Eng.
Life
Sci.
Metal Removal
Review
Microbial Processes of Heavy Metal Removal from
Carbon-Deficient Effluents in Constructed Wetlands
By D. B. Kosolapov, P. Kuschk*, M. B. Vainshtein, A. V. Vatsourina, A. Wieûner, M. Kästner, and R. A. Müller
This paper reviews the main microbial processes involved when toxic metals are removed from wastewater in constructed
wetlands. Microbial activity is thought to play a key role in the detoxification of these metals. The paper concentrates on the
microbial processes which affect the mobility, the toxicity and bioavailability of metals, namely biosorption, metal sulfide precipitation by sulfate reducers, redox transformations, and methylation, as well as microbe-plant interactions. These reactions
result in either the precipitation and accumulation of metals in wetland soils, or their volatilization and emission into the atmosphere. The possibilities of optimizing the microbially mediated reactions for the development of wetland technology are
discussed as a long-term metal retention strategy.
1 Introduction
Many metals and metalloids play a specific role in functions of living organisms as microelements (e.g., Fe, Mn, Mg,
Ni, Zn, Cu, etc.) serving as components of enzymes, structural proteins and pigments, and maintaining the ionic balance
and osmotic potential of cells. However, high concentrations
of metals can pose a severe threat to biota and human
health. The contamination of water, soil, and air by toxic
metals is a growing environmental problem all over the
world due to the activities of metal processing, surface treatment, and the mining industry, the burning of fossil fuels,
and the uncontrolled landfilling of waste [1].
Due to the toxic, persistent, bio-accumulative, and synergistic effects of some metals on biota, their cycling and fate
in the environment are of great concern. The threat of metal
pollution to public and environmental health has encouraged interest in developing systems that can remove metal
contamination from soil and water, or at least neutralize its
harmful effects. Most contaminated environments contain
mixtures of pollutants, the most troublesome components
usually turning out to be metals. Unlike organic contaminants, which can be degraded into harmless chemical species,
metals, by their very nature, cannot be destroyed. However,
they can be transformed into more (or less) mobile forms by
modifying their chemical and physical characteristics.
±
[*]
D. B. Kosolapov, Russ. Acad. Sci., Institute of Biology of Inland Waters,
152742, Borok, Nekouz, Yaroslavl, Russia; A. Wieûner, M. Kästner,
P. Kuschk (author to whom correspondence should be addressed, e-mail:
peter.kuschk@ufz.de), UFZ-Umweltforschungszentrum Leipzig-Halle
GmbH, Department of Bioremediation, Permoserstrasse 15, D-04318
Leipzig, Germany; M. B. Vainshtein, A. V. Vatsourina, Russ. Acad. Sci.,
Institute of Biochemistry and Physiology of Microorganisms, Pushchino,
Russia; R. A. Müller, UFZ-Umweltforschungszentrum Leipzig-Halle
GmbH, Umwelt- und Biotechnologisches Zentrum, Permoserstrasse 15,
D-04318 Leipzig, Germany.
Eng. Life Sci. 2004, 4, No. 5
DOI: 10.1002/elsc.200420048
The principle objective of wastewater treatment is to eliminate or reduce contaminants to levels that cause no adverse
effects on humans or the receiving environment. Several
methods for doing so already exist. However, the traditional
active treatment processes employed such as reverse osmosis, ion exchange, microfiltration and the addition of chemicals are either not very efficient or quite costly. In some
cases they are simply unfeasible. Alongside physicochemical
techniques, biotechnology also presents some interesting
possibilities. The advantages of biological approaches include higher specificity than physical and chemical methods,
and their suitability for in situ methodologies (e.g., avoiding
high energy or toxic chemical addition). Current biotechnological approaches are moving towards the application of in
situ strategies in an attempt to reduce costs and avoid pollution dispersal problems caused by the transportation of contaminated soil, sediment or water to treatment plants.
Constructed or artificial wetlands (CWs) are one of the alternatives to conventional water treatment technologies and
harbor great potential for the treatment of contaminated
waters [2±5]. This cost-effective, ecologically friendly ªpassiveº technology is used extensively worldwide to treat a
range of wastewater and effluent from a variety of sources,
including mine drainage, landfill leachates, urban stormwater, and agricultural runoff. CWs are man-made ecosystems that mimic their natural counterparts. These complex
ecosystems of plants, microorganisms and substrate act together as a biogeochemical filter, efficiently removing low
levels of contamination from very large volumes of water
and hence providing protection for natural water resources
such as rivers, lakes, estuaries and ground waters. CWs do
not require as much maintenance as conventional systems
but they do require regular maintenance, especially of pretreatment units for the removal of suspended solids in the
case of subsurface flow CWs. An important disadvantage is
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that in comparison to other treatment methods CWs usually
need a considerably large area.
By the end of the 20th century, over a thousand CWs had
been built just to treat mine drainage [6]. However, the efficiency of CWs for the treatment of metal-contaminated
waters, in particular acid mine drainage, has been shown to
be variable and often unpredictable. The rate and extent to
which a given metal is removed varies depending on the
metal involved and the type of wetland [7, 8]. Some researchers even consider CWs to be a potentially non-sustainable technology for the treatment of water contaminated
with metal owing to the nature of the non-biodegradable
pollutants to be removed [9].
The efficiency of a CW for wastewater treatment is generally evaluated on the basis of influent-effluent comparison,
the ecosystem being considered a ªblack boxº. However, this
approach does not provide any information about the biogeochemical processes occurring in the wetland. Nevertheless, the information necessary to predict long-term efficiency and to improve planning and management in CWs has
been recently improved [10, 11]. Despite the experience accumulated over years of practical application and research, a
number of fundamental aspects of exactly how CWs function
are not yet adequately understood. In particular, the removal
mechanisms for each metal and the interactions between
them are still hazy. One reason for this is that, compared to
other technologies, such as activated sludge, CWs depend on
the interaction of far more different components [5].
The aim of this article is to provide an overview of the
main mechanisms by which microorganisms remove heavy
metals from wastewater in constructed wetlands.
2 Constructed Wetland Technology
CWs have been shown to be an effective technology for
treating water contaminated by toxic metals. A major advantage of this technology is that, as a ªpassiveº treatment system, the costs of operation and maintenance are significantly
lower than for active treatment processes. CWs use natural
biogeochemical processes inherent in a wetland ecosystem
to accumulate and remove metals from influent waters. The
immobilization and mobilization of metals are effected in
wetlands through a number of processes operating independently in some situations and interactively in others, thus
making the whole process of metal removal very complex
[5, 12, 13]. These processes include a combination of abiotic
and biotic reactions which occur in oxic and anoxic zones of
CWs. Among them are sedimentation, flocculation, absorption, precipitation, co-precipitation, cation and anion exchange, complexation, oxidation and reduction, microbial
activity and plant uptake [14]. The extent to which these reactions occur depends on the type of the CW, the pH value,
redox status, influent water composition, the dominant plant
species and microbial activity.
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The way in which a wetland is constructed ultimately determines how wastewater treatment occurs and what mechanisms will be involved. CWs typically consist of four principal components participating in pollutant removal:
substrates (soil), plants, water, and associated microbial
populations. Wetlands comes in two main types:
i) surface flow (ªaerobicº) wetlands consisting of vegetation planted in shallow, relatively impermeable soil, clay
or mine spoil, and
ii) sub-surface flow (ªanaerobicº) wetlands consisting of vegetation planted in deep, permeable mixture of
substrates such as soil, peat moss, spent mushroom compost, sawdust, straw/manure, hay bales and gravel, often
underlain with limestone.
In aerobic CWs, the dominant treatment processes occur
mainly in the shallow surface layer. In anaerobic wetlands,
the water primarily flows through the substrate and treatment involves major interactions within the substrate [15]. It
has been suggested that anaerobic or sub-surface flow systems may be more successful, particularly for treating acidic
wastewaters [16].
3 Microbial Potential for the Bioremediation of
Metal-Contaminated Environments
Prokaryotic microorganisms have co-existed with metals
almost since life began and have developed abilities to cope
with the toxic metals in a variety of environments. Microorganisms strongly influence the environmental fate of toxic
metals and metalloids with a multiplicity of mechanisms affecting metal speciation and mobility. These mechanisms are
integral components of both natural biogeochemical cycles
of metals and other elements and have the potential to be
used for the treatment of solid and liquid waste [17, 18].
The ability of microorganisms to affect metal speciation
stems from their ability to affect mobilization or immobilization processes that influence the balance of metal species
between soluble and insoluble phases. As far as environmental biotechnology is concerned, solubilization may
facilitate the removal of metals from solid matrices such as
soils, sediments, dumps and industrial waste. Alternatively,
immobilization processes may enable metals to be transformed into insoluble and chemically inert forms, and can
also be employed to remove metals from aqueous solution
and thus restrict their spread and threat to the environment and human health. Microorganisms can mobilize
metals through autotrophic and heterotrophic leaching,
chelation by metabolites and siderophores, redox transformations, and methylation, which can result in volatilization. The main processes leading to the immobilization of
metals are redox reactions, which are complemented by
others such as precipitation, the sorption of metals by bacteria, algae and plants, organic substrates, and ferric hydroxides [19±23].
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Metal Removal
4 Biosorption
In CWs, the main way to recover metals from wastewaters
is through the use of the complex interactions of microorganisms with plants. Algal and microbial biomass on their
own only sequester metals temporarily and if not harvested,
this process does not contribute to a stable fixation. Microorganisms are able to concentrate metals to levels that are substantially higher than those encountered in the environment.
The processes of active (energy-dependent) and passive (energy-independent) metal uptake are termed bioaccumulation and biosorption, respectively [19]. Biosorption is the
passive metabolism-independent sequestration of metals by
interactions with live or dead microbial biomass and is an
important approach for the bioremediation of metal-contaminated environments. Many studies have demonstrated
the efficiency of metal removal by bacterial, cyanobacterial,
algal or fungal biomass under a range of conditions. The
sorption of metals to microbial cells is governed by a multiplicity of mechanisms and interactions, including ion exchange, chelation, adsorption, and entrapment, but not all of
them are fully understood [21, 24]. Biosorption may involve
various functional groups on the biomass, including carboxyl, sulfonate, phosphate, hydroxyl, amino, etc. Some bacteria
are able to accumulate metals inside their cells, forming
amorphous mineral inclusions [25]. The sorption of metals
to cells seems to play a critical role in all microbe-metal interactions. Interactions with specific groups on the surface of
the cell may enhance or inhibit metal transport and thus
their transformation. Bacteria, algae and fungi, as well as
plants and animals produce a range of specific and nonspecific metal-binding compounds that may bind toxic metal ions
and also adsorb or entrap particulate substances such as precipitated metal sulfides and oxides [22, 26]. Metal sorption
and uptake by microbial cells may be improved by natural
selection, the genetic manipulation of existing genomes, and
direct physical and chemical manipulations of cells [18].
However, the storage of metals by microbial biomass is relatively short-term due to the short life cycle of microorganisms, and so afterwards other metal retention processes need
to be involved.
5 Metal Precipitation by Sulfate Reduction
Microorganisms may efficiently immobilize toxic metals
through their precipitation, either as the result of dissimilatory reduction or by interactions with products of microbial
metabolism. Bacteria, fungi and algae produce sulfide, hydroxide, carbonate and phosphate, which may react with
metals, forming highly insoluble metal precipitates. This
strategy has attracted great interest in the bioremediation of
metal-contaminated environments.
Dissimilatory sulfate reduction and the subsequent precipitation of metal sulfides have been identified as the most important reactions in metal removal from wastewaters [27].
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By the action of sulfate-reducing bacteria H2S is formed
which precipitates metals such as iron, zinc, copper, nickel,
etc. In addition, this bacterial sulfate reduction lowers the sulfate concentration and causes an increase of the pH. Sulfate
reducers oxidize a range of organic compounds or hydrogen
coupled with the reduction of sulfate, producing sulfide. The
reaction mediated by sulfate reducers is as follows:
SO42± + 2 CH2O ® H+ + HS± + 2 HCO3±
(1)
Sulfate reduction results in a decrease in sulfate and an increase in bisulfide and HCCO3±. Hydrogen sulfide gas may
then be formed form bisulfide and hydrogen ions, especially
at pH values < 7:
H+ + HS± ® H2S (g)
(2)
The loss of H2S into the atmosphere and also the production of HCO3± represent a decrease in the acidity, raising the
pH and buffering the solution. HS± may react with a variety
of metals and result in a metal sulfide precipitate:
H2S + M2+ ® MS¯ + 2H+
(3)
(where CH2O and M2+ represent a simple organic compound and a divalent metal ion such as Fe2+, Cd2+, or Zn2+,
respectively).
Sulfate reducers may create extremely reducing conditions, which promote abiotic metal reduction. In addition,
lowering the acidity of a system as a result of the shift in
equilibrium when sulfate (dissociated) is converted into sulfide (largely protonated) can prompt the further precipitation of metals as hydroxides and increase the efficiency of
sulfide precipitation [28].
Sulfate reduction may enable both the in situ and ex situ
bioremediation of metal-contaminated environments
[20, 27]. Dissimilatory sulfate reduction is considered one of
the most important processes involved in the long-term retention of metals in artificial and natural wetlands [3, 29±32].
For instance, in wetland microcosms that have received diluted landfill leachate amended with Pb and Cd, nearly all
these metals were present in the sediments as sulfides, limiting the bioavailability and toxicity of these elements [33].
The formation of sulfides in wetlands may enable the
long-term removal of metals since the metal sulfides may
stay in the sediments permanently as long as they remain anoxic. As the solubility of most metal sulfides is very low,
even the moderate production of sulfide by bacteria can
lower metal concentrations to environmentally permissible
levels. In most sediment habitats, dissimilatory sulfate reduction is the sole mechanism responsible for the formation of
iron monosulfides and pyrite [34]. Iron monosulfides can be
destroyed by oxic and acidic conditions. Because pyrite is
much more resistant to acids, CWs in which FeS is rapidly
transformed into FeS2 would be more resistant to the effects
of environmental perturbations [35].
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Sulfate reducers are ubiquitous and tolerate a wide range
of environmental conditions. In freshwater wetlands, sulfate
reduction rates are generally limited by the amount of
sulfate. In CWs, however, sulfate concentrations are usually
greater because of wastewater loading, and the sulfate reduction rates are mostly dependent on the substrate supply.
Therefore, a major factor limiting the application of microbial sulfate reduction to the removal of metals from carbondeficient mine drainages and industrial effluents in wetland
systems is the availability of carbon and energy sources to
drive the process [36].
Marked ecological differences between the wetlands are
reflected in the composition and activity of sulfate-reducing
consortia, which display significant differences in terms of
substrate utilization, patterns of sulfide generation and metal removal from contaminated waters [31]. Fermentative,
nitrogen-fixing and sulfate-reducing bacteria were shown to
be the key components of the microbial community in a CW
treated contaminated mine water. As well as the addition of
a carbon source, sulfate reduction rates were controlled by
dissolved nitrogen concentrations [36]. In CWs, for sulfate
reduction to be effective for treating wastewater, factors
which promote the process and sulfide formation must be
maximized and destructive factors minimized. First of all, in
order to stimulate the sulfate-reducing microorganisms in
the case of carbon-deficient effluents, a proper carbon
source should be provided to enhance their growth and to
cause other bacteria to remove the oxygen from the environment.
6 Redox Transformations
6.1 Dissimilatory Reduction of Metals
Anaerobic dissimilatory metal-reducing bacteria and archaea are able to reduce a variety of metals to a lower redox
state by using them as terminal electron acceptors in anaerobic respiration. Microbial redox reactions can mobilize or
immobilize metals depending on the metal species involved.
Metal-reducing microorganisms may decrease the mobility
of certain metals, resulting in metal precipitation. If the
reduced forms of metals are less soluble, toxic and bioavailable, this approach could be used for bioremediation. The
dissimilatory reduction of iron, chromium, uranium, vanadium, technetium, gold, and some other metals and metalloids is performed by various microorganisms in a diversity
of environments and is widely used for waste treatment
[17, 21, 23, 37, 38].
The treatment of chromium-containing waste is a good
example of the bioremediation potential of microorganisms.
Although many different oxidation states of chromium exist,
Cr(III) and Cr(VI) are the most stable. While there are
natural sources of chromium in the environment, the majority of Cr(VI) originates from industrial activities. Chromium
waste is commonly associated with nuclear power plants as
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well as industrial sources, such as tanneries, rust-proofing
and metal plating. The reduction of highly soluble, toxic and
mutagenic Cr(VI) to the water-insoluble and significantly
less toxic Cr(III) results in precipitation as hydroxides and
offers a promising bioremediation strategy [39].
Cr(VI) reduction is widespread in microorganisms and occurs under both oxic and anoxic conditions, with NADH and
electron transport systems serving as the respective electron
donors [40±42]. In particular, some sulfate-reducing bacteria
may share physiological properties of both sulfate- and
metal-reducers, and grow with Cr(VI) and other metals as
electron acceptors [43]. The reduction of Cr(VI) by sulfatereducing bacteria was shown to be catalyzed by cytochrome
c3 [38]. Indirect mechanisms that also promote Cr(VI) reduction in contaminated sediments are mediated by biogenic
sulfide [44, 45] and Fe(II) [46]. A determined sequence exists
for the accumulation of metabolic products in bacterial cultures: for example, the consortium of anaerobic bacteria
started to produce sulfide after chromium (VI) and nitrate
had been exhausted [47]. Environmental factors affecting
Cr(VI) reduction were reviewed recently and include competing electron acceptors, pH, temperature, redox potential,
and the presence of other metals [41]. Some aerobic and
anaerobic bacteria are known to be able to use organic contaminants as electronic donors for Cr(VI) reduction, indicating potential for the in situ treatment of mixed waste
[48, 49].
Another priority pollutant, uranium (VI), can be reduced
to insoluble uranium (IV) by Fe(III)-reducing microorganisms, e.g., Geobacter metallireducens, this reduction in solubility forming the basis for uranium removal from contaminated water [48]. The ability to reduce U(VI) enzymatically
is not restricted to iron-reducing bacteria; other microorganisms including a Clostridium sp. and sulfur- and sulfate-reducing bacteria such as Desulfovibrio desulfuricans are also
responsible for uranium reduction. The reaction is mediated
by cytochrome c proteins and requires electron donors, such
as lactate or hydrogen. U(IV) is not reduced further, forming uraninite mineral by extracellular precipitation [38, 50].
Due to active bacterial U(VI) reduction, as well as sorption
and complexation processes, wetlands near the old mines act
as efficient filters and may contain anomalous concentrations of uranium (sometimes as high as 3000 mg/L) as well
as other metals in the sediments [51].
Desulfovibrio desulfuricans and some other bacteria may
also couple the oxidation of a range of electron donors to
the reduction of Tc(VII) (in the form of the soluble pertechnetate ion, TcO4±), mediated by periplasmic dehydrogenase.
As a result, a lower-valency oxide of Tc(IV) is precipitated
on the cell peripheries, implying the potential of this reaction for the treatment of Tc-contaminated wastewater [38].
The ability to reduce vanadium is widespread amongst
bacteria and fungi. The microbial reduction of V(V) is
responsible for the precipitation of this element in anoxic
environments and may be used to remediate vanadium-contaminated ore-processing waste streams [37]. Anaerobic
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V(V)-reducing bacteria are able to utilize a wide range of
electron donors including sugars, amino acids and hydrogen.
V(V) has been shown to be reduced to blue-colored V(IV),
and possibly further to V(III), forming a black precipitate
[52]. Various aspects of the microbial reduction of these and
other metals were reviewed in detail by Lloyd [38].
Whereas the insoluble products of the reduction of most
metals remain in the contaminated matrix, the reduction of
Hg(II) results in relatively nontoxic volatile elemental mercury (Hg(0)) being lost to the atmosphere. The volatilization
of mercury mediated by bacterial mercuric reductase
(MerA) is a good example of how metal transformations can
be exploited in wastewater treatment, and this process also
harbors the potential for environmental bioremediation.
The microbial activity effects mercury removal by volatilization to the atmosphere as an alternative to metal immobilization strategies. Natural and engineered Hg2+-resistant
bacterial strains have been successfully applied to remove
mercury from wastewater [53]. A combined method of
chemical leaching and the subsequent volatilization of mercury by bacteria has been developed that removed about
70 % of mercury from polluted Minamata Bay sediments
[54]. However, although the volatilization of mercury may
solve a local problem, there is public concern that it might
eventually contribute to global atmospheric pollution.
Fe(III)-reducing microorganisms greatly influence the biogeochemical cycles of iron and other metals, as well as the
fate of organic matter and nutrients in a variety of habitats
[17, 37, 55]. Because the most important environmental factors such as labile organic matter and reducible Fe(III) are
abundant in wetland sediments, many wetlands are considered sites of active microbially mediated Fe cycling [56]. A
wide range of anaerobic bacteria and archaea are able to
conserve energy though the reduction of Fe(III) to Fe(II).
Many of these prokaryotes also have the ability to grow
through the reduction of Mn(VI) to Mn(II). Iron- and manganese-reducing microorganisms can dissolve insoluble
Fe(III) and Mn(IV) oxides, resulting in the release of soluble
Fe(II) and Mn(II), as well as the trace metals bound by the
Fe(III) or Mn(IV) minerals. Fe(III)- and Mn(IV)-reducing
microorganisms can thus affect the fate of other contaminant metals through both direct enzymatic reduction and indirect reduction catalyzed by biogenic Fe(II) and Mn(II).
These microorganisms have a detrimental effect on wetland
efficiency as they remobilize iron and other metals and may
therefore export metals from the system. The bacterial
reductive dissolution of ferric iron minerals follows the
equation:
4 Fe(OH)3 + CH2O ® 4 Fe2+ + HCO3± + 7 OH± + 3 H2O (4)
As mentioned above, the dissimilatory reduction of
Cr(VI), U(VI), and several other metals by iron-reducing
microorganisms can result in these toxic metals being immobilized in sediments. Therefore, since the solubility of both
iron and manganese is increased by microbial reduction and
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neither of these metals poses a major toxic threat, the most
environmentally relevant reactions mediated by dissimilatory iron-reducing bacteria involve other metals.
6.2 Oxidation of Fe(II) and Mn(II)
Dissolved reduced forms of iron and manganese (Fe(II)
and Mn(II)) are oxidized by abiotic reactions and bacteria,
respectively, and then precipitated mainly as hydroxides.
The oxidation of Fe(II) is ubiquitous in metal-contaminated
environments, such as mine drainage waters and tailings
piles, drainage pipes and irrigation ditches, sediments and
bogs, and plant rhizospheres [22]. The chemical reaction involves the oxidation of ferrous iron (Fe2+) in solution to a
ferric iron (Fe3+) by iron-depositing bacteria as follows:
Fe2+ + H+ + O2 ® Fe3+ + H2O
(5)
The ferric iron then reacts with water to form an insoluble
iron hydroxide:
Fe3+ + 3 H2O ® Fe(OH)3¯ + 3 H+
(6)
The precipitation of iron and manganese oxides, caused by
abiotic oxidation and further microbial processes, is thought
to be the dominant process in metal removal in aerobic
zones of CWs. Aerobic wetlands, because of their extensive
water surface and slow flow, promote bacterial metal oxidation and subsequent hydroxylation, and thus cause the precipitation and retention of Fe, Mn, and Al hydroxides [57].
Indeed, 40±70 % of the total iron removed from acid mine
drainage by the CW was found in the form of ferric hydroxides [58]. The precipitated oxyhydroxides of iron and manganese strongly absorb other heavy metals such as Cu, Pb,
Ni, Co and Cr, thus removing these elements from treated
contaminated waters [14]. While being effective at removing
metals from contaminated water, processes of iron oxidation
and hydrolysis can also lower the water's pH [29].
Manganese is also removed from polluted water through
the microbial oxidation of the bivalent form to the tetravalent state. The Mn(IV) is then precipitated mainly as MnO2.
Manganese oxidation occurs more slowly than iron oxidation
and is inhibited by the presence of Fe(II). Consequently, Fe
and Mn precipitate sequentially rather than simultaneously
in CWs. As a result, Mn precipitation takes place mainly in
the last stages of wetland flow systems, after Fe has been precipitated. Portions of Mn(II) are also precipitated as MnS
and Mn(OH)2 [59]. In different parts of CWs, both aerobic
and anaerobic conditions occur, and the metals of concern in
wastewaters can be precipitated as (oxy)hydroxides and sulfides. For example, the CW has been shown to be effective at
treating acid mine drainage. As a result of the activity of
iron-oxidizing and sulfate-reducing bacteria in different wetland cells, more than 98 % of Fe, 95 % of Ni, and 45 % of S
were removed from the seepage [60]. In this connection,
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multi-cell wetlands using different mechanisms of metal
removal supported either by aerobic or anaerobic conditions
were recommended for wastewater treatment [61].
A striking feature of some wetland plants is the presence
of metal-rich rhizoconcretions on the roots known as Fe
plaques. These structures are formed by the abiotic and microbial oxidation of ferrous iron and are composed primarily
of Fe(III)(hydro)xides and other metals such as manganese
precipitated on the root surface. The rhizoconcretions may
be 5±10 times more enriched in Cd, Cu, Pb, and Zn than the
surrounding sediment environments [62].
7 Methylation of Metals
The biomethylation of Hg, Pb, Sn, Te, As, and Se with the
production of volatile derivatives such as dimethylmercury,
dimethylselenide or trimethylarsine is a well-known phenomenon. The methylation of metals and metalloids is mediated by a range of aerobic and anaerobic bacteria, as well as
fungi, algae, plants and animals, which enzymatically transfer methyl groups to the metals. The methylated compounds
formed differ in their solubility, toxicity, and volatility, and
may be eliminated from the system by evaporation. Most
volatile metal compounds exhibit higher toxicity than their
inorganic species since organic derivatives are lipophilic and
thus more biologically active [63].
Wetlands are favorable to the bacterial methylation of
mercury and are known contributors of methylmercury
(CH3Hg) to downstream lakes and rivers [64]. The formation of methylmercury appears to be one potential way of
mercury removal in CWs. However, methylmercury is an organic form of mercury that is lipophilic, extremely toxic, and
readily accumulated by aquatic organisms. Moreover,
methylmercury is most efficiently transferred up the food
chain to higher trophic levels [65]. Previous studies have
demonstrated that sulfate reducers are the major biotic contributors to methylmercury formation in marine and freshwater sediments [66]. Therefore, in CWs with sulfate added
to enhance metal removal through sulfide-mediated precipitation, parameters such as the composition and activity of
the sulfate-reducing microbial community need to be identified to optimize sulfate reduction while minimizing methylmercury formation. The complexation of sulfide with Hg(II)
to produce insoluble cinnabar (HgS) has been described as a
principal mechanism of decreasing mercury bioavailability
in aquatic systems [67]. Organomercurials ± such as methylmercury ± may be also detoxified by bacterial organomercurial lyase, the resulting Hg(II) then being reduced to volatile Hg(0) by mercuric reductase [68].
8 Plant-Microbe-Metal Interactions
Wetland plants or helophytes are important in many different ways for metal retention by wetlands. One of the most
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widely recognized processes for metal removal in CWs is
probably plant uptake. However, the direct uptake of metals
by plant biomass may represent only a minor proportion of
the total metals removed by some CWs. At any rate, helophytes play a critical role in metal removal via filtration, adsorption, and cation exchange, and through plant-induced
chemical changes in the rhizosphere. Plants also provide
habitat and organic compounds, which support the microbial
activity in soils and the formation of complexes with metals
[11, 12]. In wetlands, the most active microbe-metal interactions occur in the rhizosphere of plants. The rhizosphere,
which can be defined as a compartment of the soil or
sediment directly influenced by plant roots, is important in
processes determining the mobility of metals, and thus their
toxicity, bioavailability and behavior in the ecosystem.
Wetland soils are generally considered a sink for metals
and may contain very high concentrations of metals in a
reduced state in the anoxic zone, where metals are mainly
bound to sulfides. However, wetland plants alter the redox
conditions, the pH and organic matter content of sediments
and so affect the chemical speciation and mobility of metals.
Metals may be mobilized or immobilized depending on the
combination of factors involved, making it difficult to predict how plants will actually affect metal mobility under given conditions [69]. So depending on specific environmental
conditions, the extent of oxygen input via the roots and the
root deposition of organic material by water plants can cause
various redox conditions or steep concentration gradients
within the rhizosphere.
The effect of plants on sediment redox potential and thus
metal mobility depends upon characteristics of wetland
vegetation and sediment properties. If anoxic conditions
remain constant, toxic metals can precipitate out of solution
in sulfide-rich sediments (such as mine tailings and salt
marshes), potentially being permanently immobilized. Oxygen diffused from the roots to the sediments may induce
sulfide oxidation, the release of sulfide-bound metals and
acid generation, causing severe environmental problems.
On the other hand, the release of oxygen by roots may decrease metal mobility by the microbial oxidation and precipitation of Fe and co-precipitation of the other metals. The precipitation of iron (oxy-)hydroxides in the rhizosphere due to
oxygen release and microbial activity is known to potentially
lead to the formation of iron plaques (see above). This also
leads to the immobilization of other metals such as zinc and
arsenic with high adsorption affinities with the iron (oxy-)hydroxides and which may co-precipitate [70]. Vegetation contributes to the composition and concentration of organic
matter in sediments through senescence, root tissue sloughing and the release of exudates that may directly or indirectly
affect metal mobility and toxicity. The remobilization of metals may result from the acidification of the rhizosphere by
plant exudates. Changes of Eh and pH values in sediments
cause changes in metal speciation and solubility, which can
result in a flux of metals from sediments to porewater and
then to overlying water, increasing their uptake by plants.
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In turn, rhizobacteria can protect the partner plants by
either directly antagonizing a pathogen or stimulating the
defense responses of wetland plants. Symbiotic root-colonized bacteria and fungi may play a protective role, alleviating metal toxicity and also increasing the efficiency of phytoremediation by promoting the accumulation of metals and
metalloids in tissues of helophytes [71]. However, a better
understanding of the effects of rhizospheric microorganisms
on metal uptake by helophytes would promote the development of the bioremediation potential of CWs.
In CWs used to treat metal-containing effluents, although
the plants have been shown to accumulate some metals,
their concentrations may be minor compared to overall metal retention by the wetland. In a flow-through wetland treatment system constructed to treat the coal combustion waste
product leachate from a power station, most of the metals
(Fe, Mn, Co, Ni) removed were accumulated in sediment,
which constituted the largest sink. Metal accumulation
tended to be greater in the surface layer of sediments as well
as in the rhizomes of cattail (the dominant species of the
CW). The accumulation of metals in living shoot tissues of
cattail and submerged living Chara (a macroalga) tissues
were relatively minor sinks in comparison with the sediments [72]. The accumulation of Mn, Zn, Cu, Ni, and Cr by
helophytes may account for only a small percentage of the
removal of the annual metal load delivered to natural and
constructed wetlands. For example, Fe and Mn levels in the
wetland plants were only 1 and 2 %, respectively, of the annual Fe and Mn load received by these wetlands [73]. These
data suggest that the main contribution of wetland plants to
metal removal is not through the direct uptake of metals,
but instead through substrate stabilization, rhizosphere oxidation, the supply of organic matter for microorganisms, the
provision of sites for microbial attachment, and water transport into the wetland soil by plant transpiration.
Metals have been shown to accumulate in wetland soils
around the roots. The mechanisms of this phenomenon are
as follows. The oxidation of ferrous iron to ferric iron leads
to the precipitation of iron hydroxides in the rhizosphere,
which in turn leads to a decrease in the concentration gradient of dissolved iron towards the plant roots. The iron hydroxides then co-precipitate other metals, again creating a
decreasing concentration gradient of these elements towards
the roots. These gradients lead to the diffusion of metals in
the direction of the roots [74]. Plants and, to a less extent,
burrowing benthos organisms have been shown to have a
significant effect on the accumulation of metals in wetland
sediments. For example, concentrations of iron and arsenic
were significantly higher in vegetated/inhabited salt marsh
sediments compared to nearby bulk sediments [75].
Thus, while wetland sediments, which are mainly anoxic
and reduced, act as sinks for metals, they can become a
source of metal contaminants as a result of plant activities.
Plants can oxidize sediments mobilizing metals. Nevertheless, the wetlands generally act as sinks rather than sources
for metals. They are effective traps for immobilizing toxic
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metals, with relatively low export to adjacent natural aquatic
ecosystems [69, 72].
The mechanisms and efficiency of wastewater treatment
in CWs have been shown to vary seasonally and with wetland age [7, 10, 11]. Wetland ecosystems, especially in temperate regions, are characterized by periodicities of biotic
processes at diurnal, seasonal and long-term annual scales.
The growth of plants and microbial processes are subject to
high seasonal variation. Seasonal variations of environmental variables (temperature, solar radiation, organic matter
input, precipitation, evapotranspiration) cause changes in
microbial activity, which in turn affect the removal of metals
and create limits in the degree of water quality improvement
by CWs [76].
9 Conclusions
A wide range of physical, chemical and biological processes contribute to the removal of metals from wastewater
in CWs. These processes are integral components of the biogeochemical cycling of metals and nutrients in the environment and are potential mechanisms for bioremediation. Specific metal species may be removed through storage in the
wetland sediment or losses to the atmosphere. Microorganisms play the key role in affecting metal mobility, toxicity
and bioavailability. Despite the many applications of CWs,
understanding of the microbial detoxification mechanisms is
still patchy. Knowledge of the basic microbial processes controlling metal removal in wetlands will substantially promote
the improved efficiency of applications of wetland treatment
systems as a long-term treatment strategy.
While CWs have been effective at removing metals from
contaminated water, the final fate of the metals is a question
of concern, since long-term applications may lead to secondary contamination of the environment. If metal concentrations in CWs continue to rise, they may eventually pose a
threat to wildlife. The immobilized metals remain in the sediment of CWs and can be sensitive to changes to the oxidation-reduction conditions. Depending on the specific situation, changes in redox potential may increase the mobility of
metals. The redox conditions in CWs must be continuously
monitored and any restrictions necessary can be implemented concerning the use of the wetland. Knowledge of longterm metal behavior in CWs is important in order to predict
their efficiency.
Although there has been a variety of engineering modifications for CWs, much less attention has been paid to optimizing the microbial processes responsible for metal removal. An understanding of biotic interactions may prove to
be crucial to designing an effective wetland system incorporating microbial processes for achieving the highest rates of
metal removal from effluents. Using CWs for the treatment
of mixed waste containing both metals and organic contaminants is one of the main approaches.
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Further study of microbial activity in CWs ought to improve in situ strategies for the treatment of metal-contaminated wastes and may also provide a novel insight into
overall wetland structure and function. An enhanced understanding of the complex biotic and abiotic interactions
involved along with the technical possibilities will enable
wetland technologies to be used on a broader scale.
Acknowledgements
Funding for this study was kindly provided by Linkage
NATO grant no. EST-CLG-978918.
Received: June 7, 2004 [ELS 48]
Received in revised form: July 16, 2004
Accepted: July 29, 2004
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