PB95-191243
EPA No. 530-R-95-011
Final Report
Water Quality in Open Pit Precious Metal Mines
Margaret Saunders Macdonald
Glenn C. Miller
W. Berry Lyons
Department of Environmental and Resource Sciences
University of Nevada, Reno
Reno, NV 89557
December 1994
Supported by a Grant from the U.S. Environmental Protection Agency
(Grant #R820245)
v-! Pmtpction Agency
U.S. Environmental Proteouui &
Chicago/ IL° 60604-3590 Re<*cl«m,cyc,ab,e
7~^ <*\ Printed with Soy/Cancta Ink on paper that
contains at least 50% recycled fiber
-------
Abstract
\,v
There are many open pit precious metal mines in Nevada that will be filled with water
after mining is completed. This study was conducted to document the current level of
understanding of issues concerning pit water quality, and to determine where additional
research is needed. A computer-aided literature search was conducted using the keywords pit
water, open pit, groundwater, mining, arsenic, geochemistry, wallrock, evapoconcentration,
hydrothermal, acid mine drainage and various combinations of these words. There is a wealth
of information available about lakes. However, pit lakes are a new type of lake, with unique
geometry and geochemistry giving them characteristics unlike those of natural lakes. Factors
that contribute to pit water quality include flow of groundwater, water-wallrock reactions, pH,
trace element concentrations (especially arsenic), evapoconcentration and hydrothermal
activity. Pit lakes exist from the mining of minerals such as phosphate, uranium, coal,
copper, silver and gold. None of these lakes have been studied in extreme detail, but limited
studies contribute knowledge to the understanding of water quality in open pit mines. The
variability in the contributions of the factors that affect pit water quality and the unique
geology in each mine necessitate the study of each pit lake or future pit lake on an individual
basis, and make prediction of water quality difficult.
v Disclaimer
o^
This document was prepared by the Department of Environmental and Resource
Sciences at the University of Nevada, Reno, under a grant from the U.S. Environmental
Protection Agency. Although the Environmental Protection Agency supported the
development of this document and is making this report available to the public, the
information presented here is solely the product of the authors and does not necessarily reflect
the Agency's position or policy.
In the electronic version of this document, the figures and tables are included as a
separate file.
11
-------
Table of Contents
* Page
Abstract ii
Disclaimer : : ' ii
Table of Contents iii
List of Figures iv
List of Tables v
Introduction 1
Factors Contributing to Pit Water Quality 8
Groundwater flow 8
Water-wallrock reactions 10
Row in fractures 11
Sloughing of wallrocks 11
Acid versus Alkaline Pit Water 12
The acid forming process . 12
Neutralization 14
Evaluation of potential for acid production 17
Mitigation of acid mine drainage 19
Trace Elements 20
Arsenic Speciation 24
Evapoconcentration , 30
Hydrothermal Activity 31
Existing Pit Lakes 34
Phosphate Mines 34
Uranium Mines 35
Coal Mines 36
Copper Mines 38
Precious Metals: Silver Mines 50
Precious Metals: Gold Mines 52
ill
-------
Effects of Pit Water Quality on Life 55
Reclamation 59
Conclusion : ' 61
References 64
List of Figures
re Page
1 Nevada open pit mines that are water-filled, are dewatering now, or
will be dewatering in the future 2
2 Typical thermal stratification of a lake * 4
3 Seasonal changes in stratification 5
4 Groundwater movement after mining 9
5 Eh-pH diagram showing contours of dissolved iron 22
6 Arsenic removal from 300 mg/1 As solution with six different levels of
Fe/As 25
7 Solubility of scorodite, FeAsO4*2H2O, at 23°C 27
8 Eh-pH diagram for the system Fe-As-H2O showing arsenic compounds
at25°C 28
9 pH, Berkeley Pit, Butte, Montana 40
10 Dissolved Oxygen, Berkeley Pit, Butte, Montana 41
11 Fe(H) and Fe(ffl), Berkeley Pit, Butte, Montana.... 42
12 Dissolved As, Berkeley Pit, Butte, Montana 43
13 Field Eh, Berkeley Pit, Butte, Montana 44
14 Mass of solids on filter, Berkeley Pit, Butte, Montana 45
IV
-------
15 1985-89 aqueous pH, Southern Tail Pit, Equity Silver Mines, B.C 51
t
V
16 Copper, iron and zinc versus pH, Southern Tail Pit, Equity Silver Mines,
B.C 53
17 Geology of Nevada's current and future pit lakes 63
List of Tables
Table ' Page
1 Iron sulfate minerals identified at Iron Mountain, CA 15
2 Pit water quality at Ruth 47
3 Yerington pit water quality 48
4 U.S. EPA drinking water standards 49
5 Cortez Gold Mine pit water quality 54
-------
Introduction
*
:
The development of technology that allows heap leaching of disseminated gold ores by
cyanide solutions led to the profitable mining of gold ores with average grade as low as .02
ounces per ton (NBMG, 1989). This means that, on average, 50 tons of rock are removed to
produce each ounce of gold. With over 5 million troy ounces of gold produced in Nevada
each year (NBMG, 1991), a large amount of rock is removed from open pit mines. Many of
the mines are now dewatering to retrieve ore below the water table. As mining goes deeper
sulfide ore bodies are encountered rather than the oxide ore bodies found closer to the
surface. When dewatering stops and the water table in the pits returns to its original level,
the mines will contain pit lakes. Figure 1 shows open pit mines in Nevada that are currently
water-filled, are now dewatering, or will be dewatering in the future. These lakes will exist
long after mining ends, causing environmental concern for centuries.
Because open pit precious metal mining below the water table is a relatively new
phenomenon, there is a lack of information regarding water quality in these pits. The only
discussion of pit water geochemistry is that of Davis and Ashenberg (1987) who studied the
Berkeley Pit lake in Butte, Montana. The objectives of this research paper were to document
the current level of understanding of issues concerning pit water quality, and to determine
where additional research is needed.
A literature search was conducted using Silver Platter software and the Georef and
Selected Water Resources Abstracts databases. The keywords used for the search were pit
water, open pit, groundwater, mining, arsenic, geochemistry, wallrock, evapoconcentration,
hydrothermal, acid mine drainage and various combinations of these words. All information
for each citation was searched (title, abstract, keywords, etc.), and all years in the databases
were included.
There is a wealth of information available about lakes (Hutchinson, 1957; Wetzel,
1983). All of the natural lake types have been studied and classified according to
-------
Twin Crwta * tori*a
GoWsthke
100 km
SCALE
Figure 1 Location map for Nevada open pit nines requiring dewatering
-------
circulation and stratification patterns. Biologists know in detail the types of aquatic biota that
are found in 'natural lakes. Man-made reservoirs have been studied and found to have
fundamentally different characteristics from natural lakes, but those characteristics are easy to
understand since.the dynamics of natural lakes are known and both natural lakes and man-
made reservoirs are usually relatively shallow (less than 20 m deep) (Wetzel, 1983). Mining
pit lakes, however, are a new type of man-made water body due to their extreme depth (often
over 300 m deep). Their surface area to depth ratio is small, making them most similar in
shape to natural volcanic crater lakes such as Crater Lake in Oregon. Newly exposed
wallrock in pit lakes, with the mineralogy associated with ore deposits, can contribute
chemical constituents in higher concentrations than are found in most natural volcanic lakes.
What follows is a review of what is known about the behavior of natural lakes
(Wetzel, 1983), which will give an idea of the possible behavior of pit lakes. However,
definitive answers will not be available until the pits are filled with water and limnological
studies are completed on them.
Thermal stratification occurs in a lake when the surface waters are heated by solar
radiation more rapidly than the heat is dissipated by mixing. The temperature profile of a
typical stratified lake has three zones (Figure 2): warm water in the epilimnion, cold water in
the hypolimnion, and a zone of rapid temperature change with depth called the thermocline.
The maximum density of water occurs at 4°C so this is a very stable profile, with the densest
water at the bottom.
Lakes can undergo seasonal changes in stratification (Figure 3). The most stable
stratification profile, with maximum temperature difference between the epilimnion and the
hypolimnion, occurs in summer. Oxygen in the hypolimnion varies seasonally and can go to
zero under stratified conditions as shown in Figure 3, depending on whether oxygen-
consuming materials are present. An oxygen deficit in the hypolimnion is usually due to
oxidation of organic matter but can also be caused by chemical oxidation (Hutchinson, 1957).
As the surface waters cool in the fall, the temperature profile becomes less and less stable.
-------
*-. 15
25
Figure 2
EPILIMNION
10
THERMOCLINE
20
HYPOLIMNION
30
10 15 20
TEMPERATURE (°C)
25 30
Typical thermal stratification of a lake.
After Wetzel, 1983
Limnology
-------
*
SUMMER STRATIFICATION
.
i
i
>
0
:
X
»
r
I
1
mg/l
48 12
I
f
"** J
\
;
i
>
10 QgZO 30
WINTER STRATIFICATION
j
4
0
*
4*
i
f
\
.
0
_.-
/
mg/l
4 8
fee
,*
.""
g«»**
10 ^ 20
12
;
:
i
/
.*'
j
.
FALL TURNOVER
0 4 *" 8 12
|
f
0
5
10 0^20 30
- ^
«2
-^~~" Temp.
SPRING TURNOVER After Het«u 1983
0 4 8 12
|
i
1 to
>
30
f
j
i
*
5
i
j
i
I
i
;
i
i
i
O 1O _ 20 30
Limnology
Figure 3
-------
Turnover may occur, generally on a windy day when the water in the epilimnion is nearly the
same temperature as the water in the hypolimnion. The entire lake will mix and oxygen
content will be high throughout (Hutchinson, 1957). In winter, ice may form as the surface
water cools, and the lake is once again stratified. Then, as the surface warms in the spring,
the ice melts and the profile again becomes unstable due to nearly isothermal conditions. If
the lake receives sufficient energy from wind, again turnover and mixing occur, and again
oxygen contents are high throughout the profile.
The water in the Berkeley Pit in Butte, Montana did not freeze in the winters of 1988-
89 or 1989-90, but did freeze in earlier years (Huang and Tahija, 1990). In the years that the
Berkeley Pit froze, immediately after the ice melted in the spring the surface water of the pit
was a blue-green color. By midsummer, it was brown in the upper 4 m and blue-green below
4 m (Davis and Ashenberg, 1989). The blue-green color represents reducing conditions (a
Cu(II) mineral). As the surface waters became oxygenated after the ice cover was gone, iron
oxyhydroxide precipitated giving the brown color. Ice may play a role in lowering surface
metal concentrations due to exclusion of soluble ions as ice forms. The unfrozen solution
below the ice is then enriched in soluble ions and may sink due to higher density (Huang and
Tahija, 1990).
A high concentration of total dissolved solids at the bottom of a lake may cause a
dense layer of water that does not participate in mixing. This dense layer is called the
monimolimnion and the mixing layers above are called the mixolimnion. The steep
salinity/density gradient between the layers is called the chemocline. Lakes with this density
profile are called meromictic. Cryogenic meromixis occurs from "freezing out" of salts from
ice, which then precipitate and sink to deep water, as mentioned above. Crenogenic
meromixis is caused by submerged saline springs, and biogenic meromixis occurs from salts
released during the decomposition of organic matter. In a meromictic lake the mixolimnion
goes through the same circulation and stratification patterns as non-meromictic lakes.
There are many types of stratification patterns that occur in lakes which depend on
climate, morphometry and chemistry. Lakes which undergo complete circulation in the spring
and fall and are thermally stratified in the summer and winter are called dimictic lakes.
-------
Lakes in which temperatures never go below 4°C circulate at or above 4°C all winter and
stratify in the summer. These are the most common patterns among temperate lakes.
Whether or not pit lakes will mix is difficult to predict. Density differences due to
concentration of total dissolved solids, due to input from hydrothermal springs, due to
climate, and due to pit lake morphometry all contribute to the stability or lack of stability of a
lake's density profile. More studies of the limnology of existing pit lakes, are needed.
Measurements should be taken at different times of the year to provide information about
seasonal changes.
-------
Factors Contributing to Pit Water Quality
t
\
Many factors will contribute to the quality of water in the pits. The pit water may be
acidic or alkaline, arsenic concentrations may be high or low, other trace metals may be .high
or low and evapoconcentration may affect their concentrations. Groundwater flow around the
pit is important and hydrothermal water may or may not enter the system. Open pit mining
significantly increases the rock surface area that is exposed to water and water-rock reactions
are a primary consideration in prediction of water quality. In this section each of these
factors will be examined.
Groundwater flow
Groundwater flow is an important consideration in determining pit water quality.
Conditions prior to mining must be well known before any predictions can be made about pit
water. The original water table elevation will generally be the same as the ultimate water
level in the pit and thus determines, in part, the depth of the pit lake.
Figure 4 shows groundwater flow around an open pit after mining. Groundwater flow
would be expected to flow into and out of the pit through aquifers, but not everywhere as
implied in the generalized Figure 4. It is important to know depth and flow rates of all
aquifers intersected by the pit. For an open pit mine on a flat land surface, the pit will fill
and stabilize near the level of the original water table except for evaporation and additional
recharge. Alternatively, on a sloping land surface, the pit may overflow if the original water
table elevation is sufficiently high.
Schubert (1980) found that groundwater flow changed direction in coal mine spoil
material adjacent to a lake in Illinois, depending on storm precipitation. During high lake
levels after a storm the spoil material received recharge from the lake. The lake then receded
due to surface discharge and evaporation, and groundwater seeped back into the lake.
Groundwater seeps from the White's and Intermediate open cuts at Rum Jungle and
-------
FLAT LAND SURFACE
WATER TABLE
SLOPING LAND SURFACE
SROUNDWATER FLOW
CASE B
MATER HOVENENT AFTER MINING.
After Morln, 1990
Acid Drainage frcn Mine Halls
The Main Zone Pit at Equity
Silver Mines
Figure 4
-------
contributes to the pollution load in an adjacent river (Harries and Ritchie, 1988). At the
Nabarlek Uranium Mine, groundwater seeps from the pit when the water level in tailings in
the pit is higher than the surrounding groundwater level during the rainy season (McLoughlin
et al., 1988).
Dewatering may alter the natural groundwater flow regime. Cones of depression may
extend 10 km or more and subsidence may occur within the cone. Up to .5 m subsidence
was recorded within the cone of depression of a large (approx. 2.5 km2) dewatered open pit
coal mine in Poland with maximum drawdown in the pit of 200 m (Wojciechowski and
Serewko, 1985).
Direction of groundwater flow can be an important factor in determining the quality of
pit water. Wicks et al. (1991) described final-cut coal strip mine lakes which had spoil
material on one side, and bedrock on the other side. In these cases, sulfide oxidation
occurred only on the spoil material side, not on the bedrock side. In this situation, upgradient
lithology determines groundwater quality. Groundwater flowing from the bedrock into the
lake would be relatively uncontaminated. If groundwater flowed from the spoil into the lake,
lower water quality would be expected.
Water-wallrock reactions
In addition to upgradient lithology, much of the quality of pit water will depend on the
interaction between water and the wallrocks in the pit. This interaction will depend on the
amount of wallrock exposed to water, how permeable the wallrocks are, how stable the
submerged pit slopes are, as well as the redox and chemical conditions in the lake. This
section discusses the physical processes that determine the amount of wallrock area exposed
to water. The redox and chemical conditions are discussed in later sections.
The area of wallrock surface that will be exposed to water should be computable from
the mine plans and knowledge of the original water table elevation. Based on these
calculations, one can determine the fixed surface area of wallrock available for reactions.
This calculation would, however, be complicated by flow in fractures and joints and by
10
-------
sloughing of pit walls both above and below the pit water surface, which will provide new
mineral surfaces for reaction.
Flow in fractures
"Water flow in joints and fractures can greatly increase the rock surface area available
for water-rock reactions. The permeability of a system of fractures is much larger than that
of a porous medium (Bear, 1979), but the exact amount of permeability is often difficult to
determine. In addition to joints and fractures that are naturally present, some will be
introduced by blasting during active mining. A fault may have low permeability zones
associated with the fault gouge as well as zones of high permeability in the fractured rock.
Fault breccia may be more permeable than the gouge. For these reasons, faults can act as
groundwater barriers, as groundwater conduits, or both at the same time (Patton and Deere,
1971).
Sloughing of wallrocks
Most of the research on pit slope stability found in the literature pertains to stability in
dewatered pits during active mining. The factors affecting pit slope stability are rock
structure, including bedding planes, joints and faults; groundwater pressure; earthquakes;
weight of a block of rock; hydrostatic forces in joints and fractures; and damage from blasting
(Broadbent and Zavodni, 1982; Hoek, 1971; Brown, 1982; Glass, 1982; Piteau and Martin,
1982 and Holmberg and Maki, 1982). Wide fault zones may have large areas of weathered
or hydrothermally altered rock and may influence groundwater flow such that pore pressures
are excessive within and adjacent to the faults (Patton and Deere, 1971).
An additional factor in determining slope stability is the regional groundwater flow
system. Determination of whether the pit is located in a groundwater recharge area or
discharge area is important because groundwater discharge areas have a greater likelihood of
having excess pore water pressure in the walls and beneath the floor of the mine (Patton and
Deere, 1971).
11
-------
The slope stability factors discussed above apply to wall rock above the water surface
in water-filled open pits. No studies were found on the stability of submerged mine walls,
although Davis and Ashenberg (1989) indicated that the submerged walls in the Berkeley Pit
at Butte, Montana are unstable. Sloughing has filled in 38 m of the pit lake. The Ruth
(Nevada) Pit bottom is currently 37 m higher than the original pit bottom (Woodward-Clyde
Consultants, 1992) presumably due to sloughing of wall rocks. Additional studies are needed
to determine how much sloughing is actually occurring in existing pit water lakes, and how
this sloughing affects the water-rock reactions that will take place in the lakes.
Acid versus Alkaline Pit Water
Acid mine drainage (AMD) is considered to be the greatest chemical problem caused
by mining (U.C. Berkeley Mining Waste Study Team, 1988). Generation of acids during and
after mining results from the oxidation of naturally occurring minerals. Once the AMD
process begins, it is difficult to control and often accelerates and is likely to persist for
decades or centuries. Several examples exist where AMD began after closure of a mine.
After closure, it is difficult to adopt mitigating measures. Acidification of pit water is not
expected to be generally observed in precious metal mines. However, the problem may occur
in some mines and is of sufficient magnitude that a discussion on the process is given below.
The acid forming process
i
Geology of the wallrocks in the pit and adjacent rocks with which surface water
comes into contact are the main influences on the pH of pit water. The amount of acid
generated will be determined by the amount of sulfide minerals and/or ferrous iron in the
wallrocks and adjacent rocks that are available for oxidation. Oxidation of sulfides forms
acid by the following reaction sequence (Steffen Robertson and Kirsten, 1989):
FeS2 + 7/2 O2 + H2O -> Fe2+ + 2SO42 + 2KT
If conditions are sufficiently oxidizing, Fe(II) will oxidize to Fe(HI):
Fe2+ + 1/4 O2 + H+ --> Fe^ + 1/2 H2O
r
At pH values above 2.3 to 3.5, Fe(ffl) will precipitate as Fe(OH)3:
Fe3* + 3H20 --> Fe(OH)3 + 3H+
12
-------
In this reaction sequence, acidification results from the overall production of four equivalents
of hydrogen aons. At low pH values (<4), and in the presence or absence of oxygen, pyrite
is rapidly oxidized by ferric iron (Nordstrom, 1982; and Nordstrom and Alpers, 1990):
FeS2 + 14Fe^ + 8H20 --> 15Fe2+ + 2SO42' + 16H+
Pyrite oxidation by microbial activity can continue under low or undetectable oxygen
concentrations because ferric iron can be used instead of oxygen as an electron acceptor for
the bacteria (Nordstrom, 1982). The primary chemical factors which determine the rate of
acid generation are pH, temperature (the oxidation rate by oxygen doubles for every 10°C
rise, Nordstrom, 1982), oxygen content of the gas phase if saturation is less than 100%,
oxygen concentration in the water phase, degree to which a rock is saturated with water,
chemical activity of Fe3*, surface area of exposed metal sulfide, and chemical activation
energy required to initiate acid generation (Steffen Robertson and Kirsten, 1989).
Bacteria such as Thiobacillus ferrooxidans accelerate the rate of ferrous-iron oxidation
and of reduced-sulfur oxidation at pH less than 4, and thus accelerate the rate of acid
production. As long as pH is maintained above about 4 the rate of acid formation is slow
(U.C. Berkeley Mining Waste Study Team, 1988). But below pH of 4, bacteria greatly
accelerate the oxidation rate, and once the acceleration occurs it is difficult to reverse.
Bacterial growth of T. ferrooxidans is greatest in waters with pH of 3.2, and at temperatures
less than 40°C (U.C. Berkeley Mining Waste Study Team, 1988). Above 40°C bacterial
oxidation declines rapidly and above 45°C there is little or no bacterial action. Additional
factors which determine the bacterial oxidation rate include the biological activation energy,
population density of bacteria, rate of population growth, nitrate concentration, carbon dioxide
content and concentrations of any bacterial inhibitors. While T\ ferrooxidans is the most
common sulfur- and iron-oxidizing bacteria there are at least 18 other species of bacteria
which can oxidize or reduce iron or sulfur (Steffen Robertson and Kirsten, 1989). In addition
to pyrite (iron sulfide), bacteria can accelerate the oxidation rate of sulfides of antimony,
gallium, molybdenum, arsenic, copper, cadmium, cobalt, nickel, lead and zinc (Lundgren and
Silver, 1980).
Efflorescent iron sulfate minerals (Table 1) can form when sulfides are oxidized but
the acid produced is not flushed away and evaporation occurs (Nordstrom and Alpers, 1990).
13
-------
These minerals are highly soluble and "store" acid. Rhomboclase, for example, has acid in its
formula: (H$O)Fe(SO4)2*3H2O. These minerals might form in the unsaturated zone, and the
first flush of water that comes through would dissolve them, producing a pulse of acid. In
this way hundreds of kilograms of acidity could be stored within the walls of a mine (Morth
et al., 1972). This stored acidity could be flushed out by normal groundwater recharge during
mining or could be washed out during final pit flooding causing mine water to become
acidic. This acid water could be mistaken for active acid generation, which may or may not
be occurring. Therefore, it is important to determine the rate of acid generation and the
movement or lack of movement of water over the oxidation sites (Morin, 1990).
Neutralization
The acid produced by sulfide oxidation may be neutralized if the wallrocks contain
sufficient amounts of carbonate minerals:
CaCOj + IT --> Ca2++ HCCV
and
CaCO3 + 2IT -> Ca2+ + H2CO3
In order for CaCO3 to neutralize the acid water generated, it must exceed the equivalent
concentration of acid -and must also be exposed and available. If amorphous ferric hydroxide
precipitates in the area surrounding the reacting carbonate grains, the pH may again decrease
due to armoring of the calcite by the amorphous ferric hydroxide (Davis and Runnels, 1987).
14
-------
Table 1. Iron Sulfate Minerals Identified at Iron Mountain, CA
Mineral
Formula
Melanterite
Ferricopiapite
Copiapite
Roemerite
Kornelite
Coquimbite
Rhoraboclase
Voltaite
Additional Sulfates Identified
Gypsum
Chalcanthite
Halotrichite
' FenS04*7H20
Fe4m(SO4)5(OH)2*16H2O
FenFe4ra(SO4)6(OH)2*20H2O
FenFe2m(SO4)4*14H20
Fe2m(SO4)3*7H20
Fe2m(S04)3*9H20
(H3O)Fem(SO4)2*3H2O
K2Fe5nFe4m(SO4)12* 18H2O
CaS04*H2O
CuSO4*5H2O
FenAl2(SO4)4*22H2O
Nordstrom and Alpers, written
communication, 1990
15
-------
Other minerals consume acid including siderite (FeCO3), magnesite (MgCO3),
rhodochrosite (MnCO3), witherite (BaCO3), ankerite (CaFe(CO3)2), dolomite (MgCa(CO3)2),
malachite (Cu2CO3(OH)2), gibbsite (Al(OH)3),limonite/goethite (FeOOH), manganite
(MnOOH), and brucite (Mg(OH)2). Each mineral buffers to a different range of pH. Some
examples:
Mineral Buffer pH
Calcite and Aragonite 5.5-6.9
Siderite 5.1-6.0
Malachite 5.1-6.0
Gibbsite 4.3-3.7
Limonite/Goethite 3.0-3.7
(Steffen Robertson and Kirsten, 1989)
Not all of these minerals (e.g. gibbsite and limonite/goethite) can effectively bring the pH up
to the neutral range. Silicate minerals also consume acid, but not as rapidly as the carbonate
minerals. Inflow of groundwater can also supply acid-neutralizing alkalinity (Morin, 1990).
If conditions are reducing, acidity from acid mine drainage can also be neutralized by
microbially mediated sulfate reduction followed by reduced sulfur mineral precipitation
because sulfate reduction reverses the acidification process and generates alkalinity. This
process also lowers the concentration of iron and sulfate by precipitation of iron sulfides.
The equation follows:
SO42' + 2Cotsmk + 2H20 --> H2S + 2HC03-
The sulfide formed during this reduction process must be stored in a reduced form in the
sediments in order for the alkalinity generated to be permanent (Wicks et al., 1991).
Miller and Murray (1988a) outline the stages in the generation of acid mine drainage:
Stage 1 - chemical and/or biological oxidation of sulfide minerals which slowly produces
acid. The acid may be neutralized by carbonates and result in only a slight decrease in pH,
although total dissolved solids may increase. Testing programs that monitor only pH will not
identify sulfide oxidation during stage 1. Stage 2 - after the carbonates and other neutralizing
materials are consumed, the pH drops and acidophilic bacteria multiply. Stage 3 - when the
16
-------
pH drops below 3.5, bacterially catalyzed sulfide oxidation becomes effective and the rate of
,
acid generation is rapid.
Evaluation of potential for acid production
The initial characteristics of mine water do not necessarily reflect the long term
potential for acid mine drainage, and predicting the acid-forming ability of a rock type is
problematic. Also, the time for development of acid conditions can vary from less than one
day to more than 50 years (Miller and Murray, 1988a), making it difficult to monitor.
However acid producing potential must be evaluated to determine long term water quality.
The evaluation is done by static and kinetic tests.
After representative samples of all geologic units in a mine are obtained, static tests
are used to define a sample's balance between potentially acid-generating minerals (potential
acidity) and acid-neutralizing minerals (neutralization potential). Theoretically, a sample will
generate acid only if the potential acidity exceeds the neutralization potential, but as different
phases oxidize and/or dissolve the acidity/neutralization balance changes.
The general procedure for conducting static tests is first to measure total sulfur in a
sample which gives a determination of potential acidity, and then to determine the
neutralization potential. The third step is to calculate the net acid generating potential by
subtracting the potential acidity from the neutralization potential. Negative values indicate the
potential for net acidity.
There are problems associated with each step of this procedure because the mineralogy
and mineral associations in a sample are not considered. The measure of total sulfur may
overestimate the acid generating potential if some of the sulfur occurs in non-acid generating
minerals, or if some of the sulfur is not exposed and available for reaction in the sample
(Miller and Murray, 1988b; Doyle, 1991). The co-existence of pyrite with other sulfides such
as chalcopyrite and sphalerite decreases the oxidation rate (Nordstrom, 1982). The net
neutralization potential is usually obtained by crushing and blending a sample which is not
representative of conditions in the field. In an uncrushed rock, acid may be produced in a
17
-------
fracture, and may leave the rock via the fracture and never come into contact with
«
neutralizing minerals (Doyle, 1991). The determination of the potential for net acidity is not
totally reliable because some mine wastes that contain excess basic material still produced
acidic drainage {diPretoro and Rauch, 1988; Erickson and Hedin, 1988).
Despite the limitations of static tests, if they indicate that a sample is potentially acid
generating then kinetic tests should be conducted. Lawrence (1990) proposed that if the acid
neutralization potential does not exceed the acid generating potential by at least 3:1 there is
not a clear margin of safety and kinetic tests should be conducted. Kinetic tests involve
weathering samples under laboratory-controlled or on-site conditions to determine rates of
acid generation, sulfide oxidation, neutralization and metal depletion.
Steffen Robertson and Kirsten (1989) outline kinetic test procedures as follows. The
first step is to determine surface area, mineralogy and total sulfides. Particle size and
differences in surface area can account for differences in acid generation rates. The
mineralogy of a sample, specifically chemistry and crystal form of the minerals, is important
in controlling the rate of acid generation and neutralization (Hammack, 1985; Chander and
Briceno, 1988; Hammack et al., 1988). Total metal analyses are important for determining
when a metal in a sample may be depleted by leaching.
Problems are also associated with kinetic tests because laboratory tests can not
accurately duplicate field conditions and acid generation potential can be determined only by
field tests conducted for long time periods. Acid generating rock with significant amounts of
both sulfide and carbonate may yield drainage that is neutral for a long period of time before
becoming acidic (Steffen Robertson and Kirsten, 1989). Kinetic tests that simulate
weathering require at least 6 to 9 weeks to acquire meaningful data, and optimized
weathering, using a dilute acid for leaching instead of water, does not show what type of
discharge would result from normal weathering conditions (Hammack, 1985). Field kinetic
tests up to 20 weeks in length have been recommended in order to exceed the depletion of
neutralization potential and not underestimate the acid generation potential (Lapakko, 1990;
Lawrence, 1990).
18
-------
Comparison of pre-mine predictions of acidity or alkalinity from the literature with
actual results from the mines at a later time showed that accurate predictions do not
necessarily follow from testing programs (Ferguson and Erickson, 1988). A clear and correct
prediction can be made only if potentially acid-producing sulfide minerals are greatly
abundant or generally lacking relative to acid-consuming carbonate minerals.
Mitigation of acid mine drainage
There is currently no widely accepted technology which controls AMD production at
the source without indefinite maintenance (Ziemkiewicz et al., 1990). Steffen Robertson and
Kirsten (1989) discuss the three stages of control of acid rock drainage which are control of
the acid generation process, control of acid drainage migration, and treatment of acid rock
drainage. These methods all may require long term monitoring and maintenance.
Acid generation control involves excluding oxygen or water to inhibit sulfide
oxidation, or slowing the acid generation rate by controlling pH or controlling bacterial
activity. Oxygen and water can be excluded by use of impermeable covers. pH can be
controlled by adding neutralizing materials. Bacterial activity can be temporarily controlled
by addition of anionic surfactants such as sodium lauryl sulfate. Inhibition of Thiobacillus
ferrooxidans will not reduce acid formation unless the inhibitor is added to groundwater or
rainwater infiltration before contact with the pyrite occurs (Kleinmann et al., 1981). Once the
acid has left the site of pyrite oxidation, treatment with an inhibitor will not help.
Control of acid rock drainage migration involves diversion of surface water flowing
toward the acid source and prevention of groundwater flow into the acid source. These
control measures would be difficult, if not impossible, to achieve in open pit mines, although
drainage adits are recommended which drain away from the pit as a possible means of
long-term isolation of water from sulfide bearing rocks (Steffen Robertson and Kirsten,
1989). In some locations this may be a means of control in wallrocks above the surface of
the water in the pit. Surface water can be diverted away from the pit, but precipitation falling
at the pit edge will still run down the walls.
19
-------
Methods for treatment of acid mine drainage may be either active systems requiring
continuous operation such as chemical treatment plants or passive systems which operate
without human intervention such as wetlands. Steffen Robertson and Kirsten (1989) describe
in detail many types of treatment systems, some of which could be used for the pump and
treat method of pit water quality remediation.
In summary, in analyzing an open pit mine site for potential acid water, it is important
to know the geology of the wallrocks in detail. Until more work on testing methods is done,
the advantages and disadvantages of the various static and kinetic tests should be weighed in
choosing the test method that will best predict acid producing potential. If acid production is
expected, the various ways to prevent it or neutralize it can be analyzed and determination
can be made of whether or not it can be controlled before it becomes a problem.
Trace Elements
Trace elements are defined as elements that generally occur in waters at concentrations
of less than 1 mg/1 (Drever, 1988), although in mine waters they are often found at higher
concentrations. Trace elements that are of environmental concern are primarily metals and
metalloids. Their concentrations are not always easy to predict because in order to calculate
the concentration of a dissolved metal in equilibrium with a solid phase, the concentrations of
all potential complexing agents and the stability constants of the various possible complexes
must be known.
Under oxidizing conditions at high pH concentrations of dissolved trace metals that are
positively charged in solution are controlled by adsorption. Tessier et al. (1985) calculated
equilibrium constants for the adsorption of Cd, Cu, Ni, Pb and Zn onto iron oxyhydroxides in
oxic lake sediments. Adsorption substrates can include clays; organic matter; and iron,
manganese, aluminum and silicon hydrous oxides. The relative importance of each type of
substrate needs to be determined in order to identify what is controlling adsorption. Relative
importance of substrate type will depend on binding ability and abundance. Hydrous
manganese and iron oxides both have very high adsorption capacities and affinities for heavy
metals. When adsorption by manganese and iron oxides is a control on heavy metal
20
-------
concentrations in solution, the dissolved concentrations are a function of Eh and pH. Tessier
et al. (1985)*reported that adsorption of trace metals onto iron'oxyhydroxides typically
increases from near 0% to near 100% as the pH increases through a narrow critical range of
approximately 2 pH units (referred to as the adsorption edge). When the oxides are dissolved
by reduction, adsorbed metals are released. In reducing sediments trace element
concentrations in pore water are controlled by metal sulfide formation (Moore et al., 1988).
As pH increases, the negative surface charge of manganese and iron oxides increases and
metals are strongly adsorbed. Optimum adsorption occurs at high pe and pH (Drever, 1988).
Based on Eh-pH diagrams (Figure 5), in oxygenated waters dissolved iron concentrations are
expected to be low under neutral and alkaline conditions but high under acidic conditions
(Drever, 1988).
Decreasing concentrations of aluminum and iron correlate with increasing pH in
groundwater near Globe, Arizona, but not well enough to allow accurate prediction of their
concentrations based on pH alone (Stollenwerk and Eychaner, 1988). Aluminum solubility
appeared to be controlled by microcrystalline gibbsite at pH greater than 4.9. At pH less than
4.9 the authors propose that an assemblage of aluminum minerals controls Al solubility. The
solubility of iron is controlled by amorphous ferric hydroxide at all pH values.
In a stream receiving acidic effluents from a mine tailings deposit, suspended
particulate matter formed as pH increased from 3.5 to 6.5 due to mixing with groundwater of
pH greater than 7 (Karlsson et al., 1988). The particulate material consisted of FeOOH +
Fe(OH)3(am) and A1OOH + Al(OH)3(im) and silicates and was often associated with organic
matter. By the time pH was greater that 6.5 most of the iron and aluminum from the mine
effluent and most of the total metal content in the aqueous phase precipitated in the
particulate phase. About 50% of the Mn, Cu, Zn, Cd and Pb precipitated with the particulate
phase, mainly by adsorbing on the hydrous oxide precipitate. Mn and Pb also coprecipitated
to some extent. These processes gradually removed these metals from the aqueous phase with
increasing pH. Under ice-covered anoxic conditions some of the Fe stayed in the divalent
state reducing the amount of suspended matter that formed.
21
-------
Eh(V)
8
10 11
567
pH
Eh-pH diagram showing contours of dissolved iron
After Drever, 1988
Figure 5
22
-------
Acidophilic algae and bacteria from low pH environments have anionic cell walls and
can concentrate aqueous dissolved metals onto cell walls and intracellular sites (Mann et al.,
1990). Amorphous Fe and Ti concentrated at cell walls are bioprecipitated into
microcrystalline- aggregates of Ti- and Fe-oxyhydroxides which then act as scavengers for
heavy metals such as Cu, Pb, Zn, Ni, Cd and Th. The formation of the oxide minerals by
inorganic means is generally kinetically inhibited, so the biological catalysis of their
nucleation plays a major role in determining the aqueous metal concentrations in acid mine
drainage. Iron-oxidizing Thiobacilli can tolerate 0.37M aluminum, 0.15M zinc, 0.17M cobalt,
0.15M manganese, 0.16M copper, 0.1M chromium and 0.01M uranium. Lower tolerances are
found for silver (10'9M to 10'5M), mercury (0.05M) and molybdenum (0.03M). The bacteria
can not tolerate oxides of selenium, tellurium and arsenic (Lundgren and Silver, 1980).
The general conclusion that can be drawn about the aqueous geochemistry of trace
metals is that in most natural waters they are adsorbed out of solution at high pH due to
adsorption on iron oxyhydroxides, but remain in solution at low pH. Microorganisms can act
as scavengers for some heavy metals. Studies are needed in the existing water-filled pits
before generalizations about the short and long term fate of trace metals in pit water can be
made.
23
-------
Arsenic Speciation
:
v
Naturally occurring arsenic is found in groundwater in many parts of the western
United States, with moderate to high concentrations being common. Some of the highest
concentrations are associated with mining (Welch et al., 1988). Arsenopyrite (FeAsS) and
arsenic-bearing pyrite are often found with gold ore. Other sources of arsenic include
orpiment (As^) and realgar (AsS) and arsenic-rich iron oxides.
Arsenic has multiple valence states, and can form over 245 mineral compounds
making it difficult to evaluate because its behavior is complex (Lynch, 1988). Weathering of
rocks containing arsenic usually forms soluble arsenates with arsenic in the +5 oxidation state
(Boyle and Jonasson, 1973). Under anoxic conditions arsenic is usually found in the +3
state. In general, soluble As(HI) compounds are more toxic than As(V) compounds and
inorganic As compounds are more toxic than organic As compounds (Bitton and Gerba,
1984).
The major processes that control arsenic concentration and Speciation include mineral
precipitation and dissolution, solution composition, adsorption and desorption, ion exchange,
competing and complexing ions, chemical transformations, biologic activity, pH, Eh, aquifer
mineralogy, and reaction kinetics (Welch et al., 1988).
Amorphous ferric hydroxide has an extremely high adsorptive capacity for arsenic
(Pierce and Moore, 1982). However, the presence of materials such as humics, surfactants,
polyelectrolytes, silicic acid, phosphoric acid, silica particles, feldspar and montmorillonite
accelerate the growth of FeOOH crystals from amorphous, ferric hydroxide and can result in
the release of arsenic (Robins et al., 1988).
For effective removal of arsenic, the Fe/As mole ratio must be significantly greater
than one (Robins et al., 1988). Figure 6 shows arsenic remaining in solution versus pH, and
curves for increasing iron to arsenic ratios. More arsenic is removed from solution with
higher iron concentration due to adsorption on iron hydroxide and coprecipitation forming
minerals such as scorodite (FeAsO4*H2O).
24
-------
10
.1
.01
Ox
12345678
pH
Arsenic removal from 300 mg/1 As solution with six different
levels of Fe/As.
After Robins et al., 1988
Arsenic Metallurgy Fundamentals
and Applications
Figure 6
25
-------
The solubility of scorodite is lowest at pH of about 4 (Figure 7) (Krause and Ettel,
1989). When scorodite forms, arsenic is removed from solution. On an Eh-pH diagram
(Figure 8), the stability field for scorodite occurs in the high Eh, or oxidized range. Others
have also shown, that arsenic solubility and mobilization are minimized in oxidized conditions
with pH of 4-5 (Robins et al., 1988; Masscheleyn et al., 1991).
Ferguson and Gavis (1972) described the arsenic cycle in a stratified lake. In the oxic
epilimnion, reduced forms of arsenic are oxidized to arsenate which coprecipitates with ferric
hydroxide. Some arsenate is transported across the thermocline by turbulent dispersion and
convection. In the anoxic hypolimnion arsenate is reduced with the specific reduced arsenic
species formed depending on the sulfur concentration and the Eh.
When a lake is mixing, arsenic adsorbs on iron oxides on the surface of the sediments
and coprecipitates with hydrous iron oxides (Aggett and O'Brien, 1985; Mok and Wai, 1989).
When the sediment becomes reducing due to burial or increased biological activity near the
sediment/water interface, scorodite becomes soluble and Fe(IH) is reduced to Fe(II) which is
soluble at neutral pH's, so arsenic is released. The reduced iron and arsenic then can diffuse
upward until they are introduced to the lake water, or they can be reprecipitated due to
oxidation. During thermal stratification with anoxic conditions in the epilimnion, mobilization
of arsenic from the sediments occurs (Crecelius, 1975; Mudroch and Clair, 1986).
Thermodynamic calculations predict that at equilibrium in oxic water As(V) should be
the only stable oxidation state while As(IH) should be the stable form in anoxic systems.
However, arsenate and arsenite concentrations versus depth in a permanently stratified lake
did not reflect the expected thermodynamic equilibria (Seyler and Martin, 1989). While
As(V) was the dominant species in oxic waters, As(IH) was also present in significant
amounts. In anoxic waters, As(IH) was dominant but As(V) was still present. This suggests
26
-------
01
-h
r+
(D
OL oi
3 |»
<-» O>
o
§
I
o
rsi
O
O>
rt-
fs>
r»
W
Ol
As SOLUBILITY mg/l
8 *-* u. b o, S S8
IMOO
IM ID
Ol O
§0
o
s
§ i
-------
Eh
-0.5
Eh-pH diagram for the system Fe-As-H-O
snowing arsenic compounds at 25<>C.
After Mirza et al., 1988
Figure 8
28
-------
that the reduction rate from As(V) to As(ni) may be slow, causing incomplete response of the
As redox couple to reducing conditions. Similar disequilibrium speciation has been shown for
arsenic in the Baltic Sea (Andrieae and Froelich, 1984). One possible explanation of the
disequilibrium is that As(V) is reduced to As(in) by acting as an effective electron acceptor
in microbial mineralization of organic matter and that Fe(ni) competes for the role of electron
acceptor, thereby decreasing the rate of As(V) reduction (Masscheleyn et al., 1991).
Disequilibrium speciation of arsenic can also occur due to changes in pH. A diurnal
pH cycle due to photosynthesis in a perennial stream contaminated with arsenic caused a
similar diurnal cycle in arsenate concentration, although the arsenate cycle lagged behind the
pH cycle by several hours (Fuller and Davis, 1989). This indicates that the
sorption-desorption kinetics of arsenic are also complex and slow, resulting in disequilibrium.
In addition to affecting disequilibrium, biologic organisms also play other roles in the
speciation and solubilization of arsenic. Luong et al. (1981) showed that the presence of
Thiobacillus ferrooxidans increases the rate at which arsenic leaches from gold-bearing
material at least four fold over the abiotic leaching rate. Woolfolk and Whiteley (1962) and
Johnson (1972) showed that bacteria can reduce arsenate to arsenite. Wakao et al. (1988)
identified arsenic oxidizing bacteria that were aerobic and grew best between pH 3 and 4.
McBride and Wolfe (1971) identified a strain of bacteria that could reduce and methylate
arsenate to dimethylarsine under anaerobic conditions. Methylation of arsenic significantly
reduces its toxicity (Eisler, 1988a). Anderson and Bruland (1991) showed that methylated
arsenic is an important constituent in lakes. In a seasonally anoxic lake, dimethylarsenic acid
became the dominant form of dissolved As within the surface photic zone during late summer
and fall. When the lake turned over in early December methylated arsenic decreased while
arsenate increased, suggesting the dimethylarsenic acid degraded to arsenate. Anderson and
Bruland (1991) studied many lakes in California and Nevada and found methylated arsenic in
all lakes except one that was highly alkaline (Mono Lake).
The speciation and the solubility of arsenic in pit water depends on the chemistry of
the solution, redox state, pH, sorption and desorption, disequilibrium redox kinetics, and
biological contributions, among other factors. At or close to neutral pH's, arsenic will
29
-------
generally be soluble under anoxic conditions and will generally adsorb or coprecipitate with
iron under oxic conditions. The disequilibrium in redox kinetics causes arsenic to be present
in both the oxidized and the reduced form under any dissolved oxygen conditions. Optimum
removal of As species from solution occurs at pH values of about 4 to 5. Excess iron relative
to arsenic also minimized arsenic solubility. In general, oxygenated, non-alkaline conditions
cause minimum solubility and mobilization of arsenic.
Evapoconcentration
Because of the arid environment in which many mines in Nevada are situated,
evapoconcentration may have a big effect on the water chemistry in the pit lakes. As an
example of the magnitude of evaporation in Nevada, Big Soda Lake near Fallon in 1984 had
an evaporation rate of 120 cm/yr and rainfall of only 9 cm/yr (Kharaka et al., 1984).
Evaporation is a function of the vapor pressure above the water surface, the saturation vapor
pressure corresponding to the temperature of the water surface, the wind speed above the
water surface, wind shear and atmospheric stability. If the air in contact with the water is
warmer and has more moisture than the ambient air, then the air at the water surface will be
lighter and will tend to rise, creating an unstable condition. Evaporation will be greater with
increasing instability. Wind shear is a function of the water temperature and fetch distance.
Decreasing the water surface temperature increases wind shear and reduces evaporation.
Increasing the fetch distance increases evaporation (Blee, 1988).
Geochemical modeling of evapoconcentration of pit water in Barrick Goldstrike's
proposed Betze Pit resulted in an estimate of 35% increase in concentration of conservative
elements by the year 2100. The lake will reach a hydrologic steady state in about 200 years
with inflow from groundwater estimated at .05 m3/sec, outflow at .02 m3/sec and evaporation
at .03 m3/sec. The lake will reach a chemical steady state sometime after it reaches a
hydrologic steady state, and at that time concentrations of conservative elements will have
increased by a factor of 2.25 (BLM, 1991).
In order to predict the effects of evapoconcentration in arid environment pit water
lakes, information is needed about: the geometry of the drainage basins, local climate
30
-------
averages, the water balances in the pits, and the geochemistry of the pit waters. A numerical
simulation model that considers all of these factors and predicts the effects of
evapoconcentration should be developed.
Hydrothermal Activity
Nevada is located in a geologic area of high heat flow causing elevated groundwater
temperatures. Much of the geothermal activity is associated with fault systems.
Mineralization at mines is usually associated with fault systems too, so it follows that some
mine sites will have thermal groundwater. Garside and Schilling (1979) outlined the locations
of thermal waters throughout the state of Nevada.
The effects of hydrothermal activity on pit water lakes can be inferred from studies in
the literature on hydrothermal inputs to natural lakes. Yellowstone Lake in Wyoming is
oligotrophic and contains hydrothermal springs and gas fumaroles and has high geothermal
gradients within the lake bed (Klump et al,, 1988) The gases emitted by the fumaroles are
primarily carbon dioxide with traces of methane and hydrogen sulfide. The hydrothermal
waters (up to 70°C) are anoxic and high in dissolved nutrients and major ions. Mats of
microbial heterotrophs and photo- and chemolithotrophs and dense congregations of aquatic
worms, zooplankton and sponges were found surrounding some of the vents.
Instability of water in lakes is caused by differences in density, and temperature
affects density. At atmospheric pressure, water is at maximum density at 4°C. However, the
temperature of maximum density decreases with increasing pressure. Eklund (1963,1965)
defined the temperature of maximum density, T^, for a specific volume of fresh water, V, as
a function of pressure or depth as follows:
31
-------
V = V0[l-a+bO)P + cO2]
TMD = 4°C - O', O' = (b/2c)P = (.021°C/bar)P
=.(2.1 X 10-3°C/m)z
P = water pressure at depth z
a = 49.458 X lO'Vbar
b = .327 X 10-6/bar°C
c = 7.8 X lO'VC2
O = (4.00 - T)°C
Osbom and LeBlond (1974) and Chen and Millero (1977) developed a test of static stability
of water in lakes, taking into account the vertical variation of temperature, T, and salinity, S,
as follows:
2V0c(T-TMD)(dT/dz - dTA/dz) < (dV/dS)P(dS/dz)
dTA/dz = adiabatic temperature gradient =
T(2V0cpg/cp)(T-277.16 + pgzb/2c)
T = °K at depth z
p = mean water density above
cp = specific heat of water at constant temperature
Golubev (1978) showed that water temperatures in Lake Baikal at depths greater than
250 m are always less than 4°C. The bottom water (300-400 m deep) is less than 3.5°C.
Golubev found a localized region in Lake Baikal with a 1-2 m thick bottom water layer with
a temperature at least .5°C warmer than the overlying water. This represents a very unstable
heat flux due to hot spring discharge.
Williams and Von Herzen (1983) found a similar situation in Crater Lake, Oregon. At
about 295 m depth the temperature everywhere is less than 4°C. Below 295 m the deep water
temperatures increase with increasing depth. Thermal springs discharge into the deep water
and affect the temperature stratification, circulation and chemistry of the lake. Williams and
Von Herzen (1983) described thermal fluids discharging from a spring as fluid rising in a
plume and rapidly mixing the deep lake waters. The rest of the water is dense due to
dissolved solids and flows down slope where it mixes turbulently with the overlying lake
"water. Some of the water ponds on the lake floor where the mixing process is considerably
slower. The authors determined that the most important effect of the high heat input at Crater
Lake is that it caused the lake to mix vertically on a time scale of weeks to months.
32
-------
Williams and Von Herzen (1983) said that Crater Lake might be meromictic due to an
increase in suspended solids with depth, although the chemistry does not change with depth.
Thermal springs discharge about 6.35 X 109 g of dissolved solids into the lake annually. The
authors describe-the monimolimnion as well-mixed due to thermal convection. They said that
the deeper monimolimnion is not strongly influenced by periods of circulation and mixing in
the overlying mixolimnion, although it is not stagnant. The uniform chemical data show that
Crater Lake's waters are well-mixed laterally and vertically, so there must be significant
mixing between the mixolimnion and the monimolimnion. Williams and Von Herzen (1983)
describe the maximum density point as a natural top for the thermal convection system of the
monimolimnion. As thermal fluids rise (because they are less dense than the surrounding
fluids), they become diluted and cooled by the colder lake water. As they cool they approach
the temperature of maximum density at which they become negatively buoyant. Williams and
Von Herzen (1983) concluded that the standard limnologic definitions do not consider the
effects of terrestrial heat flow on the heating of deep lake water.
Hydrothermal input to pit water lakes will affect the temperature, stratification,
circulation and chemistry of the lakes. Mixing may be enhanced due to thermal convection
or suppressed by the input of dissolved solids which increase density. The elevated
temperatures will affect chemical speciation, if they are high enough, but only close to the
thermal springs because the spring water will cool as it is assimilated into the surrounding
water. Additional studies are needed on other existing lakes containing thermal springs in
order to better quantify the range, of potential effects of the springs and especially to
determine how they affect circulation of water in the lakes.
33
-------
Existing Pit Lakes
f
*
Examining some case studies of existing pit lakes can provide additional information
about the factors that affect pit water quality. Because precious metal open pit mining below
the water table is a relatively new phenomenon, most mines are still in production and
dewatering. Only a few such pit lakes exist. However, there are examples of water-filled
open pits from mining of other minerals such as uranium, phosphate, coal and copper.
Although different factors may affect pit lakes resulting from mining of different mineral
commodities, information relevant to understanding conditions in precious metal pit lakes can
be gleaned from other types of pit lakes.
Phosphate Mines
Many phosphate mine pit lakes are found in the state of Florida. As mining methods
changed, the shape of the lakes changed. Phosphate pits dug before 1920 were relatively
circular. This shape did not allow much littoral development, yet the lakes were productive
and supported large populations of game fish. The 1920's introduction of the dragline
allowed parallel excavations, 60-90 m wide, 600 m long and up to 20 m deep. Rules for the
reclamation of water bodies created by mining in Florida were intended to encourage the
development of lakes having shapes similar to natural lakes. These characteristics include
shallow basins, a well-developed littoral area, and a large portion of the water column in the
euphotic zone. Specifically, the regulations require (Florida Dept. of Natural Resources,
1975, cited by Pratt et al., 1985) that bottom slopes within 8 m of shore must not be steeper
than 4:1 (horizontal to vertical). To encourage littoral vegetation, at least 25% of the lake
surface must be within the zone of water fluctuation, or adjoining wetlands must be created.
To provide for fish bedding areas and submerged vegetation zones, at least 20% of the
surface must fall within a zone between the annual low water line and the minus 2 m annual
low water (Pratt et al., 1985). Florida's reclamation regulations also require trees along 50%
of the perimeter of reclaimed lakes (Ericson and Mills, 1986). Trees could provide nutrients
to aquatic biota in the lakes by dropping organic matter into the water. To allow for the
possibility that precious metal mine pit lakes could be used as fisheries, plans could be made
to leave benches at the proper elevation to provide littoral areas in the future pit lakes.
34
-------
Pratt et al. (1985) studied reclaimed phosphate pit lakes ranging in age from less than
one year to seven years and unreclaimed lakes over sixty years old. They found reclaimed
lakes to be dynamic systems resembling newly-formed reservoirs. These lakes tend initially
to be eutrophic with high concentrations of phosphate, nitrogen and trace minerals. Phosphate
pit lakes do not contain food webs as complex as those found in natural lakes because
biological introductions to the new lakes tend to be haphazard. Also, the biogeochemical
conditions in the lake are constantly changing during the first several years prior to
stabilization and this may pose a challenge to organisms. Precious metal mine pit lakes will
also receive haphazard biological introductions. Unlike in phosphate pits, phosphorous is
likely to be a limiting nutrient in precious metal pit lakes, so biological productivity would be
lower.
Uranium Mines
With uranium mines, there is often concern about radiation, so tailings are usually
covered and the pits backfilled. Water quality is generally not suitable for any use. The
Jackpile-Paguate uranium mine in west central New Mexico consists of three open pits which
are steep sided and partially filled with water since mining stopped in 1982 (Reith et al.,
1990). Water in each pit is slightly contaminated with trace metals, radionuclides, and
suspended sediments. Levels of contamination did not necessitate undertaking a treatment
program. Below-grade ore stockpiles were used as backfill in the pits. Radon barrier covers
consisting of 30 cm of shale and 60 cm of soil were placed on the ore-derived backfill
material.
The Nabarlek Uranium Project in the Northern Territory, Australia mined a small
high-grade ore body (McLoughlin et al., 1988). The open pit was used for tailings disposal
and contaminated water storage during milling operations. The pit had a net gain in water
storage each year due to seasonal high rainfall, and seepage from the pit occurred. To
prevent this, measures that exclude water were planned. After the tailings in the pit settle, the
mound over the surface of the pit will be leveled and a layered capping of clay and rock will
be placed, followed by topsoiling and revegetation.
35
-------
Uranium mining has produced many open pits in South Texas. Many of these are
^«
now water-filled and about 30 m deep. Kallus (1977) found pit water to have alpha and beta
sources of radiation and high concentrations of arsenic (average=.018 mg/1, maximum
value=.041 mg/1), and selenium (average=.022 mg/1, maximum value=.054 mg/1). Kallus
(1977) cited an unpublished master's thesis (Itin, 1975) that found these lakes to be unsuitable
water sources for human consumption, recreation, irrigation, stock and wildlife watering and
for fish and aquatic life. Itin (1975) found that the pit .lakes were not thermally stratified.
The water in the Whites Pit at the Rum Jungle Uranium Mine in Australia had a pH
of 4.75 in 1959, one year after mining ended, and a sulfate concentration of 180 mg/1. In
1974, after the addition of unneutralized tailings to the pit, the pH had decreased to 2.4 with
sulfate concentrations at 9000 mg/1. In 1974, the Intermediate Pit had a pH of 3.5 and a
sulfate concentration of 2000 mg/1. The Whites Pit lake is stratified with dissolved oxygen
concentrations of less than 1 mg/1 below 5 m (Goodman et al., 1981). A microbial study of
the Rum Jungle pits showed that the largest populations of Thiobacillus ferrooxidans occurred
in the sediments of the Whites Pit, indicating that they may be oxidizing sulfides
anaerobically, using something besides oxygen (such as iron) as a terminal electron acceptor.
Remediation at Rum Jungle is discussed in the Reclamation section of this paper.
Coal Mines
Both thermally stratified and non-stratified conditions exist in surface coal mine lakes
(Voelker, 1985). A coal pit water impoundment in Montana did not thermally stratify until
the fourth year after filling (Goering and Dollhopf, 1981). Lack of stratification in the first
three years may have resulted from larger inflow which produced more turnover and mixing.
Heat absorption by turbidity caused by reddish-black ferric oxides played a major role in
summer thermal stratification of coal strip mine lakes in Missouri (Parsons, 1977). Heat
budgets for turbid lakes were 65% less per unit volume than for clear lakes which did not
stratify.
Acidified coal strip mine lakes go through a series of successional stages in which the
acid, sulfate, Ca, Mg, Al,-and Fe concentrations are gradually reduced (Campbell and Lind,
36
-------
1969). This recovery has been referred to as the slow "titration" of the acidity by sulfate
reducing bacteria (Doyle, 1976). Decker (1971) showed that adding primary sewage sludge
added sufficient organic matter to greatly accelerate the recovery process. Sulfate reducing
bacteria utilize organic matter and release sulfide ions, which either combine with metal
cations and precipitate or combine with 2H+ at low pH to form H2S gas which is released to
the atmosphere (Doyle, 1976).
Acid mine water increases the rate of weathering of clay minerals, feldspars and
carbonates. Aluminum silicates produced by this mechanism tend to buffer pH (Kelly, 1988).
King et al. (1974) studied the restoration of acid lakes and found that below pH of 4.5
neutralization proceeded slowly. Once the aluminum buffer system was consumed then
neutralization was more rapid. In the pH range of 4.5 to 5.0 bacteria are no longer limited by
harsh acidic conditions so sulfate reduction occurs at a constant maximum rate (Doyle, 1976).
Concentrations of CO2 and H2S gases also increase as the bacteria growth rate
increases. Above pH = 6.4 (the pKj for carbonic acid) bicarbonate and hydrogen ions form
by dissociation:
CO2 + H20 --> H2CO3 -> H+ + HC03
This is a buffering system which maintains the pH near 6.4. Other buffers include HSO4",
H2CO3, H2S and ionic metals (especially Al3*). Each lake is unique since buffers will be
present in different concentrations and the initial amount of acid present will also vary.
When the pH is high enough for dissociation of carbonic acid from CO2, the bicarbonate
alkalinity system is established and phytoplankton and zooplankton can then form the base of
the food chain for the aquatic community. After an acidified lake has recovered, its water
chemistry is like that of an early eutrophic lake except that the sulfate:bicarbonate ratio is
high (Doyle, 1976).
Authigenic inorganic reduced sulfur minerals (mostly FeS2 and elemental S) are found
in the sediments of final-cut coal strip mine lakes (Wicks et al., 1991). The reaction for
sulfate reduction and precipitation of sulfides follows:
2Fe(OH)3(s) + SO42 + CH3COO' + H2 -->
FeS(s) + Fe2+ + 2HCO3' +-3H2O + 3OH"
37
-------
The products of this reaction, HCO3" and OH", are bases. In the strip mine lakes pore water
iron concentrations did not correlate with sediment sulfur content. Canadian Shield lakes,
however, had a strong pore water iron/sediment sulfur correlation (Carignan and Tessier,
1988). Wicks et al. (1991) attributed this difference to the availability of iron. The Canadian
Shield lakes are located on crystalline metamorphic bedrock overlain by glacial deposits, so
the source of iron is probably the weathering of crystalline bedrock. This means that a
resistant mineral is the source of the iron. Iron is released slowly, so as soon as dissolved
iron is delivered to the sediments it is consumed by the reaction. Conditions in the Canadian
Shield lakes are iron-limiting. The strip mine lakes, on the other hand, were located on
sandstone, shale and limestone bedrock. The source of iron is dissolved Fe in the pore water
of the sandstone and shale and very reactive Fe(OH)3 in the sediments of the overburden.
The sulfur mineral formation process is not iron or sulfate limited in strip mine lakes where
reactive iron and sulfate are abundant. Instead, the type of, and amount of, organic matter
controls the formation of reduced sulfur minerals.
The following can be learned from these coal lake studies that may pertain to precious
metal pit lakes:
~ stratification is unlikely during major inflow of water to the pit,
- turbidity caused by the precipitation of authigenic ferric oxides can change the
heat budget in a pit lake and cause stratification,
~ bacterial sulfate reduction can aid the recovery of acidified lakes by removing
potentially toxic metals and hydrogen ions, and
the rate of sulfate reduction can be limited by amount of organic matter in lakes
in sedimentary terrains, and limited by iron concentration in crystalline rock
terrains.
Copper Mines
Mining of the Berkeley Pit porphyry copper mine in Butte, Montana ended in 1982
and flooding of the pit began in 1983. The pit dimensions are 1.8 km by 1.4 km across and
542 m deep. In 1987 the water level in the pit was rising at a rate of 22 m per year. Davis
and Ashenberg (1989) estimate that the pit will overflow during the year 2009 and will
38
-------
intersect the alluvial aquifer by 1996. In October of 1987, geochemical sampling was
completed in. a depth profile down to 130 m in the middle of Ihe pit lake in an attempt to
characterize the aqueous solution in the pit and to define processes which may control metal
concentrations and distribution (Davis and Ashenberg, 1989).
The pH of the Berkeley Pit water ranges from 2.7 at the surface to 3.17 at depth
(Figure 9). Davis and Ashenberg (1989) attribute the lower pH at the surface to surface water
runoff. Huang and Tahija (1990) showed that the surface water source is an active tailings
pond and leach pads and that these surface sources supply most of the trace metal ions to the
pit. Dissolved oxygen decreases exponentially from 0 to 3 m, and below 3 m the system is
suboxic (Figure 10). Within the top 5 m most of the ferric iron is reduced and Fe(II)
becomes dominant (Figure 11). Ferric iron was below detection limits from 25 to 100 m
depth. Arsenic concentrations were low in the upper 15 m but increased with depth below 15
m (Figure 12). At the same depths that all of the iron is reduced, most of the arsenic occurs
in the +5 oxidation state. This is because at pH of 3, iron is reduced at an Eh of about .75 V,
while arsenic is reduced at an Eh of about .4 V. The field Eh profile for the Berkeley Pit
(Figure 13) goes from about .8 V at the surface to about .45 V at 130 m. The mass of solids
on the filter (Figure 14) is a measure of total suspended solids which increase
39
-------
1
V
0
20
40
60
2
z
>-
a.
i±f
0 80
100
120
1
BERKELEY PIT, BUTTE, MT
'
.
1
t
1 1
1 1 1 1
2 3 4 5 6 .7 8
PH
OaU from Davis and Ashenberg, 1989
Applied Geochemistry, v. 4.
Figure 9
40
-------
BERKELEY PIT, BUTTE, MT
- 60
E
80
120
7.5
Figure 10
Dissolved Oxygen (mg/1)
after Davis and Ashenberg
1989, Applied Geochem., v. 4.
41
-------
BERKELEY PIT, BUTTE, MT
i I I I I I I I
0 .1 .2 .3 .4 .5 .6 .7 .8 .9 1 1.1
Fe(II) and Fe(III) (ug/g)
Data from Davis and Ashenberg
1989, Applied Geochem., v. 4.
Figure 11
42
-------
BERKELEY PIT, BUTTE, MT
250500 750 1000
Dissolved As (ug/1)
after Davis and Ashenberg
1989, Applied Geochem., v. 4.
Figure 12
43
-------
BERKELEY PIT, BUTTE, MT «
20
40
60
80
100
120
J L
0 .1 .2 .3
.4 .5
FIELD Eh (v)
.6 .7 .8 .9
Data from Davis and Ashenberg, 1989
Applied Geochemistry, v.4.
Figure 13
44
-------
BERKELEY PIT, BUTTE, MT
20
40
60
80
100
120 .
I I I I J I i I I I I I I
.05
.1 .15
MASS OF SOLIDS ON FILTER (g)
.2
Figure 14
Data from Davis and Ashenberg, 1989
Applied Geochemistry, v. 4.
45
-------
TABLE 2
Pit Water Quality at Ruth
Arsenic
Barium
Cadmium
Chromium
Fluoride
Lead
Mercury
Nitrate
Selenium
Silver
Chloride
Copper
Iron
Manganese
pH
Sulfate
TDS
Zinc
Liberty
8/2/91
(rag/1)
<0.001
<.01
0.412
0.03
13
0.005
<0.0002
<0.1
<0.002
<.01
30
50.6
20.7
77
2.86
2860
3480
35
Ruth
9/30/86
(mg/1)
<0.004
<0.3
0.148
0.1
2.4
<0.01
<0.0001
0.02
<0.004
<0.01
6.6
31
2.94
108
3.23
2130
40
Ruth
8/2/91
(mg/1)
<0.001
0.02
<0.005
0.02
2.3
0.001
0.0002
2.4
0.04
0.04
44
17.5
0.17
0.611
8
1330
1820
0.041
Kimbley
9/24/91
(mg/1)
<0.180
0.009
<0.007
<0.010
2.61
<0.050
0.838
<1.0
<0.130
<0.020
264
0.172
0.455
0.31
7.59
1607
3310
2.43
Woodward-Clyde Consultants, 1992
47
-------
TABLE 3
Yerington Pit Water Quality
PH 8.06 8.21
TDS 638 628
Alkalinity: Total 143 134
Alkalinity :HC03 117 HO
Ca 49 230
Kg 14.3 22.3
K 16 6.9
Na 48.7 74
Cl 43 40
F 1.7 1-4
N03 as N 0.67 <0.5
SO* 240 242
Ac <.002 0.014
Ba 0.042 0.034
Cd <.002 0.008
Cr 0.004 0.02
Cu 0.731 0.232
Fe 0.581
Pb 0.011 0.012
Mn 0.09 0.076
Hg <.002 <.001
Se 0.004 <.002
Ag <.010 <.010
Zn <.030 0.081
Source: NDEP, 1991
48
-------
TABLE 4
U.S. EPA Drinking Water Standards
Constituent Max. Contaminant Level (mq/1)
Arsenic 0.05
Barium 1*0
Cadmium 0.010
Chromium 0.05
Lead 0.05
Mercury 0.002
Nitrate (as N) 10.0
Selenium 0.01
Silver 0.05
Fluoride 1.4-2.5(based on ave.ann.tmp.)
TDS 500
Chloride 250
Copper 1*0
Iron 0.3
Manganese 0.05
Sulfate 250
Zinc 5.0
pH 6.5-8.5
49
-------
The Brenda deposit in the Okanagan District in British Columbia is a molybdenum-
copper porphyry. The pit is 300 m across and more than 90 m deep. The pH of this pit is
7.3 (McCandless, 1992). Adjacent tailings pond water contained 1.5 mg Mo/1, a
concentration too high for use in irrigation, so the water was pumped into the pit. Dissolved
Mo concentrations in the pit range from 1.37 mg/1 at 1 m depth to 1.69 mg/1 at 45 m.
The War Eagle deposit in the Yukon Territory near Whitehorse is copper in skarn.
The pit is 26 m deep and 100 m across, and it filled in 12 years, beginning in 1971. No acid
was produced and the pH of the water is 7.8-8 (McCandless, 1992). Dissolved oxygen goes
to 0 below 12 m suggesting minimal turnover.
The copper pit lakes referenced in this section show the range in water quality that is
possible. They point out the importance of local geology and wallrock geochemistry in
determining ultimate pit water quality.
Precious Metals; Silver Mines
There are three pits at Equity Silver Mines Ltd. in British Columbia. The Southern
Tail Pit was mined first, and then was backfilled with material from the Main Zone Pit.
Patterson (1990) described the backfilling process, which occurred in three stages.
Backfilling to a horizon one meter below the projected flood plane was followed by
placement of a two meter layer of inert non-acid producing waste to serve as a buffer zone.
The third backfilling stage involved placing waste on top of the buffer layer and above the
water table. Since these wastes were also acid generating they were reclaimed to reduce
oxidation rates.
The pH of the Southern Tail Pit water from 1985 through 1989 in shown in Figure 15.
The pit water was initially neutral and then pH decreased to 3 by mid-1985. Back-filling
with material from the Main Zone Pit began in October, 1985. The Main Zone waste rock
had reactive neutralizing minerals, and the Southern Tail pit water pH rose to almost 6 within
50
-------
9-
8-
7 -
5. 6"
to
o
LU
| _
4 "
3 -
2
o
01
4-1
0)
._ 1
Ol
j*
u
"O
. to
,
I m»
**
. *
.*
t * '
, t
' ?
*."
***
»
'
B
1985 1986 1987 1988 1989
YEAR
SOUTHERN TAIL PIT, EQUITY SILVER MINES, E.G.
After Morin, 1990
Add Drainage from Mine Walls
The Main Zone Pit at Equity
Silver Mines
Figure 15
51
-------
a few months, then dropped again to below 3.4, increased to 7 by the end of back-filling in
1987 and varied from 7 to greater than 8 from 1987 to 1990. 'The development of acidic
conditions as the pit began to fill may be attributed to the combination of flushing stored acid
products.on the walls and within the unsaturated fracture networks downward to the bottom
of the pit and the decreasing flow of alkaline groundwater into the pit (Morin, 1990).
Copper, iron and zinc concentrations versus pH in the Southern Tail Pit are shown in Figure
16. As pH increases, iron concentrations decrease due to precipitation of FeOOH and copper
and zinc concentrations decrease due to adsorption on FeOOH, although there is a lot of
variation in the zinc data.
In 1990 the Main Zone Pit was to be mined out at a depth of approximately 200 m at
which time wastes from the third pit (the Waterline) would be backfilled. The pit would then
be flooded to cover the wastes. A dam was to be constructed at the pit entrance to raise the
water level so it would cover a portion of the wallrock that was acid generating.
Precious Metals: Gold Mines
The Nickel Plate Pit near Hedley, British Columbia contained a gold skarn deposit
The pit is 28 m deep, and less than 100 m across, and it filled in 3 years. No acid was
produced and the pH of the water is 7.8-8 (McCandless, 1992).
The Cortez Gold Mine in Nevada is in the Roberts Mountain Formation which is
limestone. This pit had an oxide ore body, and started filling with water in the early 70's.
Water quality data is shown in Table 5. The pit is currently 20-30 m deep and in the early
80's bass were planted and still survive today. The area adjacent to this pit is being mined
again so fishing is not permitted now, but it was allowed in the past. The fish are not fed by
mine personnel, indicating that there is enough primary productivity in the pit to support a
full food chain that sustains the fish. There are also reeds growing at the edge of this pit
(List, 1992).
52
-------
O
M
0»
Ui
LOG (dissolved Zn in Hg/1)
S
8
(£('
«»
SB'S?
«
r
LOG (dissolved Cu in Mg/1)
i i
<
S
« ' O
LOG (dissolved Fe in Hg/1)
^^ 2J- "T ? T Y
-------
TABLE 5
Cortez Gold Mine Pit Water Quality
sampled 9/13/90, unfiltered sample
Element Cone, (mg/1) Element Cone, (mg/1)
Ag 0.005 Mo 0.013
Al <« Na 68.100
As <« Ni <«
Au <« Pb <«
B 0.327 Pd <«
Ba 0.061 Pt <«
Be <« S 30.500
Ca 45.100 Sb <«
Cd <« Se <«
Co 0.004 Si 14.200
Cr <« Sn <«
Cu <« Sr 0.778
Fe <« Te <«
Hg 0.081 Ti 0.012
K 11.500 Tl 0.053
Li 0.212 V <«
Mg 19.400 W <«
Mn 0.001 Zn <«
<« indicates less than detection limit
M. List, Cortez Mine
Personal communication, 1992
54
-------
Effects of Pit Water Quality on Life
.
>>
The bass living in the pit lake at Cortez Gold Mine, without being fed by humans,
indicate that there are enough naturally occurring nutrients in the pit water for them to
survive. In addition, explosives used for excavating mines can later contribute nitrogen as
nutrients (Morin, 1988). If all of the necessary nutrients are present in the pit water, the other
concern for survival of aquatic biota is water quality and the concentration of elements that
are toxic. It is difficult to determine safe levels of metals for aquatic organisms because the
level at which a metal becomes toxic to a particular species depends on time of exposure,
temperature, dissolved oxygen concentration, pH, salinity, hardness, water velocity, and the
interaction between combinations of metals (U.C. Berkeley Mining Waste Study Team, 1988).
The effects of various aqueous pH levels on biota have been studied (Appalachian
Regional Commission, 1969; Bell, 1971). Also, the effects of elevated concentrations of
various elements on fish, wildlife and invertebrates have been reviewed (Eisler, 1985a, 1985b,
1986, 1987, 1988a, 1988b, 1989, 1990).
Bell (1971) tested mature larvae and nymphs of 9 species of aquatic insects
(dragonflies, stoneflies, caddisflies and mayflies) in the laboratory for their tolerance to low
pH. All of the tested species have high value as fish food. In general, Bell found that
caddisflies are very tolerant of low pH (30-day TL 50 as low as pH = 2.45). The 30-day TL
50 is the pH at which 50% of the organisms died after 30 days. Stoneflies and dragonflies
are moderately tolerant (pH 3.71 to 5.0), and the mayflies are sensitive (30-day TL 50 at pH
= 5.38). The more sensitive insects will be limited in numbers and species composition under
prolonged acid conditions. In addition, Bell found that under low pH conditions the
percentage of aquatic insects which emerge successfully also decreases. The pH at which
50% successful emergence takes place ranges from 0.52 to 2.10 pH units higher than the 30-
day TL 50 value for the species tested. These aquatic insects were generally more tolerant
than fish however. The Appalachian Regional Commission (1969) said that most data
indicate that fully developed adult fish can live in waters of pH ranging from 5.0 to 9.0.
Below pH of 5.0 the productivity of aquatic ecosystems is considerably reduced.
Copper, zinc and cadmium accumulate in fish livers in a portion of the Sacramento
River that receives acid drainage from mining areas (Wilson et al., 1981). The metals did not
55
-------
accumulate in fish flesh. Aqueous metal concentrations at one sample location were .051 mg
Cu/1, .214 mg Zn/1, and .0023 mg Cd/1. LC-50's for various aquatic organisms range from
.0006 to 250 mg/1 for cadmium, from .055 to 60.2 mg/1 for zinc, and from .005 to 13.9 mg/1
for copper (U.C. Berkeley Mining Waste Study Team, 1988).
Cadmium is extremely toxic to rainbow trout (Ball, 1967). At temperatures between
11.0 and 12.5°C at least 50% mortality occurred at concentrations between .01 and 1.0.mg
Cd/1. Concentrations between .008 and .01 mg Cd/1 were determined to be lethal to rainbow
trout after 7 days. Although the EPA drinking water standard for Cd is .01 mg/1 (Lehr et al.,
1984), Eisler (1985a) also cited lethal and sublethal effects on freshwater aquatic life at much
lower concentrations. .0008 to .0099 mg/1 of Cd was lethal to several species of aquatic
insects, crustaceans, and teleosts. .0007 to .005 mg/1 caused sublethal effects such as
decreased growth, inhibited reproduction, and population alterations. When Cd concentration
exceeds .003 mg/1 in freshwater, adverse effects on fish or wildlife are either pronounced or
probable (Eisler, 1985a). Adverse effects are most pronounced in waters of low alkalinity.
Brook trout suffered reduction in growth, survival and fecundity in water of low alkalinity
with Cd concentrations between .001 mg/1 and .003 mg/1. With increasing alkalinity the
maximum allowable Cd concentrations increased to between .007 mg/1 and .012 mg/1.
Selenium concentrations between .06 and .6 mg/1 caused sensitive species of aquatic
organisms to die (Eisler, 1985b). Freshwater algae species may fare better if sulfate is
present, because for them sulfate has a protective role against Se tpxicity. The U.S. EPA
drinking water standard for selenium is .01 mg/1 (Lehr et al., 1984).
The freshwater organisms that were most sensitive to chromium(VI) are crustaceans
and rotifers (Eisler, 1986). Reduced growth, inhibited reproduction, and other adverse effects
were found at .01 mg/1 Cr6* and at .03 mg/1 Cr*. The U.S. EPA drinking water standard for
chromium is .05 mg/1 Cr** and 170 mg/1 Cr3* (Lehr et al., 1984 and Eisler, 1986).
Lethal concentrations of mercury for aquatic organisms range from .0001 to .002 mg/1
(Eisler, 1987). .002 mg/1 is the U.S. EPA drinking water standard (Lehr et al., 1984). Eisler
also reported significant adverse sublethal effects in aquatic birds at .00003 to .0001 mg Hg/1.
The U.S. EPA 1985 criteria for protection of freshwater aquatic life calls for a maximum 4-
day average of .000012 mg/1, not to exceed an hourly average of .0024 mg/1 (Eisler, 1987).
56
-------
Eisler claims that these criteria offer only limited protection to aquatic organisms. One of the
reasons mercury has high toxicity is that its low toxicity forms can be transformed into forms
of very high toxicity through biological processes such as methylation. Mercury also is
bioconcentrated in organisms and biomagnified through food chains.
Arsenic is bioconcentrated but not biomagnified in the food chain (Eisler, 1988a). In
contrast to mercury, methylation of arsenic greatly reduces its toxicity. The aquatic chemistry
of arsenic has been discussed previously in this paper. Important points to remember are that
inorganic As is more toxic than organic As and that As3* is more toxic than As5*. Both
inorganic As and As3* are also the most mobile forms. Eisler reported that sensitive aquatic
species were damaged at .019 to .048 mg As/1. The U.S. EPA 1985 drinking water standard
is .05 mg/1 (Eisler, 1988a). The 1985 EPA criteria for protection of freshwater aquatic life is
.19 mg As34/!. Eisler points out that the effects of chronic low exposure on reproduction,
genetic makeup, adaptation, disease resistance, growth, etc. have not been, but need to be,
studied.
Adverse effects of lead on aquatic biota were found between .001 and .0051 mg/1.
Daphnids were the most sensitive organisms, showing adverse effects on reproduction at .001
mg/1 (Eisler, 1988b). The U.S. EPA drinking water standard for lead is .05 mg/1 (Lehr et al.,
1984). Organic lead compounds are more toxic than inorganic. In water, lead is most soluble
and most bioavailable at low pH, low organic content, low suspended sediment, and low salts
of Ca, Fe, Mn, Zn, and Cd.
Aquatic organisms are resistant to molybdenum salts adverse effects on growth and
survival usually only occur at greater than 50 mg Mo/1 (Eisler, 1989). Mo is bioconcentrated
by selected species of algae and invertebrates but the effect of bioconcentration on higher
trophic level organisms is not known. Freshwater fishes are extremely resistant to Mo, but
50% of fertilized eggs of rainbow trout died in 28 days at only .79 mg Mo/1. There are
currently no federal drinking water standards for molybdenum. Eisler (1989) proposed
criteria of less than .05 mg Mo/1.
Borax (Na2B4O5(OH)4*H2O) forms during evaporation of enclosed lakes and as an
efflorescent mineral on the land surface in arid regions (Hurlbut and Klein, 1977). Boron
compounds tend to accumulate in aquatic ecosystems because they are highly soluble. The
57
-------
U.S. EPA boron criteria for protection of aquatic life is .55 mg/1 (Eisler, 1990). Eisler (1990)
reported thakconcentrations greater than .1 mg B/l may affect'reproduction in rainbow trout
and greater than .2 mg/1 may impair survival of other fish species, but additional data is
needed. Boron compounds are more toxic to embryos and larvae than adults.
This discussion has mentioned some but not all of the trace metals that may occur in
precious metal pit water in toxic proportions. As mentioned, metal toxicity levels vary
depending on species as well as on time of exposure, temperature, oxygenation of the water,
pH, salinity, hardness, water velocity and interactions between combinations of metals. The
U.S. EPA Drinking Water Standards are often used as the basis to determine pit water quality.
Although pit water will rarely be used as drinking water, this discussion has shown that the
metal toxicity levels for various aquatic organisms are often below the drinking water
standard concentrations. If the U.S. EPA Drinking Water Standards continue as the basis for
comparison for pit water, aquatic ecosystems will not be protected. Although not discussed in
detail here, the effects of pit water quality on waterfowl also need to be seriously considered.
It will be impractical, if not impossible, to use netting and/or audio hazing techniques to keep
birds out of the pit lakes after mine decommissioning.
58
-------
Reclamation
>
t
Various methods have been proposed for remediation of the water quality in the
Berkeley Pit (Davis and Ashenberg, 1989; Huang and Tahija, 1990). Those methods involve
in situ neutralization. Other pits have been reclaimed by pumping the water to a treatment
plant. The reclamation phase of metal mining in sulfidic areas may be the most difficult
phase in which to control potentially detrimental effects on the environment, and pit water in
such an area might have to be treated in perpetuity in order to avoid contamination of
groundwater and surface water (Bell and Nancarrow, 1974).
The Rum Jungle uranium mine in the Northern Territory, Australia had three acidic
water-filled pits (Dyson's, White's and Intermediate) that have since been reclaimed using
three different methods (Harries and Ritchie, 1988). Dyson's open pit had tailings placed at
the bottom of the pit. The top surface of the tailings was sloped to discharge water and then
covered with a geotextile fabric and a "rock blanket" 1 m thick to carry away groundwater
and seepage water. Material from the heap leach and the most contaminated subsoils were
placed on top of the rock blanket (above the water table). This was covered with a low
permeability geotextile sealing layer, a moisture retention layer, and another sealing layer on
top.
Water in both the White's and Intermediate open pits was treated by hydroxide
precipitation to raise the pH and remove trace metals (in situ neutralization). The water in
White's open pit was then pumped through a treatment plant and returned to the open pit
where stratification occurred due to separation of the treated and untreated water. Further
treatment of the Intermediate open pit was in situ with the addition of lime, and aeration to
ensure mixing. Then the water was allowed to settle and metal hydroxide precipitates were
pumped from the bottom of the open pit to the treatment plant. Treatment sludge from the
treatment plant had to be buried. Higgs (1990) discussed the chemical stability and disposal
considerations of treatment sludge. In general, sludges consist of metal hydroxides and
gypsum. They are not always stable during storage and constituents may redissolve. If
sludges will be produced by a treatment process, their storage or disposal is an important
consideration.
59
-------
The ideal situation for reclamation of pit lakes would be to be able to predict pit water
%»
quality so well that upon closure, but before the pit begins filling, plans could be made to
protect the future water quality. This might include covering reactive materials, or adding a
layer of neutralizing material to the pit bottom. If arsenic is anticipated to be a problem, the
pH may need to be lowered with the knowledge that other elements may then become a
problem. More ideas are needed in this new area and each mine will need a unique solution.
Planning ahead could prevent an expensive reclamation problem.
60
-------
Conclusion
^v
Many of the open pit precious metal mines that are scattered throughout Western
North America, with a high concentration in Nevada, will ultimately become pit lakes. The
lakes may be me'fomictic if total dissolved solids concentrations are high enough. The lakes
may be stratified or mixing or have periods of both depending on the local climate and on
morphometry of the pits.
It is important to know the regional groundwater flow system prior to mining since
flow directions and water table elevations have direct bearing on the characteristics of the pit
lake. Water may be expected to seep from the pits during periods of high rainfall or surface
water runoff, and to flow back into pits through aquifers during periods of high evaporation.
Subsidence can occur within the cone of depression during dewatering, leading to an altered
groundwater flow regime when dewatering ends.
Reactions of water with exposed wallrock in an open pit mine are major contributors
to pit water quality. Flow in joints and fractures makes calculation of exposed surface area
difficult. Sloughing of wallrock occurs, adding new surface area for reaction.
The pH of pit water is determined by the mineralogy of the surrounding rocks.
Sulfide minerals, when exposed to oxygen and water, oxidize in a series of reactions that
produce acid. Bacteria can greatly accelerate the rate of the oxidation reactions. Dissolution
of minerals such as calcite can neutralize acid. The balance between acid producing minerals
and acid consuming minerals is one measure of the potential for acid production. Other tests
consider rates of oxidation and neutralization reactions. All of the tests have limitations that
need to be recognized before they are used to make predictions. Control of acid generation
and migration and/or treatment of acid can be difficult undertakings that need to be planned
for well in advance.
*
Trace element concentrations are often high in waters associated with mining. Most
trace elements are soluble at low pH, but adsorb onto iron oxyhydroxides which form at high
pH. Arsenic is a trace element that naturally occurs in high concentrations in groundwater in
the western United States. In pit water at or close to neutral pH's, arsenic will generally be
soluble under anoxic conditions and will generally adsorb or coprecipitate with iron under
61
-------
oxic conditions. Optimum removal of arsenic species from solution occurs at pH values of 4
to 5 with excess iron concentration relative to arsenic.
Because of the arid climate in which most precious metal open pit mines are located,
evapoconcentration will increase the concentration of trace elements in pit water.
Hydrothermal input to pit lakes will affect their temperature, stratification, circulation and
chemistry.
Data from existing water-filled pits from phosphate, uranium, coal, copper, silver and
gold mining yields valuable information about conditions that can be expected in pit lakes.
However, more limnological studies are needed in these existing pit lakes, to collect data
versus depth at different times of the year to determine seasonal patterns. All pit water data
collected should include a statement of how the samples were collected, stored and analyzed
including ionic balances so adequate modeling can be done.
Effects of pit water quality on aquatic biota, birds and humans need to be considered.
Water quality standards that protect wildlife are generally lower than the drinking water
standards, making the latter an inadequate predictor of the ultimate quality of the water for
the survival of aquatic biota. Reclamation of pit water can be difficult, unless poor water
quality was anticipated and preventive measures were undertaken.
Figure 17 shows the geology of Nevada's future and current pit lakes. By looking at
data from existing pit lakes, we saw that there is enough variability in pit water quality
among the porphyry copper deposits ~ Butte (Montana), Ruth and Yerington ~ to show that
gross generalizations about geology are not enough to predict pit water quality. It may be
safe to assume, however, that the sedimentary hosted deposits generally have calcite
associated with them, and have lower sulfide content than the porphyries, so acid may not be
a problem. Arsenic concentrations in pit water will often be high due to abundant naturally
occurring arsenic in Nevada. Each pit is a unique system and should be studied as such.
62
-------
ttuMIIQ O1"*11 "* Urrirt
t
oorphyry
Igneous hosted
O sedlmenUry hosted
km
100
Some locations
are approximate
Figure 17
63
-------
References
Aggett, J. and G.A. O'Brien, 1985, Detailed model for the mobility of arsenic in
lacustrine sediments based on measurements in Lake Ohakuri, Environmental Science
and Technology, v. 19, p. 231-238.
Anderson, L.C.D. and K.W. Bruland, 1991, Biogeochemistry of arsenic in natural waters:
The importance of methylated species, Environmental Science and Technology, v 25
p. 420-427.
Andreae, M.O. and P.N. Froelich, 1984, Arsenic, antimony, and germanium
biogeochemistry in the Baltic Sea, Tellus, v.36B, p. 101-117.
Appalachian Regional Commission, 1969, Acid Mine Drainage in Appalachia, publ. in
Washington, D.C.
Ball, I.R., 1967, The toxicity of cadmium to rainbow trout, Water Research, v. 1, p. 805-
806.
Bear, J., 1979, Hydraulics of Groundwater, McGraw-Hill, 569 p.
Bell, A.V. and D.R. Nancarrow, 1974, Salmon and mining in northeastern New
Brunswick (A summary of the Northeastern New Brunswick Mine Water Quality
Program), Canadian Mining and Metallurgy Bulletin, v. 67, p. 44-53.
Bell, H.L., 1971, Effect of low pH on the survival and emergence of aquatic insects, Water
Research, v. 5, p. 313-319.
Bitton, G. and C.P. Gerba, 1984, Groundwater Pollution Microbiology, (Wiley
Interscience), p. 163-164.
Blee, J.W.H., 1988, Determination of evaporation and seepage losses, Upper Lake Mary
near Flagstaff, Arizona, U.S. Geological Survey, Water Resources Investigations
Report 87-4250, 39 p.
BLM (U.S. Bureau of Land Management), 1991, Final Environmental Impact Statement,
Betze Project.
Boyle, R.W. and I.R. Jonasson, 1973, The geochemistry of arsenic and its use as an indicator
element in geochemical prospecting, Journal of Geochemical Exploration, v.2, p.
251-296.
64
-------
Broadbent, C.D. and Z.M. Zavodni, 1981, Influence of rock structure on stability, in C.
Brawner, ed. Proceedings of the Third International Conference on Stability in Surface
Mining, Vancouver, B.C., June 1-3, 1981, p. 7-18.
Brown, A., 1982, The influence and control of groundwater in large slopes, in C.O. Brawner,
ed. Proceedings of the Third International Conference on Stability in Surface Mining,
Vancouver, B.C., June 1-3, 1981, p. 19-41.
Campbell, R.S. and O.T. Lind, 1969, Water quality and aging of strip-mine lakes, Journal
of the Water Pollution Control Federation, v. 41, p. 1943-1955.
Carignan, R. and A. Tessier, 1988, The co-diagenesis of sulfur and iron in acid lake
sediments of southwestern Quebec, Geochimica et Cosmochimica Acta, v. 52, p. 1179-
1188.
Chander, S. and A. Briceno, 1988, The rate of oxidation of pyrites from coal and ore
sources ~ an AC impedance study, in Mine Water and Surface Mine Reclamation,
Volume I: Mine Water and Mine Waste, U.S. Bureau of Mines Information Circular
9183, p. 164-169.
Chen, C. and F.J. Millero, 1977, Effect of salt content on the temperature of maximum
density and on static stability in Lake Ontario, Limnology and Oceanography, v. 22, p.
158-159.
Crecelius, E.A., 1975, The geochemical cycle of arsenic in Lake Washington and its relation
to other elements, Limnology and Oceanography, v.20, p. 441-451.
Davis, A. and D.D. Runnels, 1987, Geochemical interactions between acidic tailings fluid
and bedrock: use of the computer model MINTEQ, Applied Geochemistry, v. 2, p.
231-241.
Davis, A. and D. Ashenberg, 1989, The aqueous geochemistry of the Berkeley Pit, Butte,
Montana, USA, Applied Geochemistry, v. 4, p. 23-36. /
Decker, C.S., 1971, Accelerated Recovery of Acid Strip Mine Lakes, M.S. Thesis,
University of Missouri.
diPretoro, R.S. and H.W. Rauch, 1988, Use of acid-base accounts in premining
prediction of acid drainage potential: a new approach from northern West Virginia, in
Mine Drainage and Surface Mine Reclamation, v. I: Mine Water and Mine Waste,
U.S. Bureau of Mines Information Circular 9183, p. 2-10.
Doyle, P.M., 1991, University of California at Berkeley, oral communication.
65
-------
Doyle, W.S., 1976, Recovery of Acid Strip Mine Lakes, in Strip Mining of Coal:
Environmental Solutions, Pollution Technology Review No. 27, Noyes Data Corp.,
New Jersey, p. 200-223.
Drever, J.I., 1988, The Geochemistry of Natural Waters, Prentice Hall, New Jersey,
437 p.
Eklund, H., 1963, Fresh water: Temperature of maximum density calculated from
compressibility, Science, V. 142, p. 1457-1458.
Eklund, H., 1965, Stability of lakes near the temperature of maximum density, Science, v
149, p. 632-633.
Eisler, R., 1985a, Cadmium hazards to fish, wildlife and invertebrates: A synoptic review,
U.S. Fish and Wildlife Biological Report 85(1.2), 46 p.
Eisler, R., 1985b, Selenium hazards to fish, wildlife and invertebrates: A synoptic review,
U.S. Fish and Wildlife Biological Report 85(1.5), 57 p.
Eisler, R., 1986, Chromium hazards to fish, wildlife and invertebrates: A synoptic review,
U.S. Fish and Wildlife Biological Report 85(1.6), 60 p.
Eisler, R., 1987, Mercury hazards to fish, wildlife and invertebrates: A synoptic review,
U.S. Fish and Wildlife Biological Report 85(1.10), 90 p.
Eisler, R., 1988a, Arsenic hazards to fish, wildlife and invertebrates: A synoptic review,
U.S. Fish and Wildlife Biological Report 85(1.12), 92 p.
Eisler, R., 1988b, Lead hazards to fish, wildlife and invertebrates: A synoptic review,
U.S. Fish and Wildlife Biological Report 85(1.14), 134 p.
Eisler, R., 1989, Molybdenum hazards to fish, wildlife and invertebrates: A synoptic review,
U.S. Fish and Wildlife Biological Report 85(1.19), 61 p.
Eisler, R., 1990, Boron hazards to fish, wildlife and invertebrates: A synoptic review,
U.S. Fish and Wildlife Biological Report 85(1.20), 32 p.
Erickson, P.M. and R.S. Hedin, 1988, Evaluation of overburden analytical methods as
means to predict post-mining coal mine drainage quality, in Mine Drainage and
Surface Mine Reclamation, v. I: Mine Water and Mine Waste, U.S. Bureau of Mines
Information Circular 9183, p. 11-20.
66
-------
Ericson, W.A. and J. Mills, 1986, Florida's nonmandatory reclamation of phosphate lands,
Current regulations and their implementation, in D.H. Graves, ed., Proceedings of the
1986'Symposium on Mining Hydrology, Sedimentology, and Reclamation, University
of Kentucky, Lexington, December 7-11, 1981, UKY BU142, p. 191-196.
Ferguson, K.D. and P. M.:Erickson, 1988, Pre-mine prediction of acid mine drainage, in
Environmental Management of Solid Waste: Dredged Material and Mine Tailings,
Springer-Verlag, New York, p. 24-43.
Ferguson, J.F. and J. Gavis, 1972, A review of the arsenic cycle in natural waters, Water
Research, v.6, p. 1259-1274.
Fuller, C.C. and J.A. Davis, 1989, Influence of coupling of sorption and photosynthetic
processes on trace element cycles in natural waters, Nature, v.340, p. 52-54.
Garside, L.J. and J.H. Schilling, 1979, Thermal waters of Nevada, Nevada Bureau of Mines
and Geology, Bulletin 91, 163 p.
Glass, C. E., 1982, Influence of earthquakes on rock slope stability, in C.O. Brawner, ed.
Proceedings of the Third International Conference on Stability in Surface Mining,
Vancouver, B.C., June 1-3, 1981, p. 89-112.
Goering, J.D. and DJ. Dollhopf, 1981, Water impoundments in mined land spoil
material, Big Sky Mine Montana, in D.H. Graves, ed., Proceedings of the 1981
Symposium on Surface Mining Hydrology, Sedimentology and Reclamation,
University of Kentucky, Lexington, December 7-11, 1981, UKY BU142, p. 331-335.
Golubev, V.A., 1978, Vertical temperature gradients and the static stability of the bottom
water of Lake Baikal, Doklady Akad. Nauk SSSR, v. 239, p. 3-5.
Goodman, A.E., A. M. Khalid arid B.J. Ralph, 1981, Microbial ecology of Rum Jungle,
II: Environmental study of two flooded open cuts and smaller associated water bodies,
Australian Atomic Energy Commission Establishment, Report AAEC/E-527.
Hammack, R.W., 1985, The relationship between the thermal activity of pyrite and the
rate of acid generation, in D.H. Graves, ed., Proceedings of the 1985 Symposium on
Surface Mining Hydrology, Sedimentology, and Reclamation, p. 139-144.
67
-------
Hammack, R.W., R.W. Lai, and J.R. Diehl, 1988, Methods for determining fundamental
chemical differences between iron disulfides from different geologic provenances, in
Mine Drainage and Surface Mine Reclamation, v. I: Mine Water and Mine Waste,
U.S. Bureau of Mines Information Circular 9183, p. 136-146.
Harries, J.R. and- A.I.M. Ritchie, 1988, Rehabilitation Measures at Rum Jungle Mine Site, in
Environmental Management of Solid Waste: Dredged material and Mine Tailings,
Springer-Verlag, New York, NY, p. 131-151.
Higgs, T.W., 1990, ARD treatment plant sludge ~ chemical stability and disposal
considerations, in Acid Mine Drainage, Designing for Closure, Proceedings of the
GAC/MAC Joint Annual Meeting, Vancouver, B.C., May 16-18, 1990, p. 427-440.
Hoek, E., 1970, Influence of rock structure on the stability of rock slopes, in C.O. Brawner
and V. Milligan, eds., Stability in open pit mining, Proceedings of the First
International Conference on Stability in Open Pit Mining, Vancouver, B.C.,
November 23-25, 1970, p. 49-63.
Holmberg, R. and K. Maki, 1982, Case examples of blasting damage and its influence on
slope stability, in C.O. Brawner, ed. Proceedings of the Third International Conference
on Stability in Surface Mining, Vancouver, B.C., June 1-3, 1981, p. 773-793.
Huang, H. and D. Tahija, 1990, Characteristics and treatment problems of surface and
underground waters in abandoned mines at Butte, Montana, in P.M. Doyle, ed.,
Mining and Mineral Processing Wastes, Proceedings of. the Western Regional
Symposium on Mining and Mineral Processing Wastes, Berkeley, California, May 30-
June 1, 1990, p. 261-270.
Hurlbut, C.S. and C. Klein, 1977, Manual of Mineralogy, 19th Edition, John Wiley and
Sons, New York, 532 p.
Hutchinson, G.E., 1957, A Treatise on Limnology, v. 1, John Wiley and Sons, New York,
1015 p.
Itin, S.C., 1975, The public health significance of abandoned open-pit uranium mines in
South Texas, Master's thesis, University of Texas, (unpublished).
Johnson, D.L., 1972, Bacterial reduction of arsenate in seawater, Nature, v.240, p. 44-45.
Kallus, M.F., 1977, Environmental impacts of uranium mining in South Texas, in M.D.
Campbell, ed., Geology of Alternate Energy Resources in the South-Central United
States, Houston Geological Society, p. 83-102.
68
-------
Karlsson, S., B. Allard and K. Hakansson, 1988, Characterization of suspended solids in a
stream receiving acid mine effluents, Bersbo, Sweden, Applied Geochemistry, v.3, p.
345-356.
Kelly, M., 1988, Mining and the Freshwater Environment, Elsevier Applied Science, London
and New York, 231 p.
Kharaka, Y.K., Robinson, S.W., Law, L.M. and Carothers, W.W., 1984,
' Hydrogeochemistry of Big Soda Lake, Nevada: An alkaline meromictic desert lake,
Geochimica et Cosmochimica Acta, v. 48, p. 823-835.
King, D.L., J.J. Simmler, C.S. Decker and C.W. Ogg, 1974, Acid strip mine lake recovery,
Journal of the Water Pollution Control Federation, v. 46, p. 2301-2315.
Kleinmann, R.L.P., D.A. Crerar and R.R. Pacelli, 1981, The biogeochemistry of acid mine
drainage and a method to control acid formation, Mining Engineering, March, 1981, p.
300-305.
Klump, J.V., R. Paddock, P.A. Anderson, C.C. Remson and J. Maki, 1988, Hydrothermal
activity in Yellowstone Lake, WY: Preliminary observations of a unique lacustrine
environment, EOS, v. 69, no. 44, p. 1109, abstract only.
Krause, E. and V.A. Ettel, 1989, Solubilities and stabilities of ferric arsenate compounds,
Hydrometallurgy, v. 22, p. 311-337.
Lapakko, K., 1990, Solid phase characterization in conjunction with dissolution
experiments for prediction of drainage quality, in Mining and Mineral Processing
Wastes, F.M. Doyle, ed., Proceedings of the Western Regional Symposium on Mining
and Mineral Processing Wastes, Berkeley, California, May 30-June 1, 1990, p. 81-86.
Lawrence, R.W., 1990, Prediction of the behavior of mining and processing wastes in the
environment, in Mining and Mineral Processing Wastes, F.M. Doyle, ed., Proceedings
of the Western Regional Symposium on Mining and Mineral Processing Wastes,
Berkeley, California, May 30-June 1, 1990, p. 115-121.
Lehr, J.H., D.M. Nielsen, and J.J. Montgomery, 1984, U.S. Federal legislation pertaining
to groundwater pollution, in G. Bitton and C.P. Gerba, eds., Groundwater Pollution
Microbiology, (Wiley Interscience), p. 353-371.
List, M., 1992, Cortez Gold Mine, oral and written communication.
Lundgren, D.G. and M. Silver, 1980, Ore leaching by bacteria,Annual Review of
Microbiology, v. 34, p. 263-283.
69
-------
Luong, H.V., J.M. Forshaug and EJ. Brown, 1981, Biogeochemistry of arsenic mine drainage,
University of Alaska, Fairbanks, Institute of Water Resources Completion Report
#79-21, OWRT Project B-045-ALAS).
Lynch, D.C., 1988, A review of the physical chemistry of arsenic as it pertains to primary
metals production, in Arsenic Metallurgy Fundamentals and Applications, R.G. Reddy,
J.L. Hendrix, and P.B. Queneau, eds., AIME Proceedings, p. 3-33.
Mann,' H., W.S. Fyfe and R.W. Kerrich, 1990, Metal accumulation and Fe, Ti-oxide
biomineralization by acidophilic microorganisms in mine waste environments, in Acid
Mine Drainage: Designing for Closure, Papers presented at the GAC/MAC Joint
Annual Meeting, Vancouver, B.C., May 16-18, 1990, p. 63-80.
Masscheleyn, P., R. Delaune and W. Patrick, 1991, Effect of redox potential and pH on
arsenic speciation and solubility in a contaminated soil, Environmental Science and
Technology, v.25, p. 1414-1419.
McBride, B.C. and R.S. Wolfe, 1971, Biosynthesis of dimethylarsine by
methanobacterium, Biochemistry, v.10, p. 4312-4317.
McCandless, R., 1992, Environment Canada, oral communication.
McLoughlin, R.P., J. Grounds and J. Weatherhead, 1988, Nabarlek Uranium Mine Design,
Construction, Operation Monitoring and Decommissioning of the Water
Management System, in Proceedings of the Third International Mine Water
Congress, October, 1988, Melbourne, Victoria, Australia, p. 775-784.
Miller, S.D. and G.S. Murray, 1988a, Prediction of time dependent factors in acid mine
drainage, in Proceedings of the Third International Mine Water Congress, October
23-28, 1988, Melbourne, Australia, p. 165-172.
Miller, S.D. and G.S. Murray, 1988b, Application of acid-base analysis to wastes from
base metal and precious metal mines, in Mine Drainage and Surface Mine
Reclamation, V. I: Mine Water and Mine Waste, U.S. Bureau of Mines Information
Circular 9183, p. 29-32.
Mirza, A.H., D. Tahija, K. Chen and H.H. Huang, 1988, Formation and stability studies
of iron-arsenic and copper-arsenic compounds from copper electrorefining sludge, in
Arsenic Metallurgy Fundamentals and Applications, R.G. Reddy, J.L. Hendrix, and
P.B. Queneau, eds., AIME Proceedings, p. 37-58.
Mok, W.M. and C.M. Wai, 1989, Distribution and mobilization of arsenic species in the
creeks around the Blackbird mining district, Idaho, Water Research, v.23, p. 7-13.
70
-------
Moore, J.N., W.H. Ficklin and C. Johns, 1988, Partitioning of arsenic and metals in
reducing sulfidic sediments, Environmental Science and Technology, v.22, p.
432-437.
Morin, K.A., 1988, Groundwater contamination from precious-metal, base-metal,
uranium, phosphate, and potash (KCL) mining operations, in Lin, C.L., ed.,
International groundwater symposium on hydrogeology of cold and temperate climates
and hydrogeology of mineralized zones, Nova Scotia Dept. Environment, Water
Resources and Planning, Halifax, Nova Scotia, p. 165-174.
Morin, K.A., 1990, Acid drainage from mine walls: The Main Zone Pit at Equity Silver
Mines, 109 p.
Morth, A.H., E.E. Smith, and K.S. Shumate, 1972, Pyrite systems: A mathematical model,
Contract report for the U.S. Environmental Protection Agency, EPA-R2-72-002.
Mudroch, A. and T.A. Clair, 1986, Transport of arsenic and mercury from gold mining
activities through an aquatic system, The Science of the Total Environment, v.57, p.
205-216.
NBMG (Nevada Bureau of Mines and Geology), 1989, The Nevada Mineral Industry 1988,
Nevada Bureau of Mines and Geology Special Publication MI-1988.
NBMG (Nevada Bureau of Mines and Geology), 1991, The Nevada Mineral Industry 1990,
Nevada Bureau of Mines and Geology Special Publication MI-1990.
NDEP (Nevada Division of Environmental Protection), 1991, files.
Nordstrom, O.K., 1982, Aqueous pyrite oxidation and the consequent formation of
secondary iron minerals, in J. A. Kittrick, D.S. Fanning and L.R. Hossner, eds., Acid
Sulfate Weathering, Soil Science Society of America Special Publication #10, p.
37-56.
Nordstrom, D.K. and C.N. Alpers, 1990, U.S. Geological Survey, written communication.
Osbom, T.R. and P.H. LeBlond, 1974, Static stability in freshwater lakes, Limnology and
Oceanography, v. 19, p. 544-545.
Parsons, J.D., 1977, Effects of acid mine wastes on aquatic ecosystems, Water, Air and
Soil Pollution, v. 7, no. 3, p. 333-354.
Patterson, R. J., 1990, Mine closure planning at Equity Silver Mines Ltd., in Acid Mine
Drainage: Designing for Closure, Proceedings of the GAC/MAC Joint Annual
Meeting, Vancouver, B.C., May 16-18, p. 441-449.
71
-------
Patton, F.D. and D.U. Deere, 1971, Geologic factors controlling slope stability in open
pit mines, in C.O. Brawner and V. Milligan, eds., Stability in open pit mining,
Proceedings of the First International Conference on Stability in Open Pit Mining,
Vancouver, B.C., November 23-25, 1970, p. 23-47.
Pierce, M.L. and C.B. Moore, 1982, Adsorption of arsenite and arsenate on amorphous
iron hydroxide, Water Research, v.16, p. 1247-1253.
Piteau, D.R. and D.C. Martin, 1982, Mechanics of rock slope failure, in C.O. Brawner,
ed. Proceedings of the Third International Conference on Stability in Surface Mining,
Vancouver, B.C., June 1-3, 1981, p. 113-169.
Pratt, J.R., J. Cairns, P.M. Stewart, N.B. Pratt and B.R. Niederlehner, 1985, Measurement
of recovery in lakes following phosphate mining, Final Report, Virginia Polytechnic
Institute and State University, Center for Environmental Studies, Publication No. 03-
045-039, 117 p.
Reith, C.C., R. Portillo, J. Millard, and D. Gonzoles, 1990, Cost optimization on the Jackpile-
Paguate reclamation program, in P.M. Doyle, ed., Mining and Mineral Processing
Wastes, Proceedings of the Western Regional Symposium on Mining and Mineral
Processing Wastes, Berkeley, California, May 30-June 1, 1990, p. 181-184.
Robins, R.G., J.C.Y. Huang, T. Nishimura and G.H. Khoe, 1988, The adsorption of arsenate
ion by ferric hydroxide, in Arsenic Metallurgy Fundamentals and Applications, .R.G.
Reddy, J.L. Hendrix, and P.B. Queneau, eds., AIME Proceedings, p. 99-112.
Schubert, J. P., 1980, Groundwater-surface water interchange in surface mined lands, Eos, v.
61, no. 48, p. 1193 (abstract only).
Seyler, P. and J. Martin, 1989, Biogeochemical processes affecting arsenic species
distribution in a permanently stratified lake, Environmental Science and Technology,
v.23, p. 1258-1263.
Steffen Robertson and Kirsten (B.C) Inc., 1989, Draft Acid Rock Drainage Technical Guide,
v. 1, prepared for the British Columbia Acid Mine Drainage Task Force.
72
-------
Stollenwerk, K.G. and J.H. Eychaner, 1988, Solubility of aluminum and iron in ground
watei; near Globe, Arizona, in Mallard, G.E. and Ragone, S.E., eds., U.S. Geological
Survey Toxic Substances Hydrology Program - Proceedings of the technical meeting,
Phoenix, Arizona, Sept. 26-30, 1988: U.S. Geological Survey Water-Resources
Investigations Report 88-4220, p. 581-591.
Tessier, A., F. Rapin and R. Carignan, 1985, Trace metals in oxic lake sediments: Possible
adsorption onto iron hydroxides, Geochimica et Cosmochimica Acta, v.49, p. 183-
194.
U.C. (University of California) Berkeley Mining Waste Study Team, 1988, Mining Waste
Study, Final Report, 416 p.
Voelker, D.C., 1985, A gazetteer of surface mine lakes, Eastern Interior Coal Province,
Illinois, U.S. Geological Survey Water Resources Investigations 84-4355, 81 p.-
Wakao, N., H. Koyatsu, Y. Komai, H. Shimokawara, Y. Sakurai and H. Shiota, 1988,
Microbial oxidation of arsenite and occurrence of arsenite-oxidizing bacteria in acid
mine water from a sulfur-pyrite mine, Geomicrobiology Journal, v.6, p. 11-24.
Welch, A.H., M.S. Lico and J.L. Hughes, 1988, Arsenic in groundwater of the western
United States, Groundwater, v.26, p. 333-347.
Wetzel, R., 1983, Limnology, Saunders College Publishing, Philadelphia, 743 p.
Wicks, C.M., J.S. Herman and A.L. Mills, 1991 Early diagenesis of sulfur in the
sediments of lakes that receive acid mine drainage, Applied Geochemistry, v. 6, p.
213-224.
Williams, D.I. and R.P. Von Herzen, 1983, On the terrestrial heatflow and physical
limnology of Crater Lake, Oregon, Journal of Geophysical Research, v. 88, no. B2,
p. 1094-1104.
Wilson, D., B. Finlayson and N. Morgon, 1981, Copper, zinc and cadmium concentrations
of resident trout related to acid mine wastes, California Fish and Game, v. 67(3), p.
176-186.
Wojciechowski, J. and K. Serewko, 1985, Land subsidence affected by dewatering of open-
pits, in Mine Water, Proceedings of the Second International Congress, v. 2,
September, 1985, Granada, Spain, p. 875-889.
Woodward-Clyde Consultants, 1992, Screencheck Environmental Assessment, Robinson
Project, prepared for Magma Copper Company, Tucson, Arizona, (Draft).
73
-------
Woolfolk, C.A. and H.R. Whiteley, 1962, Reduction of inorganic compounds with
molecular hydrogen by Micrococcus lactilvticus, part I,' Journal of Bacteriology,
v.84, p. 647-658.
Ziemkiewicz, P.P., A.H. Stiller, T.E. Rymer and W.M. Hart, 1990, Advances in the prediction
and control of acid mine drainage, in Mining and Mineral Processing Wastes, P.M.
Doyle, ed., Proceedings of the Western Regional Symposium on Mining and Mineral
Processing Wastes, Berkeley, California, May 30-June 1, 1990, p. 93-101.
74
U.S. Environmental Protection Agency
Region 5, Library (PL-12J)
77 West Jackson Boulevard, 12th Floor
Chicago, IL 60604-3590
------- |