River). High pH values are associated
with this fly ash leachate (see previous discussion). Extremely high pH
values (>12) occurred after December 1976 when ^2^0^ was added to the fly
ash to improve the efficiency of the electrostatic precipitators. Average
concentrations of silica and alkalinity were also greater in the ash basins
than in the source waters. Their concentration trends behave similarly to
pH after the addition of Na^COo commenced. In the case of silica one would
expect greater solubility of silicic acid in the pH ranges present in the
ash basin.
Phosphorus concentrations in the ash basin were normally less than the
-49-
-------
O WISCONSIN RIVER
A COOLING LAKE
D ASH BASIN
RATION
1-
Z
UJ
O
Z
o
o
DC
<
O
5
CD
O
_j
-12
-11
-10
-9
-8
-7
-1
AIK -2
-3
Si ~4
-5
-5
TOT
_e_
' Qd
^
-7-
D ° D '
n D
D
^ D D
D D
-
A A
A O 6 A o
A • ° °
n
r— i _ iC CBW A
A n H '8' Q®AS©
HOD
n H Q n g
n9gnQn"^
A A
A
O o °
I^S DO a2 ®A
° A B A ._, U
n
n D n D
JJASONDJFMA S
1976 1977
Figure 16. Monthly comparison of Wisconsin River, cooling lake, and ash
basin analyses.
-50-
-------
O WISCONSIN RIVER
A COOLING LAKE
D ASH BASIN
z
0
1-
<
QC
1-
z
LU
0
z
o
cc
1
^J
o
o
o
_J
-3
Ca
-4
-3
Mg-4
-5-
-1
Na -2
-3
-3
K
-4
n D D D D n
AA&AA@oA^QQ Q
° D D
AriOO® 0QOO
0 D ODD n
n D D
n — n
n n
a a
D D
D
n _
^ 9 Q Q o o A
@ §§8aB8 g -
J JASON
1976
D J F
M A
1977
Figure 17. Comparison of average Na, K, Mg, and Ca concentrations in the
Wisconsin River, cooling lake, and ash basin.
-51-
-------
concentrations found in the cooling pond. This indicates that phosphorus
(and organic matter) is being adsorbed on the fly ash particles deposited on
the ash delta. On several occasions anomalously high values were
observed. However, these high phosphorous values were usually associated
with abnormal turbidity during periods of high wind action. Previous
studies (Tenney and Schelberger 1970, Higgins et al. 1976) have demonstrated
the adsorptive properties of fly ash.
Figure 17 illustrates the concentrations of sodium, potassium, calcium,
and magnesium in the ash basins and cooling pond at the Columbia station.
Elevated sodium and potassium concentrations are expected since hydrolyses
of their respective oxides are attributed to be one of the reactions leading
to an elevated pH (see previous section). The sodium concentrations reach
exceptionally high concentrations in December, again resulting from the
addition of Na^COo. Concentrations of magnesium are usually below those
observed in the source waters. Previous calculations have demonstrated that
the solubility product of brucite (Mg(OH)2) is exceeded and the data suggest
that the element is retained in the ash delta as the water percolates
through the deposited fly ash. The profile after December 1976 is
completely opposite to the pH trends observed in Figure 16. The behavior
displayed by calcium demonstrates that average levels of dissolved calcium
were much higher than the source water prior to December 1976 and that after
this period calcium was also precipitating and being retained in the fly ash
delta.
Figure 18 illustrates average iron, aluminum, copper, and chromium
concentrations in the ash basins. Iron is normally near the detection limit
although high values are observed in March and April 1977. These high
values can possibly be explained as an artifact of filtration since
colloidal iron may pass through a 0.4-pm filter. Ash basin aluminum
concentrations exceeded source water concentrations by a magnitude of 100 to
1,000 at all times during the study. This is consistent with the laboratory
experiments. The concentrations of copper and chromium demonstrate that
both elements are liberated after interaction of cooling pond water with fly
ash. The elevated copper concentrations are not consistent with laboratory
equilibrium experiments and may be due to kinetic factors.
Table 13 demonstrates the results obtained from the M1NEQL computer
program for calculating thermodynamic equilibrium concentrations. Only
species present at greater than 1% of the total concentration are
reported. For these calculations the pH was held at 11.5 and redox
reactions were not considered. Many of the elements are predicted to remain
in solution as soluble complexes ; several of the cations have soluble
hydroxy complexes (Al, Fe, Cr, Cd, and Pb) and presumably would exist in
solution as anionic species. The computer results suggest that silica would
not precipitate. However, soluble silica decreases in the ash basin when
pumping is interrupted, and zeolites (aluminosilicates) have been identified
in the ash delta. (Helmke et al. 1977). Solid phases predicted by the
program include CaC03, A1(OH)3, Mg(OH)2, Fe(OH)3, Cu(OH)2, Ba SO^, and
» (Several of these solid phases were identified in the previously
-52-
-------
O WISCONSIN RIVER
A COOLING LAKE
D ASH BASIN
Fe
z
0
1-
DC
1-
•^»
LU Al
O
z
0
o
^^—
cc
_J
o
2 Cu
O
O
Cr
-4
-5
-6
-4
-5
-6
~7~
-8
-6
-7
- o o ° o
A A A *
---a--o---a— -a--
i i
r^ D r^
D D
D
- 00 &
A ®
Cyi c^3
D °
^ y a A
o o 0
1 d
- D a a
1 1 1 1 —
a a
o
0
A
— - A- S—
i i ,
D D
O
A ®
A
2 D
A
/^v
° 0
D
a a
-i 1 1
N D
1976
J F
M A
1977
Figure 18. Comparison of cooling lake and Wisconsin River dissolved metal
concentrations with those observed in the ash basin.
-53-
-------
TABLE 13. MINEQL COMPUTER PROGRAM RESULTS FOR SOLUBLE CONCENTRATIONS
OBSERVED DURING THE SEPTEMBER 1977 SAMPLING PERIOD (25°C)
Initial
Element concentration3
C03 2.6xlO~3
Cl 4.2xlO~4
H4Si04 1.5xlO~4
Al l.lxlO'4
S04 1.2xlO~2
Na 3.1xlO~2
Ca 1.3xlO~3
Mg 3.0xlO~5
K l.lxlO~4
Fe 8.0xlO~7
Cu 3.8xlO~7
Soluble
species
C03=
NaCO~
HCO~
ci-
H3SiO-4
H2Si°4
H4Si04
A1(OH)~
4
NaSO~
Na+
NaSOT
4
NaCO~
MgAl(OH)+
MgS04
K+
KSO~
Fe(OH)~
%
36
12
1
100
96
3
1
7
90
10
95
4
1
2
1
91
9
26
Precipitate %
CaC03 51
A1(OH)3 93
CaC03 98
Mg(OH)2 95
Fe(OH)3 74
Cu(OH)2 100
(continued)
-54-
-------
TABLE 13 (continued)
Initial
Element concentration3
Ba 4.4xlO~6
H4B04 4.0xlO~4
Cr 1.9xlO~6
Cd 2.5xlO~8
Pb 4.8xlO~9
Soluble
species
H4B04
Cr(OH)~
Cd(OH)~
Cd(OH)2
Cd(OH)+
Pb(OH)~
Pb(OH)2
% Precipitate
BaS04
100
100
64
34
1
61 Pb(OH)2
10
%
100
29
-55-
-------
discussed laboratory solubility studies.)
The limitations of this computer model are based primarily on the
completeness of the number of reactions and the accuracy of the formation
constants. Additionally, the model implies thermodynamic equilibrium, a
condition obviously not found in the ash basins. However, the model is
useful for predicting possible solid phases and represents a first attempt
at understanding the chemistry of the system. As stated previously, kinetic
constraints cannot be handled and adsorption reactions are not included in
the calculations.
Table 14 compares the range of potentially toxic elemental
concentraions observed in the ash basin with literature values for water
quality criteria. With the exceptions of lead and arsenic, all the elements
listed in the table could have a deleterious effect on the quality of
receiving waters. Aquatic life is normally absent in the ash basin, and
these concentrations, in conjunction with the extremely high pH values
recorded in the basin, may provide one explanation for this observation.
TABLE 14. COMPARISON OF SUGGESTED WATER QUALITY CRITERIA WITH
VALUES OBSERVED IN THE ASH BASIN
Concentrations
possibly toxic Range observed
to aquatic life in ash basin
Element (mg/liter) (mg/liter)
Al
As
B
Cd
Cr
Cu
Pb
Zn
1.0a
0.5b
0.22b
0.0001b
O.lb
0.01b
0.030b
0.030b
0.02 -
0.006 -
0.1
0.0001-
0.066 -
0.010 -
<0.002
<0.01 -
52.6
0.216
6.4
0.004
0.142
0.028
0.04
aResources Agency of California (1960),
bVan Hook and Shutts (1976).
Figures 19 and 20 indicate that the rate of aluminum and chromium
precipitation is enhanced between the ash basin and the discharge water.
This enhancement results from the present practice at the Columbia
Generating Station of neutralizing the ash basin effluent with sulfuric
acid. Despite partial removal of some elements (Al and Cr), concentrations
are sufficiently high to be viewed as potentially toxic in receiving waters
composed of a significant fraction of ashpit discharge. It may be
advantageous, using current facilities, to neutralize ash basin effluent
-56-
-------
September 13, 1977
As
Al
mfl/i
Cr
mg/i
Cu
mfl/|
.20
.15
.10
.05
.50H
.10
.05
.025 H
Cd
mg/ .002-
COOLING ASH
POND DELTA
1st 1st 2nd 2nd
BASIN BASIN BASIN BASIN
DISCHRG
DISCHARGE
Figure 19. Soluble element concentrations at various stations at the
Columbia plant for September 13, 1977.
-57-
-------
As
mg/,
Al
Cr
mg/,
Cu
Cd
mg/,
B
mg/,
.200-
.150-
.100-
.050-
30-
20-
10-
.10-
.05-
.005-
.004-
.002-
5-
COOLING ASH 1st
POND DELTA BASIN
1st 2nd
BASIN BASIN
DISCHARGE
2nd DISCHARGE
BASIN
Figure 20. Soluble element concentrations at various stations at the
Columbia plant for September 27, 1977.
-58-
-------
prior to the final settling basin. This process may facilitate
sedimentation of flocculated aluminum with subsequent partial removal of
several additional toxic elements.
The second objective of the study of the Columbia ash basin was to
characterize the surfaces of the sediments in the ash basins. Suspended
sediments, collected in sediment traps during the September 1977 sampling
period, were examined by scanning electron microscopy (SEI1) and X-ray
diffraction.
Figure 21 is a (SEM) picture of fly ash particles blown on shore in the
second fly ash basin. The characteristically spherical fly ash particles
serve as a nucleus for precipitation reactions in the basin. Figure 22
represents a picture of precipitate removed from the float of a sediment
trap in the second basin. X-ray emission and X-ray diffraction analyses of
this sample suggest that it is composed primarily of calcite. The crystal
structure is not the normal cubical crystal observed for this mineral phase;
instead this crystal has a structure similar to calcite containing trace
quantities of magnesium (Berner 1975).
X-ray diffraction analyses were made of sediment samples obtained in
the first and second ash basins and of the windblown precipitate and float
precipitate pictured in Figures 21 and 22. Qualitative surface X-ray
emission analyses of all samples demonstrated the presence of Al, Si, S, K,
Ca, Mg, Ba, and Fe in varying degrees. However, Ca usually predominated
(particularly in the sediment float sample).
Figure 23 illustrates the diffraction peaks observed in the windblown
precipitate. Calcite (CaCO.j), quartz (Si02), and gypsum (CaS0^.2H20) peaks
can be readily observed. Barite (63804), magnetite ^6304), apatite
(Ca5(PO^)3(F,Cl,OH)), and potassium chabazite (CaA^Si^O^. 6H20) may also be
present although overlapping peaks tend to obscure the relative X-ray
intensities characteristic of the respective minerals. Additional peaks have
not been identified at this time.
,Figure 24 depicts the X-ray diffraction peaks observed in the
precipitate that was scraped off the walls of the sediment trap float
positioned in the second ash basin. Evidence for the presence of calcite
(CaCOo) and magnetite (FeoO,) is apparent although peaks for the other
mineral phases discerned in Figure 23 were not observed. Of particular
significance is the absence of the quartz peaks present in all the other
diffraction analyses. This suggests that quartz in the sediments can be
attributed to resuspended shore material or that quartz was originally
present in the discharged fly ash. Many of the peaks present in this figure
have not been identified.
Figures 25 through 28 display the results obtained from X-ray
diffraction analyses of suspended sediments collected by traps in the first
basin (near the ash delta) and the second basin of the ash pond. The
figures indicate that the sediments become more complex (with additional
-59-
-------
Figure 21. Scanning electron microscope picture of windblown precipitate
from the shore at the second basin of the Columbia ashipt
(magnification = 100) .
Figure 22. Scanning electron microscope picture of the precipitate formed
on the sediment trap float in the second basin of the Columbia
ashpit (magnification = 100).
-60-
-------
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-66-
-------
mineral phases) and Increasingly crystalline with distance from the initial
deposition site. Both patterns demonstrate the presence of quartz, gypsum,
and calcite although a magnetite peak is absent from the analyses of
sediment from the first basin. Additional peaks remain to be identified.
Groundwater Contamination From The Ash Basin
Contamination of groundwater by fly ash leachate has been documented by
Theis and RLchter (1978). The authors presented evidence which suggested
that equilibrium groundwater concentrations could be partially predicted
utilizing current adsorption models. This model was run for elemental
concentrations in the presence of various metal oxides of a relatively well-
characterized soil. However, the present literature dealing with the
movement of contaminants through various soild systems suggests that several
mechanisms (that is, other than adsorption) may be responsible for
controlling elemental mobilities in groundwater. Korte et al. (1975)
demonstrated that the mobility of several transition metals is highly
dependent on the chemical form of the element and the iron and manganese
content of the soil. Thus, some elements may be partitioned by adsoprtion
on the various soil components (that is, organics, clays, or hydrous
oxides), and other elements may precipitate. Regardless of whether
adsorption or precipitation constitutes the primary removal mechanism, one
may predict that a given soil volume eventually becomes saturated (given a
constant input). The movement of chemicals through groundwater systems can
thus be viewed as a natural chromatographic process with the more
conserative elements traveling faster than transition metals. This
phenomenon could be exploited to diagnose whether groundwater contamination
will be a problem. That is, if elements such as Na, Cl, or SO/ do not
appear in the vicinity of fly ash disposals, the more toxic elements such as
Pb, Cd, Zn, and Cu would also be immobilized.
Groundwater infiltration at the Columbia ash basin is not well
understood. Some ash basin water percolates into the lower soil profile
(Anderson and Andrews 1976). There is also evidence, however, that ash
basin water infiltrates directly through the dikes and remains perched on
the surface of the adjacent sedge meadow. This infiltration may provide one
explanation for the difficulty in providing conclusive evidence of ash basin
contamination in water samples taken from wells placed at incremental
distances from the ash basin. Some of the wells may intersect contaminated
sub-surface water flows while others may not.
Some evidence, however, suggests that ash basin leachate is influencing
the water quality of the surrounding groundwater. For example, well-water
samples obtained prior to plant operation (April 1975) demonstrated boron
concentrations of less than 0.09 mg/liter. Subsequent analysis of samples
demonstrated concentrations ranging from <0.09 mg/liter to 0.82 mg/liter
(analysis performed by Swanson Fjivironmental, Inc.). Analysis of our well-
water samples obtained in September 1977 (Table 15) suggested boron
concentrations of approximately 0.3 mg/liter. Boron may be present in the
ash basin at concentrations in excess of 5 mg/liter (Table 14) and is
considered relatively inert with regard to precipitation reactions (Hem
-67-
-------
1971). In addition, the current literature suggests that the element is
relatively mobile in the soil column (Braustein et al. 1977). Thus boron
contamination of groundwater is apparently occurring around the ash basin at
the Columbia Generation Station. The situation is far from well understood,
however. Analyses of the well-water samples obtained in September 1977
showed elevated concentrations of Na, Mg, Ca, As, Al, SO,, Cl, and Si in the
well closest to the dike. However, Mg concentrations in the ash basin were
lower than in any water system in the area. This phenomenon is most likely
caused by leaching from the soil itself. Data in Table 15 also suggest that
gradients exist for SO/, Ca, Mg, and Na concentrations in four progressively
distant well-water samples obtained north of the ash basin. However, K
concentrations displayed a reverse gradient, and the highest boron
concentration was found in the well farthest from the ash basin. The data
from this project indicate that groundwater concentrations are neither
spatially nor temporally consistent with ash basin concentrations.
Future modfeling of this system will require more detailed information
regarding sub-surface water movements around the ash basin in addition to
better soil characterization. Any sampling must also be done in conjunction
with detailed records of plant operations since interruptions in ash basin
flows strongly influences infiltration into the soil column.
TABLE 15. CHEMICAL CONCENTRATIONS OF SELECTED DISSOLVED COMPONENTS
IN GROUNDWATER SAMPLES FROM THE ASHPIT DIKE AND THE SEDGE MEADOW
Top well
Middle well
Bottom well
Sedge meadow
SO,
(ppm)
36
10
20
18
As
(ppm)
<0.001
<0.001
<0.001
<0.001
B
(ppm)
0.3
0.3
0.3
0.7
Ca
(ppm)
77.0
60.5
42.5
9.6
Cr
(ppm)
<0.005
<0.005
<0.005
<0.005
K
(ppm)
1.6
2.6
3.4
3.6
Mg
(ppm)
49
37
22
11
Na
(ppm)
6.2
5.0
1.0
1.9
68
-------
SECTION 7
ENVIRONMENTAL IMPACT ON NATURAL WATER SYSTEMS
In considering sources, mechanisms of transport, and sinks of materials
in the environment, simple input/output mass balance models are often
useful. Such simple models can offer initial insights into the general
behavior of an element and often establish a framework for subsequent
research. In the case of the cooling pond, we felt that the most useful
information from such an exercise would be data on the amount of material
that accumulates. It must be kept in mind that several blowdowns occurred
during the sampling period. The overall effect from such efforts on mass
balance calculations would tend to decrease the residence times of elements
in the cooling pond, especially for the dissolved components. The
interpretation of the calculated fluxes should thus be made with some
caution.
Fly ash particles serve as the foremost mode of transport for trace
elements entering the environment as a result of coal combustion. Moreover,
trace metals are preferentially concentrated on the smallest fly ash
particles (Natusch and Wallace 1974; Natusch et al. 1974). This
concentration is of obvious concern for those respirable size particles
released to the atmosphere, but the effects and transport of micron-size
particles in an aquatic system are not as well documented. These small ash
particles probably constitute a major portion of the suspended particulate
matter in ash ponds. The following section presents a mass balance for
selected elements in the cooling basin and a general discussion of the
effects of fly ash discharge on receiving waters.
CHEMICAL MASS BALANCES OF THE COOLING POND AND ASH BASIN
Table 16 presents a mass balance for several elements in the cooling
pond (including suspended particulate matter). Calculations were based on
averaged data from neutron activation analysis of suspended sediments (Table
17), plant pumping records (Table 1), and values obtained from dissolved
element analysis (see data acquisition report). It was assumed that
leaching of solid phase elements into the groundwater represents an
insignificant loss to the system except for K, Na, SO^, and Cl. Riase
partitioning of As, Ca, and Sb also indicate that these elements are
amenable to groundwater transport. Because CaCOg precipitation is such a
strong function of Ca concentration and temperature (see previous
discussion), it is very difficult to predict how much of this will be
transported away via groundwater.
-69-
-------
TABLE 16. MASS BALANCE FOR SELECTED ELEMENTS IN THE COOLING POND
Element
Al
As
Ba
Ca
Co
Cu
Fe
K
Na
Sb
Si
Zn
S04
P04
Cl
Earticulate
matter
Input
(kg/yr)
22,600 (5)a
104 (95)
800 (66)
565,000 (99)
10 (20)
34 ( 8)
13,600 (23)
40,800 (93)
220,750 (99)
6 (94)
80,000 (32)
320 (25)
300,000 (99)
16,000 (14)
280,000 (99)
270,000
Amount in
pond (kg)
2,380
22
125
136,000
2
36
1,800
8,570
46,360
1
10,100
34
63,000
290
58,800
56,700
Output
(kg/yr)
4,760
44
232
271,000
4
73
3,600
17,100
92,700
3
23,500
67
126,000
580
117,000
84,700
Percentage
Groundwater remaining
outflow(kg/yr) in pond
79
58
71
52
60
—
74
14,600 22
84,700 20
50
71
79
115,000 16
96
110,000 19
70
aNumbers in parentheses represent the average percentage "dissolved"
concentration, (that is, <0.4 m).
The model indicates that the cooling pond acts as a repository for the
major fraction of elements associated with solid phases. Except for the
nonreactive elements (Na, K, SOA, and Cl), the cooling pond retains 70%, or
more, of Al, Ba, Fe, Si, Zn, and PO/. The calculations also suggest that a
significant amount of Cu is added to the system from plant activities
(39 kg/yr). This addition is presumed to result from the in-plant cooling
system. Inspection of Table 14 indicates that some elemental enrichment of
particulate matter occurs between the cooling pond intake and plant
outflow. That is, particulate matter at the plant outflow exhibits higher
concentrations of many insoluble elements. We speculate that chlorination
and heating partially destroys the organic matrix of the particulate
material.
The mass balance calculations derived from cooling pond input and
output elemental concentrations indicate that: (1) Copper is released from
the plant in signficant quantities; (2) most cations (except Na and K) are
effectively retained by solid phase association; (3) suspended particulate
matter and bottom sediuments are efficient in maintaining relatively low
soluble metal concentrations; and (4) the majority of P is accumulating in
the sediments but could be available during anoxic conditions.
-70-
-------
TABLE 17. NEUTRON ACTIVATION RESULTS FOR SUSPENDED SEDIMENTS
COLLECTED IN THE WISCONSIN RIVER AND COOLING POND (SEPTEMBER 1976)a
Wisconsin River stations
Element 107
As 2213
Ba 380±50
Ca 34,000±7,000
Ce 66±1
Co 11.6±0.4
Cs 2.110.3
Eu 0.8110.02
Fe 47,8001500
Hf 2.610.1
K 11,0001900
La 32.610.4
Lu 0.30610.014
Na 3,210140
Nd 43110
Rb 3716
Sb 3.110.2
Sc 7.810.2
Se 1.910.7
Sm 6.0410.06
Tb 0.8410.05
Th 4.7110.14
Yb 2.0110.13
Zn 340110
Suspended
sediment
(mg/liter) 13.9
108
2413
1,330180
19,00016,000
6112
13.010.5
2.010.4
1.0910.04
54,2001700
2.610.2
12,3001800
30.410.6
0.3010.02
3,730170
4618
2.9±0.3
8.6510.09
5.210.2
0.6010.05
7.310.2
1.710.2
32219
13.1
109
3013
500150
17,00015,000
6311
13.310.4
2.010.3
1.0010.03
52,8001500
2.910.2
13,40011,000
30.510.5
0.29110.014
5,4001100
3519
4416
3.510.6
8.4410.08
5.3010.05
0.5910.04
7.710.2
2.0710.12
509113
13.6
Cooling pond stations
105
1812
560120
17,00013,000
18.110.7
5.910.2
0.8610.14
0.30610.01
20,2001200
0.9910.05
7,7001500
10.610.3
0.08510.010
2,720130
2014
1.210.1
2.3710.03
1.5510.03
0.2210.03
3.3610.14
4.410.6
18816
12.3
106
1613
760130
6.910.3
24,4001300
1.0410.12
10,70011,100
11.710.6
0.12610.013
1,730150
2517
2615
1.510.2
3.0410.03
2.1210.04
0.2910.03
4.7110.14
0.7010.07
422111
10.6
-71-
-------
Table 18 presents a mass balance for several elements in the ash
basin. Calculations were based on averaged data from neutron activation
analysis of suspended sediments (collected at the ash basin intake and
discharge), plant pumping records (Table 1), and data from the monitoring
effort. Groundwater fluxes could not be calculated due to lack of accurate
flow data. Similarly, no provisions have been made regarding the
intermittent fly ash discharge. The data should, nevertheless, reflect
typical input/output levels at the ash basin before Columbia Unit II was
operational.
TABLE 18. MASS BALANCE FOR THE ASH BASIN
Input (kg/yr)
Element
Al
As
Ba
Ca
Co
Cu
Fe
K
Nab
Sb
SO,
Si
Zn
Cooling pond
2,380 ( 4)a
22 (69)
125 (22)
136,000 (99)
2 (57)
36 (99)
1,800 <28)
8,570 (95)
46,360 (98)
1 (97)
63,000 (99)
10,100 (12)
34 (44)
Fly ash
8,100,000
2,563
1,040,040
16,724,850
1,281
10,240
3,891,410
520,000
323,000
1,158
1,152,980
27,018,700
9,005
Total
8,102,380
2,585
1,040,165
16,860,850
1,283
10,276
3,893,210
528,570
369,360
1,159
1,215,980
27,028,800
9,039
Rarcent
Output retained
(kg/yr) in ash basin
59,170 (99)
315 (99)
1,015 (97)
98,541 (99)
—
49 (99)
95 (62)
9,107 (99)
1,400,000 (99)
—
492,520 (99)
10,450 (94)
12 (87)
>99
88
>99
>99
—
>99
>99
98
—
—
59
>99
>99
aNumbers in parenthesis represent average percentage "dissolved"
concentrations (that is, <0.4 m).
Na2COo is used in electrostatic precipitators.
(1)
The mass balance calculations around the ash basin indicate that:
Most of the elements reach the ash basin in particulate form and remain
there; those discharged in significant quantities are in a soluble form.
This is due to efficient settling of fly ash and precipitates and because
high pH values solubilize certain elements; (2) with the exception of SO/
and As (and possibly Cr, B, Se, and V), at least 99% of the measured
elements are retained within the ash basin; (3) a significant quantity of
is added to the system and passes through. The additional Na comes from
NaCOo which is used to improve electrostatic precipitator efficiency; and
Na
(4) although the addition of H^SO^ to the ash basin effluent is effective in
lowering the pH, it results in flocculation and settling of Al in the
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-------
discharge creek. Concurrent removal of other elements is also possible (by
adsorption and coprecipitation).
The elemental discharge data from the ash basin has also been combined
with concentration and flow data from the Wisconsin River in an effort to
discern whether their releases influence the water quality of the river. It
must again be kept in mind that these calculations are based on yearly
averaged data, that some elements will deposit in Rocky Run Creek (for
example, Al and Ba), and that no distinction has been made for soluble and
particulate phases. The results from these calculations should, however, be
representative to within a factor of two or three and are presented in
Table 19.
TABLE 19. POSSIBLE ELEMENTAL CONCENTRATION CHANGES IN
WISCONSIN RIVER DUE TO ASH BASIN DISCHARGE3
Element % concentration change
Al
As
B
Ba
Ca
Cr
Cu
Fe
K
Na
Si
So A
Zn
0.97
1.1
3-5
0.5
0.6
1-4
0.5
0.003
0.08
2.5
0.05
0.6
0.01
a 12
Calculations based on an annual Wisconsin River flow of 5.3x10
liters/year (Andrews and Anderson 1976).
These calculations indicate that the elemental discharge from the ash
basin has a negligible effect on the water quality of the Wisconsin River.
Although Na, B, and Cr may represent possible exceptions, it would not be
feasible to measure concentration changes on site, because the accuracy and
precision of analytical procedures are usually between 5 and 10%. Similar
conclusions would hold for other elements not presented in Table 18 (such as
Pb, Cd, V, Se, Sb, and U). Although several elements are elevated in the
ash basin effluent, the dilution ratio between the discharge creek and
Wisconsin River is about 1:270. Any significant biological effect, resulting
from ash basin discharge of chemicals, should thus confined to Rocky Run
Creek.
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THE EFFECT OF FLY ASH DISCHARGE AND STORAGE ON RECEIVING WATERS
This study has shown that the largest fraction of the trace elements
entrained in this system reside in the particulate form. However, when
these particles are discharged from a primarily inorganic system to
receiving waters of different origin, the trace element distribution may be
altered. A pH decrease might cause some fraction of the trace elements
(cations) to be released to the dissolved phase despite their strong
electrostatic attraction to the ash surface. In addition, the organic
matter present in receiving waters will tend to solubilize trace metals such
as such as Cd, Cu, Pb, and Zn by chelation and complexation reactions. The
extent of these organo-metallic interactions clearly depends upon the
composition of the receiving water. Natural waters, such as the Wisconsin
River, that contain large amounts of humic and fulvic acids should
experience these types of organo-metallic interactions.
Acidic rainfall on ash disposal sites could also mobilize trace
elements. Extensive data is not available describing rainwater leaching of
trace elements from fly ash. In a preliminary series of experiments,
conducted to estimate this leaching potential, Natusch (1975) observed that
5 to 30% of the toxic elements (for example, Cd, Cu, and Pb) are
leachable. This equilibrium leaching study indicates that at least 10% (as
per p. 91) of the cadmium would be solubilized based on acid rain pH values
of 3.0 to 5.0 reported for the northern United States by Galloway et al.
(1976), Beamish and Van Loon (1977), and Murphy (1974). Of course, if this
leaching occurred in the ash pond it would be suppressed by solution
equilibria as described previously. However, in other ash landfill areas the
metallic burden of leachate waters may be directly exposed to the
surrounding environment. Benninger et al. (1975), using Pb as a heavy
metal tracer, demonstrated the strong sequestering of dissolved trace metals
210
by organic rich soils. Less than 2% of the Pb supplied to the terrain
was transmitted to groundwaters. Similar results of efficient trace metal
scavenging by suspended particulate matter in the Colorado River was
observed earlier by Rama and Goldberg (1961). Using these data, Benninger
et al. (1975) calculated that only 0.02%/yr of the metals residing in the
Susquehanna River watershed are transported out of the system under normal
weathering and erosion conditions. This calculation implies that adsorption
processes and soil organo-metallic interactions should scavenge a large
fraction of the trace elements in fly ash leachate waters.
Some evidence indicates that groundwater can be contaminated by ash
basin water. Well samples adjacent to the Columbia ash basin may
sporadically exhibit higher dissolved concentrations of arsenic, boron and
sulfate than groundwaters farther away from the dike. Obviously, there is a
finite quantity which may be retained by soil underlying and surrounding an
ash disposal site. When this potential is exceeded, contamination of
groundwater and close-by streams, rivers, or lakes is possible.
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Fly ash particles, on the other hand, may act as a sink for trace
elements under certain environmental conditions. Adsorption of trace
cations may occur when the pH is above 7.55. These pH conditions are
typical for natural water systems (Stumm and Morgan 1970). The hydrous
oxide surface coating of aluminum and iron encompassing the ash particles
provide an excellent substrate for adsorption of trace metals. Such
surfaces have a large adsorption capacity for many species (Fox 1968, Shukla
et al. 1971). These interactions may be significant in the surface organic
microlayer (SOM) of freshwater lakes where the atmospheric input of
particulates, including fly ash, is important. After deposition, subsequent
trace element enrichment of these anthropogenic particulates will
undoubtedly occur. These particles may then return to the atmosphere by
bubble injection (Maclntyre 1974, Wallace and Duce 1975) or interact with
biological and chemical components of the lake (Elzerman 1976).
Other elements like calcium, potassium, magnesium, and sodium will be
readily released to the dissolved phase as fly ash particles enter natural
waters either by natural weathering processes or atmospheric deposition.
These elements are of no apparent toxicological concern, but a pH rise may
occur in natural waters having an insufficient buffering capacity to
neutralize their hydrolytic dissolution reactions. Any pH increase due to
these reactions should be a localized effect. The magnitude of such a pH
change is explicitly a function of the fly ash/water ratio and the buffering
capacity of the particular receiving water.
A more important effect with regard to the overall water quality may be
the lowering of the phosphorus content by co-precipitation with calcium
and/or magnesium phases. This precipitation phenomena would decrease the
fertility of a water body. Furthermore, this process may be enhanced in
calcareus lakes commonly encountered in the north central U.S. Previous
studies performed by Tenney and Echelberger (1970) and Higgins et al. (1976)
indicate the vehicle for inorganic phosphorus removal from lake waters after
fly ash addition is precipitation of amorphous calcium-phosphates. Both of
these studies, however, employed large solid to solution ratios (e.g.
10 g/liter). Dilution of ash pond waters by receiving waters may effectively
decrease solid-solution ratios of ash pond discharges so this phenomena
becomes unimportant during normal flow periods. In fact, only a few
mg/liter suspended solid load exists downstream of the ash pond drain at the
Columbia site (Andren et al. 1976).
The input of phosphorus and silica to a water body by fly ash
dissolution could be important in terms of influencing the biological make-
up of that system. Fly ash particles may contribute large amounts of silica
and perhaps in some cases significant quantities of phosphorus. At pH
values of natural water systems (7.0 to 9.0) phosphorus should be released
from the ash particle to the dissolved phase. Since the phosphorus content
of the Columbia fly ash is less than 1%, it is probably too small a quantity
to be of importance for most freshwater lakes. It is very likely that the
particles containing the phosphorus may sink to the sediments of rivers or
shallow lakes before much of the phosphorus dissolves. It also appears that
prior dissolution of surface constituents, i.e., sulfate coating, is
necessary before phosphorus becomes directly available to the bulk
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-------
solution. Several days in an aquatic environment may be required for this
coating to dissolve (Talbot 1977, tfelmke et al. 1977).
Silica should dissolve from the fly ash over a wide pH range, as was
demonstrated by its pH dependence. Diatoms may utilize this silica as it
dissolves. The remaining silica will probably be deposited in the
sediments, where it may then be incorporated into the silica cycle of the
lake (Vigon 1976). Although silica comprises about 23% of the Columbia fly
ash, introduction of a large quantity of fly ash would most likely be needed
before this became a major silica supply for diatoms. Lastly, a desirable
lake management goal of establishing a viable diatom population to control
objectional algae blooms in eutrophic lake waters might include lake
treatment with fly ash. Further research is needed, however, to ensure that
toxic trace elements, (including transistion metals and anions like arsenic
or boron) are adequately removed by pretreatment or are in low enough
concentrations to constitute its use in such lake restoration projects.
-76-
-------
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TECHNICAL REPORT DATA
(Please read Instructions on the reverse before completing)
1. REPORT NO.
EPA-600/3-80-076
2.
3. RECIPIENT'S ACCESSION NO.
4. TITLE AND SUBTITLE
ELEMENT FLOW IN AQUATIC SYSTEMS SURROUNDING COAL-FIRED
POWER PLANTS
Wisconsin Power Plant Impact Study
5. REPORT DATE
JULY 1980 ISSUING DATE.
6. PERFORMING ORGANIZATION CODE
7. AUTHOR(S)
Anders Andren, Marc Anderson, Nicholas Loux, Robert
Talbot
8. PERFORMING ORGANIZATION REPORT NO.
3. PERFORMING ORGANIZATION NAME AND ADDRESS
Department of Water Chemistry
University of Wisconsin
Madison, WI 53706
10. PROGRAM ELEMENT NO.
IBA820
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, MN 5580>l
13. TYPE OF REPORT AND PERIOD COVERED
14. SPONSORING AGENCY CODE
EPA/600/03
15. SUPPLEMENTARY NOTES
16. ABSTRACT
Water quality parameters of a 192-ha (480-acre) cooling pond adjacent to the
Columbia Generating Station, Portage, Wisconsin, have been investigated. Analyses
were made for major and minor elements, nutrients, pH, alkalinity, 02, chloroogranics,
phenols, and polyaromatic hydrocarbons. Similar parameters were also measured in the
nearby fly ash discharge basin and its associated drainage stream. Laboratory
dissolution and precipitation studies of fly ash were performed in an effort to under-
stand the chemistry of the discharged ash water and its potential effects on receiving
waters. Mass balance acalculations were made and are presented to ascertain whether
the cooling pond acts as an efficient sink for inorganic and organic compounds, and
if so, what the fate of these compounds is. Data presented in this report are also
discussed in terms of plant operating characteristiqs. Remedial procedures are pre-
sented which could alleviate present and anticipated problems.
17.
KEY WORDS AND DOCUMENT ANALYSIS
DESCRIPTORS
b.lDENTIFIERS/OPEN ENDED TERMS C. COSATI Field/Group
Fly ash leachate
Elemental analyses
Organic analyses
Ecosystems effects
Land use
Wisconsin power plant
;. study
Cooling ponds
Waste sinks
07/B
07/C
18. DISTRIBUTION STATEMENT
Release to public
19. SECURITY CLASS (This Report)
unclassified
21. NO. OF PAGES
94
20. SECURITY CLASS (This page I
unclassified
22. PRICE
EPA Form 2220-1 (Rev. 4-77)
PREVIOUS EDITION IS OBSOLETE-,
-O4-
U.S. GOVERNMENT PRINTING OFFICE: 1980--657-165/0092
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