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

                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

                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.

                                   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

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

15     1985-89 aqueous pH, Southern Tail Pit, Equity Silver Mines, B.C	            51
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

       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
           100 km
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
Figure  2
                              10        15       20
                                   TEMPERATURE (°C)
25       30
                          Typical thermal stratification  of a  lake.
                                                      After Wetzel, 1983









48 12


••"** J


10 QgZO 30










4 8


10 ^ 20



0 4 *" 8 12



10 0^20 30

— - ^
— «2
-^~~—"— Temp.
SPRING TURNOVER After Het«u 1983

0 4 8 12

1 to




O 1O _ 20 30

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
       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
                                                                       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


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,

       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).

       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
       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
At pH values  above 2.3 to 3.5, Fe(ffl) will precipitate as Fe(OH)3:
                              Fe3* + 3H20 --> Fe(OH)3 + 3H+


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).


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).


       The acid produced by sulfide oxidation may be neutralized if the wallrocks contain
sufficient amounts of carbonate minerals:
                             CaCOj + IT  --> Ca2++ HCCV
                            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).

Table 1.  Iron Sulfate Minerals Identified at Iron Mountain, CA
Additional Sulfates Identified
   '•  FenS04*7H20
      K2Fe5nFe4m(SO4)12* 18H2O

                                            Nordstrom and Alpers, written
                                            communication, 1990

       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

             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


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


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).

       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.

       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


 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.

10      11
Eh-pH diagram showing contours of dissolved iron
                                          After Drever, 1988
Figure  5

       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

Arsenic Speciation
       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,

       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).



                  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

       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

OL oi

3 |»

<-»• O>


                                            As SOLUBILITY  mg/l

                             8  *-*       u.  b       o,  S       S8
                                                                          Ol  O
                                                                             §   i

                       Eh-pH diagram for the system  Fe-As-H-O
                         snowing arsenic compounds at 25<>C.
                                                         After Mirza et al.,  1988
Figure  8

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


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.


       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


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:

                     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.

       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.

                                   Existing Pit Lakes
       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.

       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

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

       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.

       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

       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,


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"


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

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


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

0 80



1 1

1 1 1 1
2 3 4 5 6 .7 8

OaU from Davis and Ashenberg, 1989
Applied Geochemistry, v. 4.
Figure 9

                              BERKELEY PIT,  BUTTE, MT
                    -  60
Figure  10
                             Dissolved Oxygen (mg/1)
                                             after Davis and Ashenberg
                                             1989, Applied Geochem., v.  4.

                                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

                                  BERKELEY PIT,  BUTTE, MT
                                250500        750       1000
                                   Dissolved As (ug/1)
                                                   after Davis and Ashenberg
                                                   1989, Applied Geochem., v.  4.
Figure  12

                                BERKELEY PIT,  BUTTE, MT      «
                                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

                               BERKELEY PIT, BUTTE, MT
       120 .
                   I    I   I    I   J    I    i    I    I    I    I    I    I
   .1                  .15

Figure 14
                                                      Data from Davis and Ashenberg, 1989
                                                      Applied Geochemistry, v. 4.

         TABLE 2
Pit Water Quality at Ruth


   Woodward-Clyde Consultants,  1992

                       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

                        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

       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



7 -

5. 6"

| _
4 "

3 -
2 —

._ 1

• "O
. to

• ,
I m»

.• •*

t * '

, t
' •?
» • ••




1985 1986 1987 1988 1989
                                             After Morin, 1990
                                             Add Drainage from Mine Walls
                                             The Main Zone Pit at Equity
                                             Silver Mines
Figure 15

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).

                                                    LOG (dissolved Zn  in Hg/1)
                       «»  ••
                                                                                                           LOG (dissolved Cu in Mg/1)
                                                                                                  i    i
                                                                                                                   «— '    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

                           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


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).


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,

       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


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.

       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

       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


       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.

       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


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

       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.

                                  ttuMIIQ O1"*11 "••*   Urrirt
          • Igneous  hosted

          O sedlmenUry hosted
                                                                             Some locations
                                                                             are approximate
Figure  17

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