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Proc Natl Acad Sci U S A. 1999 March 30; 96(7): 3455–3462. | PMCID: PMC34288 |
Copyright © 1999, The National Academy of Sciences Negative pH, efflorescent mineralogy, and consequences for
environmental restoration at the Iron Mountain Superfund
site, California D. Kirk Nordstrom*† and Charles N. Alpers‡ *United States Geological Survey, 3215 Marine Street, Boulder, CO 80303-1066; and ‡United States Geological Survey, Placer Hall, 6000 J Street, Sacramento, CA 95819-6129 The Richmond Mine of the Iron Mountain copper deposit contains some
of the most acid mine waters ever reported. Values of pH have been
measured as low as −3.6, combined metal concentrations as high as 200
g/liter, and sulfate concentrations as high as 760
g/liter. Copious quantities of soluble metal sulfate salts such
as melanterite, chalcanthite, coquimbite, rhomboclase, voltaite,
copiapite, and halotrichite have been identified, and some of these are
forming from negative-pH mine waters. Geochemical calculations show
that, under a mine-plugging remediation scenario, these salts would
dissolve and the resultant 600,000-m3 mine pool would have
a pH of 1 or less and contain several grams of dissolved metals per
liter, much like the current portal effluent water. In the absence of
plugging or other at-source control, current weathering rates indicate
that the portal effluent will continue for approximately 3,000 years.
Other remedial actions have greatly reduced metal loads into downstream
drainages and the Sacramento River, primarily by capturing the major
acidic discharges and routing them to a lime neutralization plant.
Incorporation of geochemical modeling and mineralogical expertise into
the decision-making process for remediation can save time, save money,
and reduce the likelihood of deleterious consequences. Mining of metallic sulfide ore deposits (primarily for Ag, Au, Cu,
Pb, and Zn) produces acid mine waters with high concentrations of
metals that have harmful consequences for aquatic life and the
environment. Deaths of fish, rodents, livestock, and crops have
resulted from mining activities and have been noted since the days of
the Greek and Roman civilizations. Mining and mineral processing have
always created health risks for miners and other workers. In addition,
mining wastes have often threatened the health of nearby residents by
exposure to emissions of sulfur dioxide and oxides of As, Cd, Pb, and
Zn from smelter stacks and flues, metal-contaminated soils, and waters
and aquatic life with high concentrations of metals. As with most forms
of resource extraction, human health risks accompany mineral
exploitation. In 1985, the U.S. Environmental Protection Agency (EPA) estimated that
50 billion tons (45 × 10 12 kg; 1 ton =
907 kg) of mining and mineral processing wastes had been generated in
the United States and about 1 billion tons would continue to be
generated each year ( 1). More recently, the EPA has described 66
“damage cases” at their web site ( www.epa.gov, search for
Mining and Mineral Processing Wastes, accessed Sept. 9,
1998) in which environmental injuries from mining activities in the
U.S. are detailed. Government records indicate that many millions,
perhaps billions, of fish have been killed from mining activities in
the U.S. during this century ( 2). Incidents of arsenic poisoning in
residents of Thailand result from arsenic contamination of the shallow
groundwaters because of weathering of mine wastes ( 3). A mine flood
disaster in Spain occurred in April 1998 in which about 6 million
m 3 of acid water and sulfide tailings escaped
from a breached impoundment and covered about 6,500 acres of farmland
and river banks along a 70-km reach of the Guadiamar River with
fine-grained sulfides (details at www.csic.es). Numerous rivers,
estuaries, and reservoirs throughout the world have been used as
dumping grounds for the large volumes of waste produced during mineral
extraction and processing. Mineral processing, in addition to fossil
fuel and metal utilization, has increased the concentration of selected
metals and nonmetals in the atmosphere. The emissions of As, Cd, Cu,
Pb, Sb, and Zn from anthropogenic sources are all greater than
emissions from natural sources, sometimes several times higher ( 4, 5). Acid mine drainage is produced
primarily by the oxidation of the common iron disulfide mineral pyrite.
Pyrite oxidation is a complex process that proceeds rapidly when this
mineral and other sulfides are exposed to air. A simplified
representation of this chemical process is given by the reaction of
pyrite with air and water,
in which the product is a solution of ferrous sulfate and sulfuric
acid. The dissolved ferrous iron continues to oxidize and hydrolyze
when the mine water is no longer in contact with pyrite surfaces,
producing additional acidity. Iron- and sulfur-oxidizing bacteria,
especially Thiobacillus ferrooxidans, are known to catalyze
these reactions at low pH, increasing reaction rates by several orders
of magnitude ( 6). These processes occur naturally and, indeed, natural
acidic drainage is well known from many locations ( 7). Mining has the
overall effect of dramatically increasing the oxidation rates by
providing greater accessibility of air through mine workings, waste
rock, and tailings, by creating greater surface area exposure through
blasting, grinding, and crushing, and by concentrating sulfides in
tailings. The overall rates of sulfide oxidation and metal release in
areas affected by mining are estimated to be orders of magnitude faster
than natural rates. Another process, sometimes overlooked, plays an important role in the
environmental consequences of mining: the formation of soluble,
efflorescent salts. Acid ferrous sulfate solutions often become so
enriched through rapid pyrite oxidation and evaporation that soluble
salts form. These often appear as white, blue-green, yellow to orange
or red efflorescent coatings on surfaces of waste rock, tailings, and
in underground or open-pit mines. Acidity and metals, formerly
contained in the acid mine water, are stored in the salts, which can
quickly be dissolved by a rising groundwater table or be dissolved when
exposed to rain and flowing surface waters, and then infiltrate to
groundwaters. The Iron Mountain Mine Superfund site is an extreme
example of how the formation of soluble efflorescent minerals can make
certain remediation alternatives much more risky and potentially
disastrous than might otherwise be imagined. Iron Mountain is located in Shasta County, California,
approximately 14 km northwest of the town of Redding (Fig.
), in the southern part of the
Klamath Mountains. “Iron Mountain Mine” is really a group of
mines within Iron Mountain that include Old Mine, No. 8 Mine,
Confidence-Complex, Brick Flat Open Pit Mine, Mattie Mine, Richmond and
Richmond Extension Mines, and Hornet Mine. Ag, Au, Cu, Fe, Zn, and
pyrite (for sulfuric acid production) were recovered at various times
beginning in the early 1860s and ended with the termination of open-pit
mining in 1962. Iron Mountain was once the largest producer of Cu in
the state of California, and now it produces some of the most acidic
waters in the world. Prior to the late 1980s when major remediation
efforts began, approximately 2,500 tons of pyrite weathered every year
from one mine alone (the Richmond Mine) and water containing about 300
tons per year of dissolved Cd, Cu, and Zn drained from the site into
the Sacramento River. During periods of high runoff, sudden surges of
acid mine waters into the Sacramento River have caused massive fish
kills, which state and federal agencies have investigated since 1939.
More than 20 fish-kill events have occurred in Sacramento River
receiving waters since 1963, with at least 47,000 trout killed during a
single week in 1967 ( 8). Furthermore, the town of Redding (with
approximately 100,000 residents) receives its drinking water from the
Sacramento River, downstream from the Iron Mountain site. Large,
metal-rich sediment deposits containing toxic porewaters have built up
in Keswick Reservoir, where the Spring Creek drainage from Iron
Mountain empties into the Sacramento River. A brief history of mining,
water management, and environmental action at Iron Mountain is outlined
in Table 1.
| Table 1Brief chronology of Iron Mountain mining and
envrionmental activities |
The mineral deposits are primarily massive sulfide lenses as much as
60 m thick containing up to 95% pyrite, variable amounts of
chalcopyrite and sphalerite, and averaging about 1% Cu and about 2%
Zn. Some disseminated sulfides occur along the south side of the
mountain. The deposits at Iron Mountain and elsewhere in the West
Shasta mining district are Devonian in age and have been classified as
Kuroko type, having been formed in an island arc setting in a marine
environment ( 9). The country rock is the Balaklala Rhyolite, a
keratophyric rhyolite that has undergone regional metamorphism during
episodes of accretion of oceanic crust to the continent. The brittle,
fractured nature of the altered volcanic bedrock gives rise to a
hydrologic conditions dominated by fracture-flow at Iron Mountain. The
mineral composition of the rhyolite is albite, sericite, quartz,
kaolinite, epidote, chlorite, and minor calcite; consequently it has
little buffering capacity. Kinkel and others ( 10), Reed ( 11), and South
and Taylor ( 12) have documented the chemical and isotopic compositions
of ore, gangue, and country-rock minerals in the West Shasta mining
district. Weathering of massive sulfide deposits at and near the
surface has given rise to large gossan outcrops, enriched in Ag and Au.
The 10 million tons of gossan in place prior to mining is the residue
from at least 15 million tons of massive sulfide that weathered
naturally. A total of 7.5 million tons of sulfide ore was mined at Iron
Mountain, and remaining reserves are estimated at approximately 15
million tons ( 13), so the overall size was at least 37.5 million tons
prior to weathering. Preliminary paleomagnetic data on iron oxides in
the gossan show portions with reversed polarity, indicating the gossan
began forming at least 780,000 years ago. Secondary enrichment in the
upper zones of the massive sulfides resulted in high concentrations of
Cu (5–10%) and Ag (about 1 oz/ton). This enrichment took place at
or near the water table during gossan formation. Three main massive sulfide ore bodies, the Brick Flat, the Richmond,
and the Hornet, include most of the oxidizing sulfides causing the
current water-quality problems. These ore bodies are thought to be
parts of a single massive sulfide body about 0.8 km long, over 60
m wide, and over 60 m thick that was offset by two normal faults
(Fig. ). All three of these
bodies have been mined, and the consequences include large changes in
the hydrogeology, resulting in highly contaminated waters from tunnels,
tailings, and waste rock piles.
Acid Effluent from the Richmond Mine Conditions at Iron Mountain are nearly optimal for the production
of acid mine waters, and this mine drainage is some of the most acidic
and metal-rich reported anywhere in the world ( 14, 15). In the Richmond
Mine, about 8 million tons of massive sulfide remain ( 13). At current
weathering rates it would take about 3,200 years for the pyrite in the
Richmond ore body to fully oxidize. The massive sulfide deposit is
about 95% pyrite and is excavated by tunnels, shafts, raises, and
stopes which allow rapid transport of oxygen by air advection. The
sulfides are at or above the water table so that moisture and oxygen
have ready access. Airflow is driven by the high heat output from
pyrite oxidation. About 1,500 kJ of heat is released per mole (120
g) of pyrite. Air enters the main tunnel, heats up in the mine, then
travels up through raises and shafts to the surface. The average flux
of acid mine drainage from the Richmond portal indicates that about
2,400 mol of pyrite is oxidized every hour, producing about 1 kW of
power or almost 9,000 kW per year. Water temperatures as high as 47°C
have been measured underground, and the amorphous silica geothermometer
( 16, 17) would suggest temperatures of at least 50°C in the
subsurface. In the early days of mining at Iron Mountain, fires were
frequent during underground excavation, and temperatures of
430 oF (221°C) were recorded at the ore surface
( 18). A considerable amount of historical data exist for effluent composition
and discharge from the Richmond Mine because it is the largest single
source of dissolved metals (both in terms of concentration and in terms
of flux) in the Iron Mountain district. The Richmond ore body was
discovered about 1915 but it was not mined on a large scale until the
late 1930s and the war years (1940–1945). Regular monitoring of the
Richmond Mine effluent by the California Regional Water Quality Control
Board in cooperation with the EPA began in 1983. A summary of the data
for discharge, pH, and Cu and Zn concentrations for 1983–1991 is shown
in Table 2. Further compilation
and details of Richmond portal effluent composition and discharge can
be found in Alpers et al. ( 19).
| Table 2Richmond Mine portal effluent
characteristics,1983–1991 |
The variability in the Richmond effluent with time can be seen quite
clearly for the 1986–1987 monitoring period. Fig.
covers the time period of late
November 1986 to April 1987 and shows the rainfall (at Shasta Dam), and
the consequent changes in Richmond Mine discharge and copper and zinc
concentrations. One explanation for the large increase in copper
concentrations is the dissolution of underground soluble salts from the
flushing effect of meteoric recharge (see below). An observed increase
in temperature with increased discharge may be the result of the
dilution of concentrated sulfuric acid, the dissolution of soluble
salts, and increased pyrite oxidation.
One of the obvious options for remediation of the Richmond Mine was to
plug it. Many mines have been plugged, but the consequences have not
been consistently favorable. The EPA wanted to know what the
consequences of plugging the Richmond Mine might be; for example, what
would the composition of the resultant mine pool be? There was,
however, no basis on which to speculate without some idea of the
underground conditions. Hence, one of the activities of the Second
Remedial Investigation Phase (1986–1992) under the Superfund Program
was an underground survey of the Richmond tunnel and part of the mine
workings. Prior to underground renovations in 1989–1990, the last
underground tour, to the best of our knowledge, was in 1955 (Don White,
U.S. Geological Survey, personal communication, 1989). The last mining
had occurred in the late 1940s. Other than an occasional inspection by
a company employee, there had been no recorded observation of the
underground workings for 35–40 years. After underground renovations,
entry was safe, and on September 10–12, 1990, water and mineral
samples were collected. They revealed extremely acidic seeps with pH
values as low as −3.6 and total dissolved solids concentration of more
than 900 g/liter. The chemical compositions of five of the most acidic waters found
underground in the Richmond Mine during 1990–1991 are shown in Table
3. These concentrations are the
highest ever recorded for As, Cd, Fe, and SO 4 and
nearly the highest for Cu and Zn in groundwater. The high subsurface
temperatures have induced considerable evaporation, which, in addition
to pyrite oxidation, has caused the high concentrations of dissolved
metals and sulfate.
| Table 3Compositions of five extremely acid mine water samples
from the RichmondMine |
The reporting of negative pH values has been controversial, and for
several good reasons. The conventional definition of pH based on the
former National Bureau of Standards criteria and defined buffer systems
limits the range of definable and measurable pH values to that of 1 to
13. Outside this range, the concept and measurement of pH are difficult
at best. Furthermore, a new definition of pH must be used that is
consistent with the conventional definition, different buffers must be
used, and electrode performance and interferences must be determined.
The most acceptable model for activity coefficients at present for
defining pH below 1.0 is the Pitzer ion-interaction approach ( 20, 21).
Acid mine waters are solutions of sulfuric acid, so the Pitzer model
applied to sulfuric acid ( 22, 23) could serve as a definition for pH.
Standardized sulfuric acid solutions would then serve as buffer
solutions for calibration and the remaining question is the performance
of standard glass membrane electrodes under these extreme conditions.
Several Orion Ross glass membrane electrodes and a Sargent–Welch glass
membrane electrode all performed well and could be calibrated up to a
sulfuric acid concentration of about 8 molal. Another difficulty facing
the definition of pH below 0.0 is scaling of individual ion activity
coefficients. There is no generally accepted procedure for defining
individual ion activity coefficients without some arbitrary
assumptions. Two common methods with the Pitzer approach include
“unscaled” Pitzer equations, and “MacInnes scaled,” using
the MacInnes assumption ( 24). The MacInnes assumption is simpler, more
flexible for a wide range of complex chemical compositions, and is more
consistent with conventional speciation models applied to natural
waters ( 24). It could be argued that the MacInnes assumption becomes
less defensible at high concentrations where the unscaled approach
should be more appropriate, but there is no obvious justification for
using one approach over the other and the choice remains arbitrary. In
the present investigations, the MacInnes scaling was used primarily
because geochemists who have applied the Pitzer method to the
interpretation of brines and saline waters find the MacInnes assumption
more consistent with conventional practice. If the unscaled approach is
used, the resultant pH values begin to differ significantly from
MacInnes scaling for sulfuric acid solutions with pH values below
−0.5. For example, at a sulfuric acid concentration of about 5.0 molal
a scaled pH would be −2, whereas the unscaled pH would be notably
higher, about −1.2. Some of these negative-pH mine waters were in apparent equilibrium with
prominent soluble salts. For example, a stalactite of zincian-cuprian
melanterite had water dripping from the tip that had a pH of −0.7
(Table 3 and Fig. ). Enormous
quantities of highly soluble iron sulfate salts were found as
efflorescences and precipitates, coating walls, ceilings, and floors of
the mine and growing out of muck piles in colorful assemblages.
Identification of these soluble salts made it possible to estimate what
the composition of a mine pool formed by mine plugging might be.
Soluble Salts and Consequences of the Mine-Plugging Scenario Ten soluble iron sulfate salts plus gypsum and chalcanthite were
identified in the Richmond Mine. These minerals and their idealized
formulae are listed in Table 4,
with the iron salts in approximate sequence downward from the early
formed to the later formed. Rhomboclase was found as stalactites and
stalagmites (Fig. ), and
clusters of coquimbite, römerite, copiapite, and voltaite
crystals were common throughout the mine (Fig.
). Rhomboclase was rarely found
without voltaite crystals.
| Table 4Idealized formulae of sulfate minerals found in the
RichmondMine |
As long as an acid mine water is in contact with pyrite, the dissolved
iron will remain in the ferrous state because of the strong reducing
capacity of the pyrite. Rapidly flowing mine water will still maintain
a high proportion of ferrous iron because the oxidation rate is often
slow enough relative to the flow rate of the water. Consistent with
this expectation, the only iron sulfate salts containing exclusively
ferrous iron, melanterite, rozenite, and szomolnokite, are found close
to pyrite sources and associated with more rapidly flowing waters.
Ferric-bearing minerals are found to form in more stagnant conditions
and can be considered to be hydrologic “dead-ends,” where much of
the Fe II has had time to oxidize to
Fe III. Additional evidence for this mineralogical
evolution is the observation that melanterite is the first-formed
mineral when typical acid mine water is allowed to evaporate under
ambient conditions and rhomboclase and voltaite are the last formed
( 25). A copper–zinc partitioning study of melanterite demonstrates that
melanterite prefers copper over zinc ( 15). The consequences of this
partitioning are that portal effluents will tend to have higher ratios
of Zn/Cu during the dry season when melanterite is forming
underground and lower Zn/Cu ratios in the wet season when these salts
are dissolved and flushed from the mine workings. This trend is seen in
the historical data on the Richmond Mine effluent ( 15). Dissolution of these soluble, iron sulfate salts (with variable amounts
of copper, zinc, cadmium, and aluminum substituting for the iron) can
generate acidic solutions with high concentrations of dissolved metals.
During the rising limb of a stream discharge in central Virginia after
the onset of rain, Dagenhart ( 26) showed that rapid increases in the
concentrations of Cu, Zn, Fe, and Al resulted from the dissolution of
efflorescent salts found on upstream tailings and waste rock piles.
This phenomenon must be common at mine waste sites and is likely to be
an important cause of fish kills associated with periods of high
runoff, especially after prolonged dry periods. Now we consider the
consequences of dissolution of the enormous quantity of salts in the
Richmond Mine in a mine-plugging scenario. The chemical composition of the mine pool created by plugging the
Richmond Mine can be estimated by allowing these salts to dissolve in a
volume of water equivalent to the void space created by the underground
workings. The exact proportion of the different type of salts is not
known, but the results of the calculations are not particularly
sensitive to this factor. The amount of salts stored underground is a
more critical factor, and so that was considered a variable.
Computations were made by inputting the mineral compositions to the
phreeqe program (ref. 27, now superseded by
phreeqc, ref. 28) for a range of salt volumes.
phreeqe can calculate the speciation and chemical
equilibrium for mass transfer processes such as precipitation,
dissolution, oxidation–reduction reactions, ion exchange, and gas
addition or removal ( 29). The results are shown in Fig.
, where the resultant pH in the
mine pool is plotted against the volume of added salts under two
scenarios: active infiltration (actively injecting clean water) and
passive infiltration (letting the groundwater naturally fill the void
spaces). The latter scenario gives a worse picture because passive
infiltration would allow more pyrite oxidation and the buildup of more
acid waters. The salts probably occupy about 1% of the volume of the
mine workings based on visual inspection from the limited subsurface
survey. As can be seen in Fig. , however, an error in this value makes
little difference. The consequences are that a mine pool of about
600,000 m 3 with a pH at or below 1, with many
grams of dissolved metals per liter (much like the current portal
effluent), would likely form at or near the top of the groundwater
table, in a rock with almost no neutralization capacity, and in which
the hydrologic flow is governed by fractures, excavations, and
drill-holes. Thus, plugging presents a remediation scenario that has a
high degree of risk with potentially dangerous results.
It has been common engineering practice to plug abandoned or inactive
mines without monitoring, modeling, or even considering the physical
and chemical consequences. Major leaks or failures at plugs, widespread
and disseminated seeps of enriched acid mine waters, and increases in
subsurface head pressures of more than 100 m have occurred. For
some mine sites, plugging may ultimately prove to be successful, but
more careful planning and peer review are essential to lessen the
probability of disastrous results. Regulatory Investigations and Remediation Several investigations and regulatory actions at Iron Mountain
have been initiated by California State agencies over the last few
decades. These are too lengthy to summarize here. Since the original
listing of Iron Mountain on the National Priorities List in 1983, the
EPA has authorized four Records of Decision (RODs) and has considered
numerous options for remediation. A condensed version of the main
remedial alternatives is as follows:
- No action
- Surface-water diversion
- Lime neutralization
- Capping (partial or complete capping of the mountain to prevent
infiltration)
- Enlargement of Spring Creek Debris Dam (acid water storage and release
structure)
- Ground-water interception
- Air sealing
- Mine plugging
- On-site leaching and solution extraction
- Continued mining under environmentally safe conditions
- Combined alternatives
Surface-water diversions have been installed to divert clean
headwater streams around contaminated areas. The waters that are the
largest sources of metal loadings have been captured and diverted to a
lime neutralization plant. In the late 1980s, an emergency lime
neutralization plant with a capacity of about 60 gallons per minute
(gpm; 1 gallon = 3.8 × 10−3
m3) was installed to handle the worst flows from
the Richmond and Hornet portals. By December of 1992, this plant had
been expanded to handle 140 gpm, but was operated only 4 months per
year during highest flows. In July of 1994 a new plant with a
capacity of about 1,400 gpm began operation at the Minnesota Flats
tailings site. In 1996 it was upgraded to 2,000 gpm, and high-density
sludge treatment was added. Now it accepts drainage from Slickrock
Creek (pumped from Old Mine and No. 8 Mine workings) as well as the
Richmond and Hornet Mine portal effluents. The decision to build the larger treatment plant and to treat
discharges from the Lawson tunnel (Hornet Mine) was also influenced by
geochemical modeling. Opinion was divided as to whether the flow of
acid mine water from the Lawson tunnel originates from the Richmond
Mine by spillage or leakage or whether the Hornet ore body produces its
own contaminant effluent. An ore chute and a raise that connected the
two mines were identified from the old mine maps (Fig. ). Because of
its proximity to the surface and the collapsed nature of the mine
workings, it was generally agreed that the Hornet Mine itself could not
be effectively plugged. However, consultants proposed that plugging the
Richmond Mine would stop or greatly reduce the flow from the Lawson
tunnel. There was also reason to believe that during the intervening
years since mining ceased, cave-ins and other ground failures had
largely cut off direct connections from the Richmond to the Hornet. Alpers and others ( 19) studied the historical data on
rainfall–discharge relationships between the two mines, Zn/Cu ratios
as a signature of reactions within each mine site, and mass balance
calculations for the two portal effluents. The most definitive method
of determining the possible influence of the Richmond Mine water on the
Lawson tunnel effluent was a mass balance approach. Using the known
water compositions discharging from each mine and knowing the
composition of the minerals that are reacting to form the effluent
waters, it is possible to calculate the mass amounts of minerals
dissolved or precipitated to produce these waters by using the
balance program ( 30). Mineral reaction signatures were
developed for each mine effluent separately, and then Richmond effluent
was mixed with clean ground water and allowed to precipitate and
dissolve additional minerals to determine if it was possible to derive
the Lawson effluent from the Richmond. No version of this mass balance
model produced a water that matched the Lawson effluent. Next, Richmond
effluent was also mixed with Lawson effluent, and geochemical reaction
was allowed, to see how much effluent each mine could be contributing
to the Lawson. The model results indicated that not more than about 2%
of the Richmond effluent could be present in the Lawson effluent.
Therefore, the Hornet Mine is producing its own effluent independently
of the Richmond Mine. Even if the Richmond Mine were successfully
plugged, water from the Hornet would continue to be a significant
problem and it would have to be treated. The fourth Record of Decision, issued in September of 1997, selected
the construction of a dam on Slickrock Creek. This structure will
capture the largest remaining loads of Cu and Zn and divert them to the
neutralization plant for treatment. The remaining remediation is now
focused on Boulder Creek, lower Spring Creek, Spring Creek Reservoir,
and the metal-enriched sediments that formed in Keswick Reservoir from
the neutralization of acid mine waters for nearly 50 years. The EPA and the potentially responsible parties remain in legal
contention over the appropriate final remediation approaches to be used
at Iron Mountain and the costs. Both the U.S. Government and the
potentially responsible parties have funded a considerable number of
investigations, remediation efforts, legal fees, and oversight
management. The loads of copper, zinc, and cadmium into the Sacramento
River have been reduced by 80–90%, and further remediation is in
progress or being planned. The main challenge that remains is how to
find a permanent (and passive) treatment solution in light of the fact
that the mine drainage will continue for approximately 3,000 years
unless the sulfide ore is mined out. Prevention and control of contamination at mine sites is a
challenging task, and remediation of large inactive mine sites such as
Iron Mountain has proven to be extraordinarily difficult, complex, and
expensive, not to mention litigious. The physical and chemical nature
of the site makes it difficult to assess the effectiveness of
remediation and the relative risks and costs of various alternatives
and their contingencies. There are no easy solutions to these types of
environmental problems, but several important points can be made about
cleanup of mine waste sites on the basis of our experiences at Iron
Mountain. First, there is tremendous value to having a technical advisory team of
multidisciplinary professionals, without an obvious conflict of
interest, to advise the regulatory agencies, to review data, and to
make recommendations. Mine sites and their contaminants are complex
functions of the geology, hydrology, geochemistry, pedology,
meteorology, microbiology, and mining and mineral processing history,
and their remediation is subject to considerations of economic
limitations, available technology, and potential land use. Furthermore,
the risks of failed remediation or no action are often poorly known.
Assessing such risks involves toxicology, epidemiology, wildlife
biology, and dealing with public perception. To ignore professionals in
these areas, who can contribute both to the wisest choice of
remediation strategies and to public awareness and education is to
invite mistakes. Second, the effectiveness of a remedial alternative usually cannot be
easily quantified or predicted. Hence, we must admit that remediation
is experimental. Research is required to effect the best and most
appropriate remediation available at a given time for a given site.
Both long-term and short-term remediations are needed. For the short
term, we need to fill in the knowledge gaps, especially as they pertain
to a particular site. For the long term, we need to continue to develop
better remediation techniques and mining and processing techniques that
can utilize mine wastes and mineral deposits of lower grade.
Mineralogical and geochemical knowledge make it possible to foresee the
potential consequences of a remedial option and to plan a remediation
strategy. The results of long-term research by the U.S. Geological
Survey provided technical tools (computer programs for geochemical
modeling and procedures for measuring pH) that could be used to answer
important questions regarding remediation scenarios. Third, it would seem prudent to proceed on mine waste cleanup in a
phased, iterative approach. Our natural inclination is to identify the
worst part of a hazardous waste site and attempt to clean it up. For
Iron Mountain, there is no single remedial solution that would clean up
90% of the problem on a permanent and maintenance-free basis (with the
exception of completely mining the mountain). There are, however,
several options (most of which have been exercised) that are low risk
and low cost and should reduce the discharge of acid mine waters. These
options can be instituted while deliberations and research continue to
find the long-term solution. Fourth, mine waste sites commonly contain low-grade resources that are
potentially mineable—it requires the right technology to make resource
recovery economic. In an age of increasing recycling, recycling
strategies should be applied to mine sites. Many mine wastes have
already undergone further metals extraction and others could be
stockpiled or tested for new uses. Additional research into metal
recovery from acidic solutions could also provide economic incentive to
recycling metals from mine drainage waste streams. Finally, Iron Mountain has been an extraordinary and extreme
environment in which to study and document the processes of acid mine
water production and efflorescent mineral formation, the value of which
goes far beyond just the immediate remediation needs. The processes and
properties found at Iron Mountain are probably commonplace at metal
sulfide mine and mineral processing sites, but usually on a smaller
scale. We now have some direct observations of the composition of water
that produces efflorescent minerals. We have some idea of the
consequences of efflorescent mineral dissolution when a mine is
plugged. We can estimate the geochemical consequences of various
remediation scenarios for mine sites with better confidence. Unraveling
the dynamic processes that affect water–mineral interactions is often
critical to solving hazardous waste problems in the hydrogeologic
environment. We are grateful to personnel of Region 9 of the U.S. EPA,
especially Rick Sugarek, for their continued support of our
investigations on this project and to personnel of CH2M
Hill for their help and assistance in our efforts to answer technically
challenging questions. We thank the California Regional Water Quality
Control Board in Redding and all the state agencies that have worked on
Iron Mountain for their cooperation and support. Roger Ashley and Katie
Walton-Day (U.S. Geological Survey) and James Hanley and Carol Russell
(EPA) provided helpful reviews. We also acknowledge Rick Sugarek for
making helpful suggestions on the manuscript. | | EPA | Environmental Protection Agency |
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