Remediation of Acid Mine Drainage-Impacted Water
Curr Pollution Rep
Remediation of Acid Mine Drainage-Impacted Water
Abhishek RoyChowdhury 0 1 2 4
Dibyendu Sarkar 0 1 2 4
Rupali Datta 0 1 2 4
0 Department of Earth and Environmental Studies, Montclair State University , Montclair, NJ 07043 , USA
1 Rupali Datta
2 Abhishek RoyChowdhury
3 Dibyendu Sarkar
4 Department of Biological Sciences, Michigan Technological University , Houghton, MI 49931 , USA
The formation of acid mine drainage (AMD), a highly acidic and metal-rich solution, is the biggest environmental concern associated with coal and mineral mining. Once produced, AMD can severely impact the surrounding ecosystem due to its acidity, metal toxicity, sedimentation and other deleterious properties. Hence, implementations of effective post-mining management practices are necessary to control AMD pollution. Due to the existence of a number of federal and state regulations, it is necessary for private and government agencies to come up with various AMD treatment and/or control technologies. This review describes some of the widely used AMD remediation technologies in terms of their general working principles, advantages and shortcomings. AMD treatment technologies can be divided into two major categories, namely prevention and remediation. Prevention techniques mainly focus on inhibiting AMD formation reactions by controlling the source. Remediation techniques focus on the treatment of already produced AMD before their discharge into water bodies. Remediation technologies can be further divided into two broad categories: active and passive. Due to high cost and intensive labor requirements for maintenance of active treatment technologies, passive treatments are widely used all over the world. Besides the conventional passive treatment technologies such as constructed wetlands, anaerobic sulfate-reducing bioreactors, anoxic limestone drains, open limestone channels, limestone leach beds and slag leach beds, this paper also describes emerging passive treatment technologies such as phytoremediation. More intensive research is needed to develop an efficient and cost-effective AMD treatment technology, which can sustain persistent and long-term AMD load. FeS2 þ 3:5O2 þ H2O→ Fe2þ þ 2SO24‐ þ 2Hþ
Acid mine drainage (AMD); Active AMD treatment; Passive AMD treatment; Phytoremediation
While coal and mineral mining is an important
revenuegenerating industry, several environmental consequences are
associated with it. The formation of a metal-rich acid solution
known as acid mine drainage (AMD) is a major
environmental problem associated with mining operations. Once exposed
to AMD, the quality of adjacent surface water degrades
drastically and eventually becomes unsuitable for sustaining
biodiversity. Additionally, soils exposed to AMD become
structurally unstable and highly prone to erosion [
AMD is produced due to the oxidation of pyrite (FeS2). In
the presence of oxygen and water, pyrite oxidizes to form
Fe2+, SO42− and H+ ions .
The produced Fe2+ ion then reacts with O2 to form Fe3+.
This reaction is facilitated by the sulfur-oxidizing bacteria
(Thiobacillus thiooxidans, Thiobacillus ferooxidans) as they
utilize the produced energy from this reaction for their
Fe2þ þ 0:25 O2 þ Hþ→ Fe3þ þ 0:5 H2O
In addition, the produced Fe3+ further oxidizes pyrite to
form Fe2+, SO42− and H+ ions.
FeS2 þ 14 Fe3þ þ 8H2O→15 Fe2þ þ 2SO24‐ þ 16Hþ
The abiotic rate of pyrite oxidation by Fe3+ is much higher
than oxidation by O2 and water. Due to the production of H+
ions, the pH of the whole system drops drastically and
becomes highly acidic. If the pH of the system remains over
3.5–4.0 standard units, Fe3+ precipitates in the form of
Fe(OH)3. The yellow-orange colored precipitation of iron
hydroxide is known as “yellow boy.”
Fe3þ þ 3H2O→ FeðOHÞ3 þ 3Hþ
The overall stoichiometric pyrite oxidation reaction can be
written as [
FeS2 þ 3:75 O2 þ 3:5H2O→FeðOHÞ3 þ 2SO24‐ þ 4Hþ þ heat
Due to high acidity, the mobility of the metals in the
environment increases significantly. Extensively acidic pH (as low
as 2–4 standard units) coupled with metals toxicity can elicit
severe impacts on aquatic biodiversity [
mine sites often accelerate the AMD generation process and
may require decades of proper management practices to
reclaim. The adverse environmental impacts of AMD can exist
forever if not addressed. The current number of abandoned
mines in the USA is estimated to be more than 557,000
], many of which are active sources of AMD.
Approximately 15,000 to 23,000 km of streams are currently impacted
by AMD in the USA [
2, 4, 11, 14, 15
], which also represents a
direct threat to human health. Due to its complex nature and
wide array of consequences, AMD is termed a “multiheaded
], and taming this beast is a challenging task.
According the US Forest Service (2005), the estimated cost of
cleaning up AMD-impacted sites on National Forest System
(NFS) land is around $4 billion. Between the years 1998 and
2003 around $310 million was spent on AMD-impacted NFS
land clean-up services [
]. Currently, several AMD
prevention and remediation technologies are in effect at various
AMD-impacted sites. The objective of this paper is to review
the commonly used AMD treatment technologies based on
their working principles and efficiency.
AMD Treatment Technologies
AMD treatment technologies can be divided into two major
categories: (1) prevention or source control techniques and (2)
remediation techniques. While the former focuses on
prevention of AMD generation and migration by controlling its
source, the later focuses more on the mitigational
measurements of produced AMD.
Prevention or Source Control Technologies
Safe disposal and storage of post-mining overburdens and
tailings play a vital role in AMD control. Several
sourcecontrolling techniques are available to prevent AMD
formation. As pyrite-bearing mine wastes produce AMD in the
presence of water and oxygen, one way to prevent AMD
production is the exclusion of either one or both of them from the
system. Co-disposal of pyritic materials along with some
benign material (waste rock, limestone) is the most common
practice to reduce AMD production from mine waste
]. The mixing of large waste rocks with fine tailings
is practiced sometimes which possesses higher moisture
content and hence reduce oxygen penetration through mine
]. Depending upon the neutralization potential
(NP) of the soil type, pyritic wastes are mixed with alkaline
amendments such as limestone to reduce acidity of the overall
]. Besides limestone, materials such as
fluidized bed combustion (FBC) ash and Kiln dust with higher
NP (20–70 %) are also used as alkaline amendments. In
addition to their ability to increase the net alkalinity of the system,
these materials also transform into a cement-like hard
substance which acts as a barrier and stabilization material [
]. Flooding/sealing of underground mines [
underwater storage of mine tailings and land-based storage in sealed
waste heaps are some of the commonly used techniques to
prevent AMD migration to local water bodies [
diversion of surface and groundwater from acid-producing pyritic
waste piles is another important AMD prevention approach.
Diversion ditches, grout barriers and slurry walls are some of
the techniques used to control water migration through mine
16, 17, 26
]. Encapsulation, capping and sealing of
sulfidic mine sites with non-sulfidic topsoil layer [
are often used to reduce water penetration (rainfall and runoff)
through mine spoils. Single- (for semi-arid regions) or
multilayer (for high-rainfall regions) soil covers are used for
encapsulation. The capping materials consist of a clay layer to
prevent oxygen penetration and an alkaline layer to provide a
hard capsulated barrier to prevent water from reaching the
waste piles. A coarse layer is often present to drain the
infiltrated water [
]. A vegetative top layer provides
stabilization to the overall system and retains moisture [
As sulfur-oxidizing bacteria play a vital role in the AMD
generation process, the use of bactericides such as anionic
surfactants is also a common practice. The bactericides, which
are often applied as liquid amendment or spray, can control the
AMD formation only for a limited time period [
major disadvantage of these expensive preventive
technologies is their ineffectiveness in the long-term. Most of these
techniques have failed to protect the environment against long
and persistent AMD pollution.
AMD remediation technologies can be divided into two
categories: active treatment and passive treatment.
Active Treatment Technology
The responsibility to clean-up abandoned mine sites is borne
by both private operators and government agencies. A number
of federal and state laws such as the National Historic
Preservation Act of 1966, the Clean Air Act of 1972, the Endangered
Species Act of 1973 and the Surface Mining Control and
Reclamation Act of 1977 are currently in effect in the USA
to regulate the standards of the post-mining water discharges
into the surrounding ecosystems [
]. The US Forest
Service even has the authority to administer the Comprehensive
Environmental Response Compensation and Liability Act of
1980 on National Forest System lands through an Executive
Order (No. 12580) passed in 1987 . The addition of
various acid-neutralizing and metal-precipitating chemical agents
into AMD water is a common practice to meet the effluent
discharge limits within a short time span. A wide range of
chemical agents such as limestone (CaCO3), hydrated lime
(Ca(OH)2), caustic soda (NaOH), soda ash (Na2CO3), calcium
oxide (CaO), anhydrous ammonia (NH3), magnesium oxide
(MgO) and magnesium hydroxide (Mg(OH)2) are being used
during the active treatment of AMD water worldwide [
The efficiency of each of the chemicals depends on factors
such as the site specificity (seasonal variation), daily AMD
load and metal concentration. Hence, the selection of
appropriate chemical agent is very important for the success of the
One of the major advantages of the active treatment process
is that unlike the passive treatment facilities, it does not require
any additional space or construction. Furthermore, the active
treatment process is fast and effective in removing acidity and
metals. The other advantage of the active treatment technique
is the lower cost associated with handling and disposal of
sludge in comparison to passive treatment techniques [
Although the active treatment process has several advantages,
it is not favored due to its limitations. The major disadvantage
of the active treatment process is that it requires a continuous
supply of chemicals and energy to perform efficiently. Costly
chemicals and engaging sufficient man power to maintain the
system increases the overall cost of this technology
significantly. The efficiency of these systems is completely
dependent on its regular maintenance and chemical supply,
which makes it difficult to control for most of the remotely
located abandoned mine sites. The efficiency and cost of the
systems also vary with the type of neutralizing agent used.
Limestone is inexpensive but less soluble in water and hence
less effective than the other chemical agents. Chemicals such
as hydrated lime are also inexpensive but ineffective if higher
pH (~9) is required for precipitation of metals like Mn [
]. Although NaOH is approximately 1.5 times more
effective than lime, NaOH is almost nine times more expensive
. Due to their extremely hazardous nature, chemical
agents such as NaOH and anhydrous ammonia need special
attention during handling. Also, the use of excessive ammonia
can create problems such as nitrification and denitrification in
receiving water bodies [
Passive Treatment Technology
Passive AMD treatment technologies can be classified into
two groups: conventional and emerging technologies. The
conventional passive treatment technologies such as
constructed wetlands and anaerobic sulfate-reducing bioreactors
have been used for a long time. Emerging technologies such
as phytoremediation are also being investigated for efficient
Conventional Passive Treatment Technology
1. Constructed Wetlands
Constructed wetlands are one of the most commonly used
passive AMD treatment technologies. There are two types of
wetlands: aerobic and anaerobic. Aerobic wetlands are
shallow water bodies (<30 cm in depth), which provide sufficient
retention time to oxidize and precipitate subsequent metal
hydroxides. Wetland plants such as Typha sp., Juncus sp. and
Scirpus sp. regulate the water flow, stabilize and accumulate
the metal precipitates, maintain the microbial population and
increase the aesthetic value of the contaminated site [
Wetland plants involve two major mechanisms to remove
heavy metals from AMD: phytoextraction and rizhofiltration.
In phytoextraction, metal-hyperaccumulating plants uptake
metals from wetland substrate and store them in their root
and/or shoot. In rhizofiltration, plants absorb, adsorb or
precipitate metals in the root zones (rhizosphere) [
studies often reported that the amount of metal retention inside
the wetland cells is higher than the metal uptake in the plant
]. Plants such as Typha latifolia, Scirpus
validus, Phragmites australis and Oryza sativa form plaques
in their root epidermis by producing metal oxide and
hydroxide precipitates that prevent the translocation of metals in the
plant tissues [
]. Although formation of Fe oxide and
hydroxide plaques in plant root zones is more common, Al
and Mn plaques are also reported by researchers [
Aerobic wetlands are more efficient in removing Fe, Al and
Mn in comparison to other metals. The Fe retention rate in
aerobic wetlands can vary from 0.13 to 96 % of the initial Fe
40, 42, 46, 47
]. Wetland plants such as T. latifolia,
Lemna minor, Nuphar variegatum and Potamogeton
epihydrus can remove 29–56 % of the initial Al load [
High Mn retention (~76 %) is also demonstrated by plants
such as Desmostachya bipinnata [
]. Both Al and Fe are
mainly stored in the root zone, but the distribution of Mn is
often noticed in the entire plant body. High acidity removal
(43 %) and an increase of the pH from 2.9 to 7.1 are also
observed inside the aerobic wetlands [
]. The efficiency
of wetlands in treating AMD depends on factors such as the
seasonal variations, the acidity and metal load and the dissolve
or soluble metal concentration gradient [
40, 42, 50, 51
Cost-effectiveness is one of the major advantages of aerobic
wetlands. The cost of aerobic wetlands ranges from $23 to $7,
000/t/year in terms of removal of 0.1 to 27 t/year of acidity over
a 20-year life span [
]. The amount of metal retention is
always higher than metal extraction in aerobic wetlands. Studies
showed that aerobic wetlands possess high retention capacity
for different metals such as 69 kg Al/year, 8089 kg Fe/year and
130 kg Mn/year [
]. The efficiency of the aerobic wetland
systems decreases if the influent water has a pH<5. Hence,
aerobic wetlands are always associated with other passive
treatment systems such as anoxic limestone drains (ALDs) or
vertical flow wetlands (VFWs) and receive net alkaline AMD
water from them [
18, 35, 49, 52
]. Aerobic wetlands cannot
remove sulfate [
] and are less effective when metal
concentrations are very high in the system [
Anaerobic wetlands are built with organic-rich
substrates, which provide reducing conditions and
neutralizing agents such as limestone. Often anaerobic wetlands
are constructed underground and are devoid of
vegetation. In this kind of a system, net acidity of AMD water
is removed by the dissolution of limestone and the
metabolism of iron- and sulfate-reducing bacteria. The
organic-rich substrates are prepared by mixing of
biodegradable products such as manure with straw, peat and
sawdust. This mixture serves as a long-term food source
for the indigenous anaerobic iron- and sulfate-reducing
bacteria due to their slow biodegradation rates. A variety
of manures such as chicken, cow and horse litter and
mushroom compost are used as substrates for the
microbial community [
17, 18, 53, 54
]. Sometimes, the
anaerobic wetlands are engineered as the reducing and
alkalinity-producing system (RAPS)  or as the
successive alkalinity-producing system (SAPS) (where
multiple RAPS are used) [
]. In this type of system, AMD
first flows downward through a compost layer, which
removes dissolve oxygen (DO) and facilitates iron and
sulfate reduction. Subsequently, the AMD passes through
a limestone and gravel bed, which adds alkalinity. To
precipitate and retain the iron hydroxides, water from
the RAPS system is channeled through a settling pond
or aerobic wetland. In anaerobic wetlands, the sorption
of metals occurs on the organic substrates through
exchangeable or complexation reactions. Initially, 50–
80 % of metal removal from the AMD inside the
anaerobic wetland system takes place due to sorption, which
decreases over time due to the substrate saturation [
]. The retention of metals as of oxide, hydroxide,
carbonate and sulfide precipitates also occurs in anaerobic
wetlands. Unlike sorption reactions, precipitation of
metals is not time-limited and depends on the density
and volume of the wetland cells. The total Fe removed
from AMD water by anaerobic wetland systems is
dominated by Fe hydroxides (~50–70 %) and Fe sulfides
(~30 %). Iron hydroxide often reduces to Fe2+ by
anaerobic iron-reducing bacteria, and this reaction increases
the pH of the system.
FeðOHÞ3 þ 0:5 H2→ Fe2þ þ 2OH‐ þ H2O
Anaerobic sulfate-reducing bacteria produce iron mono
and disulfides while reducing the sulfate present in the
AMD water. The reduction of sulfate also increases the pH
of the system [
17, 53, 58–60
Anaerobic wetlands can remove approximately 0–67.9 t/
year of net acidity and costs between $341 and $4762/t/year
]. The removal of sulfate and increase in pH are some of
the major advantages of the anaerobic wetlands. The
anaerobic wetlands can also reduce the acidity and Fe concentration
of the AMD water by 3–76 and 62–80 %, respectively [
The major disadvantage of the anaerobic wetlands is the
decrease of its efficiency over time. The saturation of substrates
occurs within a span of 1–7 months as most of the available
exchangeable and complexation sites become saturated with
metals. Sometimes, the addition of organic matter is required
to revive the filtering efficiency of the wetland [
The efficiency of the anaerobic wetlands also changes with
seasonal variation and wetland age [
]. The lifetime of
the system can be severely affected if the plants above the
ground penetrate the system’s protective cover through their
roots and introduce oxygen to the anaerobic layers [
A pilot passive treatment plant was constructed in 1994 at
Wheal Jane Mine in Cornwall, England, for long-term AMD
treatment. The project was unique because it employed both
aerobic and anaerobic wetland facilities. After appropriate
lime dosing, AMD water was allowed to pass through serious
of anoxic cell, anoxic limestone drain, five aerobic cells,
anaerobic cell and rock filter. Data show that this kind of hybrid
system is capable of removing Fe and sulfate between 55 and
92 %, and 3 and 38 %, respectively. This system can also
remove other metals such as Cd, Cu and Zn depending on
the pretreatment and flow rate of the AMD [
2. Anaerobic Sulfate-Reducing Bioreactors
Anaerobic sulfate-reducing bioreactors are another type of
widely used passive treatment technology, which involves
sulfate-reducing bacteria to remediate AMD.
Sulfatereducing bacteria are a group of chemoorganotrophic and
strictly anaerobic bacteria, which is primarily represented by
t h e g e n e r a o f D e s u l f o v i b r i o , D e s u l f o m i c ro b i u m ,
Desulfobacter and Desulfotomaculum.
Anaerobic sulfate-reducing bioreactors are made up of
a thick layer of organic-rich materials mixed with
limestone. An additional thin layer of limestone is also used
under the organic layer, which provides the additional
alkalinity and also supports the underlying drainage
channels. The AMD passes vertically through the organic
layer and limestone bed and is discharged through the
drainage system. The organic layer serves as the
substrate of sulfate-reducing bacteria. In this layer,
sulfatereducing bacteria reduce SO42− to H2S and oxidize
organic matter (CH2O) to bicarbonate ions (HCO3−) [
Sulfate-reducing bacteria use the energy produced in this
reaction for their growth and development.
SO24‐ þ 2CH2O→H2S þ 2HCO3‐
The reaction of AMD with limestone causes limestone
dissolution and produces HCO3− and Ca2+.
CaCO3 þ Hþ→Ca2þ þ HCO3 ‐
The produced HCO3− further reacts with H+ ions and
produces CO2 and water. Hence, the consumption of the H+ ions
results in the increase of the pH of the overall AMD water. At
high pH, metals start to precipitate in the form of metal
sulfides, oxides, hydroxides and carbonates.
In the anaerobic sulfate-reducing system, the most
common form is metal sulfide precipitation [
HS‐ þ M2þ→MS þ Hþ
In reaction [
], M2+ represents divalent metals such as
Fe2+, Cu2+, Pb2+ and Zn2+ and MS represents the produced
metal sulfide. Metals can also precipitate in the form of
hydroxide or carbonate [
Fe2þ þ HCO‐ →FeCO3 þ Hþ
Al3þ þ 3H2O→AlðOHÞ3 þ 3Hþ
Mn2þ þ HCO‐ →MnCO3 þ Hþ
Thus, sulfate-reducing bioreactors help in reducing acidity,
metal and sulfate concentration of AMD water and improve
the overall water quality. The efficiency of an anaerobic
sulfate-reducing bioreactor depends on various factors. The
amount of sulfate removed is dependent on the available
surface area and hydraulic retention time (HRT), while the rate of
sulfate removal is dependent on the initial sulfate
concentration in AMD [
]. Studies have been conducted to test the
efficiency of sulfate-reducing bacteria under various pH
levels. Researchers found that pH in the range of 5–8 is best
for optimum activity of the sulfate-reducing bacteria, as the
inhibition of sulfate reduction and the increase in the solubility
of metal sulfides occur at low pH [
]. Some studies also
found that although at low pH (2.8–3.5) sulfate-reducing
bacteria can survive due to their acid tolerance, their sulfate
removal efficiency dropped to 14–35 % [
]. Several studies
have been conducted to characterize the sulfate-reducing
bacterial community. Researchers found that the type of
sulfatereducing bacterial community change through time depending
on the nature of the wastewater and the type of the food
sources. Species such as Desulfovibrio desulfuricans and
Desulfobulbus rhabdoformis are dominant in a
sulfatereducing bioreactor [
]. Change of dominant bacterial
community from iron oxidizing Betaproteobacteria in
pretreated AMD water to sulfur-oxidizing Epsilonproteobacteria
and complex carbon degrading Bacteroidetes and Firmicutes
phylums in post-treated water is also observed .
Studies have been conducted to evaluate the efficiency of
the sulfate-reducing bioreactors. It is observed that the
efficiency varies from 39 to 82 % removal of the initial SO42−
load (900–2981 mg/L) [
bioreactors exhibit a high metal removal ability, and they can remove
98–99 % of initial Cu [
], 85–90 % of initial Fe [
] and 95–99 % of initial Al [
] load from the AMD
water. A net decrease in acidity and increase in pH of the
influent AMD water can also be achieved through the
70, 72, 75, 78
The activity of sulfate-reducing bacteria is the rate-limiting
factor of the anaerobic sulfate-reducing bioreactors. A near
neutral pH, reducing environment, continuous supply of
organic carbon and sulfate, solid support for microbial
attachment and the formation and retention capacity of precipitated
metal sulfides are some of the key factors of an efficient
sulfate-reducing bioreactor. Extremely low pH (below 3.5)
severely impacts the efficiency of the sulfate-reducing bacteria
]. Low temperature also impacts the acclimatization of the
sulfate-reducing bacteria significantly, but after
acclimatization, they can be active and functional even in the cold
climates (1–16 °C). A decrease in overall efficiency of
sulfatereducing bioreactors has been observed during the winter
]. Despite their higher sulfate and metal removal
efficiency, the sulfate-reducing bioreactors often fail to
perform over long-term mainly due to the exhaustion of the
substrates required for sustaining the sulfate-reducing
3. Other Commonly Used Passive Treatment Techniques
Anoxic limestone drains (ALD) are one of the commonly
used passive AMD treatment systems. ALDs are typically
30 m long, 1.5 m deep and 0.6–20 m wide underground
systems filled with limestone. Only anoxic water is introduced in
the ALDs, which are impervious to air and water. In ALD,
limestone reacts with AMD water and produces CO2 which
cannot escape from the system and raises the overall alkalinity
]. Due to the anoxic condition, the iron remains in the
reduced form inside the ALDs, and the formation and
precipitation of iron hydroxide does not occur. The optimal
performance of the ALD can be attained if the AMD channeled
through it contains no ferric iron, aluminum or DO. The pH
of ALD systems needs to be 6.0 because under more acidic
conditions, metals like Fe and Al precipitate as hydroxides
and form coats or armors on limestone [
]. Thus, iron
hydroxide precipitation severely impacts the efficiency of the
ALDs. ALDs can produce up to 275 mg/L of net alkalinity
in comparison to 50–60 mg/L of net alkalinity produced by an
open system in equilibrium [
]. A decrease of acidity by 50–
80 % can be achieved through ALDs [
]. The major
drawback of ALD is its longevity. The presence of ferric iron
and Al in AMD water can form hydroxide precipitates which
reduce the permeability and efficiency of the ALD systems
. Typically, ALDs are used as a part of the hybrid passive
treatment system in corporation with the aerobic and
anaerobic wetlands [
17, 18, 81
Vertical flow wetlands (VFW) or permeable reactive
barriers (PRB) are another type of passive AMD treatment
system. In a VFW or PRB, AMD water flows through an
organicrich layer followed by a limestone bed before discharging
through a drainage system. The VFW systems reduce ferric
to ferrous iron and decrease the amount of DO. Sulfate
reduction and Fe sulfide precipitation can take place in this system.
A series of drainage pipes placed below the limestone layer
carry the water to aerobic ponds where ferrous ions oxidize
and precipitate [
Limestone leach beds (LSB), slag leach beds (SLB) and
open limestone channels (OLC) play a significant role in
various AMD passive treatment systems. LSBs are ponds
constructed to receive waters with little or no alkalinity and
dissolved metals. These ponds are packed with limestone and
designed to have retention time of at least 12 h. The limestone
layer can be replenished when necessary. Alkalinity in this
system can reach 75 mg/L [
]. In SLB ponds, a bed of steel
slag fines is used to remediate AMD water, which need to be
devoid of metals such as Fe, Al and Mn. This system can
produce alkalinity up to 2000 mg/L, and the overall system
is easy to replenish [
]. OLCs are open channels or trenches
lined with limestone. In OLCs, limestone coated with Fe and
Al hydroxides are used to decrease the limestone dissolution
over time. The performances of the OLCs are dependent on
different variables such as slopes, pH, flow velocity and
thickness of the coating of limestone [
]. OLCs can remove 4–
69 % of acidity, 72 % of Fe and 20 % of Mn and Al from
AMD water [
17, 35, 83
]. OLCs are generally constructed with
the combination of other passive treatment systems. The
major advantage of OLC is its low-cost as it does not require any
maintenance once constructed properly [
The construction of the passive treatment technologies
depends on several factors such as characteristics of waste, flow
rate, size of the construction area, local topography and
environment. Figure 1 provides a decision-making tree for passive
treatment systems based on the characteristics of the influent
AMD water. Most of the time, the adaptation of a hybrid
system is necessary to achieve the regulatory standards before
discharging the AMD water into the local water bodies. The
installation costs of the conventional passive treatment
technologies are very expensive, and these systems also require a
periodic monitoring and maintenance . The passive
treatment facilities also generate a considerable amount of sludge,
and the removal and disposal cost of the sludge is also very
Emerging Passive Treatment Technology: Phytoremediation
Phytoremediation is an emerging passive AMD treatment
technology. Researchers and remediation practitioners are
evaluating phytoremediation-based alternatives because of
the higher costs associated with conventional AMD
remediation approaches. Phytoremediation can be applied to both
AMD-impacted soil and water. As eroded AMD-impacted
soils generally end up in surrounding water bodies and elevate
the risk, remediation of both soil and water is very important.
Phytoremediation of contaminated mine sites mainly involves
two mechanisms: phytoextraction and phytostabilization. In
the phytoextraction process, plants extract heavy metals from
the contaminated sites and store the extracted metals in their
biomass. On the other hand, phytostabilization provides a
vegetative cover to highly erosion prone and heavily
contaminated acid sulfate soils [
]. Sometimes, due to the
presence of heavy metals in high concentration, complete
metal removal cannot be possible. In such conditions,
phytostabilization immobilizes the metals and traps them in
plant root zones, which minimizes the metal exposure to the
surrounding ecosystems. The extensive root systems of the
plants also protect the soils against erosion and leaching.
Several metal tolerant plant species have been used to
r e m e d i a t e c o n t a m i n a t e d m i n e s i t e s . S u c c e s s o f
phytoremediation depends on the proper selection of the
metal-hyperaccumulator plants. Hyperaccumulator plants
generally accumulate metals in their aboveground biomass
Fig. 1 Decision-making tree for
the design of passive treatment
system [Redrawn after 35, 84–86]
at a concentration that is 100-fold greater than the
nonhyperaccumulator plants. Generally, these plants accumulate
up to or more than 0.1 % of metals such as Cu, Pb, Cd, Cr, Ni
and Co or 1 % of metals such as Zn and Mn in their dry
]. The high-accumulation factor (AF) and
hight r a n s l o c a t i o n f a c t o r ( T F ) a r e a l s o s o m e o f t h e
hyp eraccu mu latio n chara cteristic s. More than 400
hyperaccumulator plant species belonging to families such
as Brassicaceae, Asteraceae and Poaceae exist, which can be
used in metal-contaminated mine sites [
36, 41, 87
]. Table 1
presents some of the most commonly used plants for
remediation of AMD-impacted sites.
In China, a wide range of plant species (Chrysopogon
zizanioides, Sesbania rostrate, P. australis, Cyperus
alternifolius, Leucaena leucocephala, Panicum repens,
Gynura crepidiodes, Alocasia macrorrhiza and Chrysopogon
aciculatus) have been used to phytoremediate AMD water
highly contaminated with Zn, Pb and SO42− [91, 97, 98].
Plants like C. alternifolius and C. zizanioides possess very
high-acid tolerance characteristics. An increase of pH from
Analyze Mine Water for
Physical & Chemical
Characteristics, Flow Rate
and Loading Rate
Low DO, Low Fe
and Low metals
Analyze DO, Fe, Heavy Metals
High DO, High Fe
and High metals
Net Alkaline Water
Effluent Discharge Parameter Evaluation:
Regulatory Standard Achieved?
Adaptation of other
2.4 to 7.5 and 80 % removal of its initial sulfate concentration
are also noticed during the study . In Australia, plant
species like Juncus usitatus, Lomandra longifolia, Cynodon
dactylon, Pteridium esculentum, Acacia decurrens and
Melaleuca alternifolia are used for remediation of metals such as
Fe, As, Cd, Cu, Pb and Zn from both AMD-impacted soil and
water . All of the plant species thrived well under the
acidic conditions (pH ranged from 2.9 to 5.6), and species like
C. dactylon can accumulate metals like Cd (14 mg/kg), Pb
(658 mg/kg) and Zn (828 mg/kg) in its biomass. Species like
J. usitatus and L. longifolia can also accumulate significant
amount of Cd in their biomass (26 and 21 mg/kg,
respectively). Another potential plant species for remediation of Cd- and
Zn-contaminated mine sites is Thlaspi caerulescens. Studies
reported that T. caerulescens can accumulate as high as 50–
250 mg/kg Cd and 13,000–19,000 mg/kg Zn while growing in
AMD-contaminated sites [87, 92]. Due to the low biomass
production, T. caerulescens is not an ideal plant for
phytoremediation. On the other hand, plants like Cichorium
intybus L. and C. dactylon are potential phytoremediation
candidates for Pb-contaminated mine sites. C. intybus and
C. dactylon can accumulate as high as 800–1500 and 400–
1200 mg/kg Pb in their biomass, respectively . In a similar
study, it is observed that Atriplex halimus L. can accumulate
830 and 440 mg/kg of Cd and Zn, respectively, in its biomass
while growing on mine tailing under greenhouse condition
. Another commonly used plant species for mine site
remediation is C. zizanioides, commonly known as vetiver
grass. Due to their physiological characteristics and high
tolerance of metals such as Al, Mn, Fe and Zn [95, 96] and heavy
metals such as As, Pb, Hg and Cd, vetiver can be used
efficiently to restore metal-contaminated sites . Vetiver can
tolerate Fe concentrations even up to 63,920 mg/kg .
Vetiver can remediate iron ore tailings contaminated with high
concentration of metals such as Fe, Zn, Mn and Cu and can
accumulate as high as 545–1197 mg/kg Fe, 302–531 mg/kg
Zn, 415–648 mg/kg Mn and 13–66 mg/kg Cu in its root and
shoot. High-mean translocation factors for Mn (0.86), Fe
(0.71), Zn (0.69) and Cu (0.55) can be observed in vetiver’s
t i s s u e [ 9 6 ] . U s e o f s o i l a m e n d m e n t s l i k e D T PA
(diethylenetriamine pentaacetic acid) and compost mixture
increases the metal uptake ability of vetiver. Vetiver possesses
a massive root system, which can stabilize the erosion prone
acid sulfate soil. So, planting vetiver on metal-contaminated
mine soils can stabilize the soil and improve the overall soil
quality [96, 98]. Once established, vetiver grass can grow on
the acidic soils with continuous acidity production by sulfidic
minerals . In a study conducted in Queensland Australia, it
was found that vetiver systems are able to control bank
erosion while growing on acid sulfate soil . The study
showed that planting vetiver stabilized the edges of the
channel and also promoted the establishment of other plants on the
steep slopes, helping to prevent erosion and preventing the
collapse of the highly acidic soil into the channel streams.
Vetiver can trap sediments and pollutants from runoff water,
which improves the overall water quality. The increase of pH
and decrease of Fe concentration in water were also observed
during the study .
Phytoremediation of AMD-impacted soil and water has
shown positive results and fueled extensive research in this
field worldwide. The major advantages of phytoremediation
are that it is cost-effective and environment-friendly. The
success of phytoremediation is primarily dependent on the plant
availability of the metals. Due to factors such as soil
properties, metal species, loading level and soil-ageing, the amount
of plant available metal varies significantly. Several chemical
agents and soil amendments such as EDTA
(ethylenediaminetetraacetic acid), EDDS (ethylenediamine-N,N′-disuccinic
acid), compost and DTPA have been applied to increase the
plant available metal fraction in the soil. Most of the
phytoremediation studies were performed in either under
greenhouse conditions or in the field on a pilot scale. Hence,
more extensive field-based research is required to optimize
this emerging technique.
Remediation of AMD is a challenging proposition that is
dependent on several factors such as the daily AMD load, flow
rate, net acidity and metal concentration. The pre-mining
analysis of the neutralization potential (NP) of soil through acid
base accounting (ABA) helps to predict the nature of AMD
and to adapt best AMD management practices. A number of
AMD prevention and remediation technologies are being used
worldwide to prevent AMD pollution in both active and
abandoned mines. Long-term monitoring of the constructed
systems is necessary as AMD pollution can exists for decades.
Most of the conventional passive AMD remediation
technologies are ineffective and/or expensive for long-term and
persistent AMD load. Hence, a search for an effective, viable and
sustainable AMD remediation technology is ongoing.
Emerging passive treatment technologies such as phytoremediation
have the potential to be successful and are attractive because
of sustainability and cost-effective aspects of their
implementation. However, most of the research in this area so far has
been limited to greenhouse or pilot-scale field studies. Further
long-term research is needed in order for this promising
technology to be widely implemented in AMD-impacted areas.
Acknowledgments This study is supported by the United States
Department of the Interior, Office of Surface Mining Reclamation and
Enforcement under OMB No.: 4040–0004. ARC gratefully acknowledges
the PhD Program in Environmental Management for Doctoral
Assistantship and the Center for Writing Excellence (CWE) at Montclair State
University for proofreading the manuscript.
Conflict of Interest On behalf of all authors, the corresponding author
states that there is no conflict of interest.
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