Remediation technologies for heavy metal contaminated groundwater.

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Review
Remediation technologies for heavy metal contaminated groundwater
M.A. Hashim a,*, Soumyadeep Mukhopadhyay a, Jaya Narayan Sahu a, Bhaskar Sengupta b
aDepartment of Chemical Engineering, University of Malaya, Pantai Valley, 50603 Kuala Lumpur, Malaysia
b School of Planning, Architecture and Civil Engineering, Queen’s University Belfast, David Keir Building, Belfast BT9 5AG, UK
a r t i c l e i n f o
Article history:
Received 6 January 2011
Received in revised form
17 May 2011
Accepted 3 June 2011
Available online 25 June 2011
Keywords:
Groundwater
Heavy metals
Remediation technology
Soil
Water treatment
a b s t r a c t
The contamination of groundwater by heavy metal, originating either from natural soil sources or from
anthropogenic sources is a matter of utmost concern to the public health. Remediation of contaminated
groundwater is of highest priority since billions of people all over the world use it for drinking purpose.
In this paper, thirty five approaches for groundwater treatment have been reviewed and classified under
three large categories viz chemical, biochemical/biological/biosorption and physico-chemical treatment
processes. Comparison tables have been provided at the end of each process for a better understanding of
each category. Selection of a suitable technology for contamination remediation at a particular site is one
of the most challenging job due to extremely complex soil chemistry and aquifer characteristics and no
thumb-rule can be suggested regarding this issue. In the past decade, iron based technologies, microbial
remediation, biological sulphate reduction and various adsorbents played versatile and efficient reme-
diation roles. Keeping the sustainability issues and environmental ethics in mind, the technologies
encompassing natural chemistry, bioremediation and biosorption are recommended to be adopted in
appropriate cases. In many places, two or more techniques can work synergistically for better results.
Processes such as chelate extraction and chemical soil washings are advisable only for recovery of
valuable metals in highly contaminated industrial sites depending on economical feasibility.
� 2011 Elsevier Ltd. All rights reserved.
1. Introduction
“Heavy metal” is a general collective term, which applies to the
group of metals and metalloids with atomic density greater than
4000 kg m�3, or 5 times more than water (Garbarino et al., 1995)
and they are natural components of the earth’s crust. Although
some of them act as essential micro nutrients for living beings, at
higher concentrations they can lead to severe poisoning (Lenntech,
2004). The most toxic forms of these metals in their ionic species
are the most stable oxidation states e.g. Cd2þ, Pb2þ, Hg2þ, Agþ and
As3þ in which, they react with the body’s bio-molecules to form
extremely stable biotoxic compounds which are difficult to disso-
ciate (Duruibe et al., 2007).
In the environment, the heavy metals are generally more persis-
tent than organic contaminants such as pesticides or petroleum
byproducts. They can become mobile in soils depending on soil pH
and their speciation. Soa fractionof the totalmass can leach to aquifer
or can become bioavailable to living organisms (Alloway, 1990;
Santona et al., 2006). Heavy metal poisoning can result from
drinking-watercontamination (e.g. Pbpipes, industrialandconsumer
wastes), intake via the food chain or high ambient air concentrations
near emission sources (Lenntech, 2004). In the past decade, Love
Canal tragedy in the City of Niagara, USA demonstrated the devas-
tating effect of soil and groundwater contamination on human pop-
ulation (Fletcher, 2002). The diffusion phenomenon of contaminants
through soil layers and the change in mobility of heavy metals in
aquifers with intrusion of organic pollutants are being studied in
moredetails in recentyears (Cuevaset al., 2011;Satyawali et al., 2011).
Over the past few decades, many remediation technologies were
applied all over the world to deal with the contaminated soil and
aquifers. Many documents and reviews on these technologies for
remediating organic and inorganic pollutants are available (Diels
et al., 2005; Evanko and Dzombak, 1997; Khan et al., 2004;
Mulligan et al., 2001; Scullion, 2006; USEPA, 1997; Yin and Allen,
1999). Review on heavy metal removal from waste waters is also
published recently (Fu and Wang, 2011). Apart from the report by
USEPA (1997), no document reviewing the heavy metal
Abbreviations: AMD, acid mine drainage; BET, Brunauer, Emmett and Teller;
BSR, biological sulphate reduction; EDTA, ethylenediaminetetraacetic acid; FMBO,
ferric and manganese binary oxides; GAC, granular activated carbon; HA, humic
acid; HFO, hydrous ferric oxides; HRT, hydraulic residence time; ISBP, in-situ bio-
precipitation process; PHC, peanut husk carbon; PRB, permeable reactive barrier;
SPLP, synthetic precipitation leaching procedure; SRB, sulphate reducing bacteria;
TCLP, toxicity characteristics leaching procedure; UF/EUF, ultrafiltration/electro-
ultrafiltration; USEPA, United States Environment Protection Agency; ZVI, zero
valent iron.
* Corresponding author. Tel.: þ603 7967 5296.
E-mail address: alihashim@um.edu.my (M.A. Hashim).
Contents lists available at ScienceDirect
Journal of Environmental Management
journal homepage: www.elsevier .com/locate/ jenvman
0301-4797/$ e see front matter � 2011 Elsevier Ltd. All rights reserved.
doi:10.1016/j.jenvman.2011.06.009
Journal of Environmental Management 92 (2011) 2355e2388

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remediation technologies for groundwater in the recent times is
available. A technology functioning successfully under some oper-
ating conditions, inherently possess some limitation by virtue of
which it may not function as effectively in other conditions. So,
a document, summarizing all the applied and emerging technolo-
gies for heavy metal groundwater and soil remediation, along with
their scopes, advantages and limitations will come handy for the
scientific research community for designing newer technologies as
well as in the decision making process of the heavy metal affected
community trying to fish out a suitable solution for their problem.
So, in this review, we have focused on the removal of only heavy
metals from groundwater, i.e. thewater which is located in soil pore
spaces and in the fractures of rock units. Groundwater is entirely
related with the soil through which it flows. So, in the course of our
review, we came across many soil remediation technologies, some
of which were relevant to groundwater remediation as well have
been discussed.
In the past, some technologies were applied for removing only
petroleum products, some for inorganic solvent removal, while
some were earmarked for heavy metal removal. Of late, this barrier
has been diminishing as researchers around the world are
combining various technologies to achieve desirable result. All the
reviewed technologies have been classified under three categories
viz Chemical Technologies, Biological/Biochemical/Biosorptive
Technologies and Physico-Chemical Technologies. In some cases,
these technologies overlapped. However, this is the consequence of
the changing face of the science and technology in the modern
world where interdisciplinary studies are gaining ground over
compartmentalized field of studies.
2. Heavy metals in ground water: sources, chemical property
and speciation
Heavy metals occur in the earth’s crust and may get solubilised
in ground water through natural processes or by change in soil pH.
Moreover, groundwater can get contaminated with heavy metals
from landfill leachate, sewage, leachate from mine tailings, deep-
well disposal of liquid wastes, seepage from industrial waste
lagoons or from industrial spills and leaks (Evanko and Dzombak,
1997). A variety of reactions in soil environment e.g. acid/base,
precipitation/dissolution, oxidation/reduction, sorption or ion
exchange processes can influence the speciation and mobility of
metal contaminants.. The rate and extent of these reactions will
depend on factors such as pH, Eh, complexation with other
Table 1
Speciation and chemistry of some heavy metals.
Heavy metal Speciation and chemistry Concentration limits References
Lead Pb occurs in 0 and þ2 oxidation states. Pb(II) is the more
common and reactive form of Pb. Low solubility compounds
are formed by complexation with inorganic (Cl�, CO2�3 , SO
2�
4 , PO
3�
4 )
and organic ligands (humic and fulvic acids, EDTA, amino acids).
The primary processes influencing the fate of Pb in soil include
adsorption, ion exchange, precipitation and complexation with
sorbed organic matter
Surface agricultural soil: 7e20 ppm (Bodek et al., 1988; Evanko and
Dzombak, 1997; Hammer and
Hammer, 2004; Smith et al., 1995;
WHO, 2000)
Soil levels: up to 300 ppm
USEPA, Maximum
Contaminant Level
(MCL) in drinking water: 0.015 ppm
Chromium Cr occurs in 0, þ6 and þ3 oxidation states. Cr(VI) is the dominant
and toxic form of Cr at shallow aquifers. Major Cr(VI) species
include chromate ðCrO2�4 Þ and dichromate ðCr2O
2�
7 Þ (especially
Ba2þ, Pb2þ and Agþ). Cr (III) is the dominant form of Cr at low
pH (<4). Cr(VI) can be reduced to Cr(III) by soil organic matter,
S2- and Fe2þ ions under anaerobic conditions. The leachability of Cr(VI)
increases as soil pH increases
Normal groundwater concentration:
< 0.001 ppm
(Lenntech, 2004; Smith et al., 1995)
Lethal dose : 1e2 g
MCL of USEPA in drinking water:
0.1 ppm
Zinc Zn occurs in 0 and þ2 oxidation states. It forms complexes with
a anions, amino acids and organic acids. At high pH, Zn is bioavailable.
Zn hydrolyzes at pH 7.0e7.5, forming Zn(OH)2. It readily precipitates
under reducing conditions and may coprecipitate with hydrous
oxides of Fe or manganese
Natural concentration of Zn in soils:
30e150 ppm
(Evanko and Dzombak, 1997;
Lenntech, 2004; Smith et al., 1995)
Concentration in plant: 10e150 ppm
Plant toxicity: 400 ppm
WHO limit in water: 5 ppm
Cadmium Cd occurs in 0 and þ2 oxidation states. Hydroxide (Cd(OH)2) and
carbonate (CdCO3) dominate at high pH whereas Cd2þ and aqueous
sulphate species dominate at lower pH (<8). It precipitates in the
presence of phosphate, arsenate, chromate, sulphide, etc. Shows
mobility at pH range 4.5e5.5
Soil natural conc: >1 ppm (Matthews and Davis, 1984;
Smith et al., 1995)Plant conc: 0.005e0.02 ppm
Plant toxicity level: 5e30 ppm
USEPA MCL in water: 0.005 ppm
Arsenic As occurs in �3, 0, þ3, þ5 oxidation states. In aerobic environments,
As(V) is dominant, usually in the form of arsenate (AsO4)3-. It behaves
as chelate and can coprecipitate with or adsorb into Fe oxyhydroxides
under acidic conditions. Under reducing conditions, As(III) dominates,
existing as arsenite (AsO3)3� which is water soluble and can be
adsorbed/coprecipitated with metal sulphides
MCL in drinking water- (Bodek et al., 1988;
Smith et al., 1995)USEPA: 0.01 ppm
WHO: 0.01 ppm
Iron Fe occurs in 0, þ2, þ3 and þ6 oxidation states. Organometallic
compounds contain oxidation states of þ1, 0, �1 and �2. Fe(IV) is
a common intermediate in many biochemical oxidation reactions.
Many mixed valence compounds contain both Fe(II) and Fe(III) centers,
e.g. magnetite and prussian blue
Tolerable upper intake level (UL) -
Dietary Reference Intake (DRI):
(Holleman et al., 1985; Medscape)
For adults: 45 mg per day
For minors: 40 mg per day
Mercury Hg occurs in 0, þ1 and þ2 oxidation states. It may occur in alkylated
form (methyl/ethyl mercury) depending upon the Eh and pH of the
system. Hg2þ and Hg2þ2 are more stable under oxidizing conditions.
Sorption to soils, sediments and humic materials is pH-dependent
and increases with pH
Groundwater natural conc:
>0.0002 ppm
(Bodek et al., 1988;
Smith et al., 1995)
USEPA regulatory limit in drinking
water: 0.002 ppm
Copper Cu occurs in 0, þ1 and þ2 oxidation states. The cupric ion (Cu2þ) is the
most toxic species of Cu e.g. Cu(OH)þ and Cu2(OH)22þ.In aerobic alkaline
systems, CuCO3 is the dominant soluble species. In anaerobic environments
CuS(s) will form in presence of sulphur. Cu forms strong solution
complexes with humic acids
Soil natural conc: 2e100 ppm (Dzombak and Morel, 1990;
LaGrega et al., 1994)Normal range in plants: 5e30 ppm
Plant toxicity level: 30e100 ppm
USEPA MCL in water: 1.3 ppm
M.A. Hashim et al. / Journal of Environmental Management 92 (2011) 2355e23882356

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dissolved constituents, sorption and ion exchange capacity of the
geological materials and organic matter content. Ground-water
flow characteristics is vital in influencing the transport of metal
contaminants (Allen and Torres, 1991; Evanko and Dzombak, 1997).
The toxicity, mobility and reactivity of heavy metals depend on
its speciation, which again depends upon some conditions e.g. pH,
Eh, temperature, moisture, etc. In order to determine the speciation
of metals in soils, specific extractants are used to solubilize different
phases of metals. Through sequential extraction with solutions of
increasing strengths, a precise evaluation of different fractions can
be obtained (Tessier et al., 1979). The chemical form and speciation
of some of the important metals found at contaminated sites are
discussed in Table 1.
3. Technologies for treatment of heavy metal contaminated
groundwater
Several technologies exist for the remediation of heavy metals-
contaminated groundwater and soil and they have some definite
outcomes such as: (i) complete or substantial destruction/degrada-
tion of the pollutants, (ii) extraction of pollutants for further treat-
ment or disposal, (iii) stabilization of pollutants in forms less mobile
or toxic, (iv) separation of non-contaminated materials and their
recycling from polluted materials that require further treatment and
(v) containment of the polluted material to restrict exposure of the
wider environment (Nathanail and Bardos, 2004; Scullion, 2006).
In this review, we have divided the treatment technologies into
the following classes: i. Chemical Treatment Technologies ii. Bio-
logical/Biochemical/Biosorptive Treatment Technologies, iii.
Physico-Chemical Treatment Technologies.
The technologies that have been used over the past few years
and are undergoing further tests in laboratory are also discussed.
The overall classification has been pictorially presented in Fig. 1.
3.1. Chemical treatment technologies
Groundwater contaminants are often dispersed in plumes over
large areas, deep below the surface, making conventional types of
remediation technologies difficult to apply. In those cases, chemical
treatment technologies may be the best choice. Chemicals are used
to decrease the toxicity or mobility of metal contaminants by
converting them to inactive states. Oxidation, reduction and
neutralization reactions can be used for this purpose (Evanko and
Dzombak, 1997). Reduction is the method most commonly used
(Yin and Allen, 1999). All the chemical treatment processes dis-
cussed in this section are summarized in Table 2.
3.1.1. In-situ treatment by using reductants
When groundwater is passed through a reductive zone or
a purpose-built barrier, metal reductions may occur. Based on both
laboratory and field studies, an appropriately created reduced zone
can remain in reducing conditions for up to a year (Amonette et al.,
1994; Fruchter et al., 1997). Manipulation of sub-surface redox
conditions can be implemented by injection of liquid reductants,
gaseous reductants or reduced colloids. A six-step enhanced design
methodology as proposed by Sevougian et al. (1994) for in-situ
chemical barriers is shown in Fig. 2.
Yin and Allen (1999) enlisted several soluble reductants such as
sulfite, thiosulphate, hydroxylamine, dithionite, hydrogen sulphide
and also the colloidal reductants e.g. Fe(0) and Fe(II) in clays for soil
remediation purpose.
3.1.1.1. Reduction by dithionite. Dithionites can reduce redox sensi-
tivemetals such as Cr, U and Th to less toxic oxidation states (Yin and
Allen, 1999). Dithionites can be injected just downstream of the
contaminant plume to create a reduced treatment zone formed by
reducing Fe(III) to Fe(II) within the clay minerals of the aquifer
sediments. The flowing contaminants will either be degraded or be
immobilizedwhile passing through the zone. Amonette et al. (1994)
conceptualized the dithionite ion as two sulfoxyl ðSO�2 �Þ radicals
joined by a 2.39 pm SeS bond which was considerably longer, and
hence weaker than typical SeS bonds (2.00e2.15 pm). Thus, S2O
2�
4
tends to dissociate into two free radicals of SO�2 � : S2O
2�
4 ¼ 2SO
�
2 �.
Although direct reduction of trivalent structural Fe(III) in clay
minerals by dithionite and strongly alkaline solutions was proposed
by Sevougian et al. (1994) for smectite (Equation (1)), it was also
likely to be caused by thehighly reactive free radicals SO�2 � as shown
in equation (2) (Amonette et al., 1994; Sevougian et al., 1994).
2Ca0:3
�
Fe2ðIIIÞAl1:4Mg0:6
�
Si8O20ðOHÞ4nH2Oþ2Na
þ
þS2O
2�
4
þ2H2O42NaCa0:3
�
FeðIIIÞFeðIIÞAl1:4Mg0:6
�
Si8O20ðOHÞ4nH2O
þSO2�3 þ4H
þ
(1)
Fig. 1. Classification of groundwater heavy metal remediation technologies.
M.A. Hashim et al. / Journal of Environmental Management 92 (2011) 2355e2388 2357

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Table2
Chemicaltreatmenttechnologies:comparativeoverview.
TechnologyScopeConditionsand
modesofapplication
AdvantageDisadvantageMechanismandprocessSelectedreferences
1.In-situreductionprocesses
1.1ReductionbydithionitesRedoxsensitive
elements(Cr,U,Th)
alkalinepHandhigh
permeabilityofsoil.
Injectedinaquifer.
Activeoverlarger
area;Longlastingeffect
Toxicgasintermediate;
Handlingisdifficult
Reductiveprecipitation
atalkalinepH
(Amonetteetal.,1994;
Fruchteretal.,1997;
Sevougianetal.,1994)
1.2ReductionbyH2S(g)Redoxsensitive
metals(Cr)
In-situapplicationby
carriergasmedium
Nosecondarywaste
generation
Toxicgasintermediate;
Gasdeliverytoaquifer
isdifficult
Sulphideoxidizedtosulphate
andmetalisprecipitated
ashydroxide
(ThorntonandAmonette,
1999;Thorntonand
Jackson,1994)
1.3ReductionbyFebasedtechnologies
1.3.1UsingZVIand
ColloidalFe
CrO2�
4,TcO
�
4,UO2þ
2,AsInjectionofFe0colloid
intreatmenttrench
orinaquifer
Canbeinjectedindeep
aquiferswithouttoxic
exposure;Regenerationpossible
Productionoftoxic
intermediateinaquifer;
Modellingisdifficult;
Barrierintegritycannot
beverified
Reductiveprecipitationof
heavymetalsandsorption
onsurfaceadsorptionsitesofZVI
(Cantrelletal.,1995;
Gillhametal.,1993;
Manningetal.,2002;
SuandPuls,2001)
1.3.2CrremovalbyferroussaltCr(III),Cr(VI)Acidifiedferroussulphate
solninjectedin
wellsandtrenches
BesidesCrO2�
4,itcan
treatTcO
�
4,UO2þ
2andMoO
�2
2.
Cr(III)maygetoxidized
totoxicandmobileCr(VI)
underalkalinecondition;
Lesserdepthofsoil
Reductiveprecipitation
ofCr(VI)asCr(OH)3orasthe
solidsolutionFexCr2�x(OH)3
(CL:AIRE,2007;Hong
etal.,2007;Pulsetal.,
1999;Seamanetal.,1999)
2.Soilwashing
2.1In-situsoilflushingAwiderangeof
heavymetalse.g.
Cr,Fe,Cu,Co,Al,
Mn,Mo,Ni.
Surfaceflooding,sprinklers,
basininfiltration
systems,leachfields,
injectionwells
Suitableforusingathighly
contaminatedindustrialsites
Mobilizedcontaminants
mayescapeintoenvironment
ifnottrappedproperly.
Washingsolutiontreatment
isdifficult
DesorptionofmetalsatlowerpH
andrecoveringofleachateby
pumpandtreatsystemfromaquifer
(McPhillipsandLoren,
1991;Mooreetal.,1993;
USEPA,1995,1997)
2.2In-situChelateFlushingPb,Cd,Cr,Hg,Cu,
Zn,Fe,As
In-situinjectionof
chelatese.g.EDTA,NTA,
DTPA,SDTC,STC,K2BDET
Ligandsactatverylowdose;
Stablecomplexesformed;
Chelatescanberegenerated
Somechelatesarepersistant,
toxic;Expensiveprocess
Formationofstablechelate
complexesbetweenchelate
andcontaminants
(Blueetal.,2008;Hong
etal.,2008;Limetal.,
2004;Warshawsky
etal.,2002)
2.3In-situremediationby
selectiveionexchange
HeavyMetalsand
TransitionMetals
In-situuseofsynthetically
preparedtypeII
SIRsandionexchange
resinsinPRBs
selectivelyremovelowlevel
ofmetalionsfromcontaminated
aquifer,despitehighconc
ofnaturalcomponent
Highcost;Extremely
contaminantspecific
Liquideliquidextractionandion
exchangeprocessinvolving
aseparatesolidphase
(Korngoldetal.,1996;
Warshawskyetal.,2002)
3.In-situchemicalfixationPb,Asandother
metalsin
agriculturalsoil.
Usingredmudandmixture
ofFeSO4,CaCO3,
KMnOandCa(H2PO4)2
LaboratoryapplicationZnandCdmetalsmaybe
mobilizedwithincrease
insoilacidity
Stabilizationofmetalsbyoxidizing
andtrappinginthestructure.
(Lombietal.,2002;
Yangetal.,2007)
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Clay�FeðIIIÞþ4SO�2 �4Clay�Fe
ðIIÞ
þ2S2O
2�
4 þH2O/2SO
2�
3
þS2O
2�
3 þ2H
þ (2)
Fruchter et al. (1997) proposed the use of alkaline solution
buffered with carbonate and bicarbonate while injecting
dithionite to reduce the effect of produced Hþ on soil pH. Fe(II)
once produced, would reduce the migrating redox-sensitive
contaminants e.g. CrO2�4 , U, Tc and some chlorinated solvents
(Fruchter et al., 1997). Yin and Allen (1999) commented that
alkaline pH and high permeability of soil are absolute necessity
for this process to work. However, dithionate was found to be
difficult to handle and generation of toxic gases may be
a hazard. Creation of a reductive treatment zone is shown
in Fig. 3.
3.1.1.2. Reduction by gaseous hydrogen sulphide. Gaseous hydrogen
sulphide (H2S gas) was tested for in-situ immobilization of chro-
mate contaminated soils by Thornton and Jackson (1994), although
the delivery of H2S gas to the contaminated zone posed to be
somewhat difficult. Nitrogen could be used as a carrier gas for the
delivery and control of H2S gas during treatment and also for
removal of any unreacted agent from the soil after treatment. The
H2S reduced Cr(VI) to Cr(III) state and precipitated it as an oxy-
hydroxide solid phase, itself being converted to sulphate as indi-
cated in Equation (3) (Thornton and Jackson, 1994). Due to very low
solubility of sulphate and Cr(III) hydroxides, secondary waste
generation was not an issue.
8CrO2�4 þ 3H2Sþ 4H2O/8CrðOHÞ3þ3SO
2�
4 (3)
Fig. 2. Creation of a reductive treatment zone, adapted from Fruchter et al. (1997).
Fig. 3. Enhanced design methodology for in-situ chemical barrier, adapted from Sevougian et al. (1994).
M.A. Hashim et al. / Journal of Environmental Management 92 (2011) 2355e2388 2359

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This gaseous treatment is conceptually similar to soil venting.
Researchers at the U. S. Department of Energy (1996) proposed
a design for the construction of injection and withdrawal wells for
in-situ gaseous treatment with H2S (Fig. 4).
Thornton and Amonette (1999) leached 90% of Cr(VI) dispersed
in soil within a column by using 100 ppm aqueous solution of H2S.
The residual Cr(VI) was found to be sequestered in unreacted grain
interiors under impermeable coatings formed during H2S treat-
ment. Thus, this technology may be experimented for chromate
contaminated aquifer treatment as well.
3.1.1.3. Reduction by using iron based technologies. Iron based
technologies for remediation of contaminated groundwater and
soil is a well documented field. The ability of iron as Fe(0) and Fe(II)
to reduce the redox sensitive elements have been demonstrated at
both laboratory scale and in field tests (CL:AIRE, 2007; Kim et al.,
2007; Ludwig et al., 2007; Puls et al., 1999). More iron based
removal processes will be discussed under the sub-sections 3.2.2.4,
3.3.1.1.1, 3.3.1.1.4, 3.3.1.2.1, 3.3.1.2.2, 3.3.1.3.2, 3.3.2.5 and 3.3.2.7.
3.1.1.3.1. Zero-valent colloidal iron (colloidal ZVI1). ZVI (Fe0) was
found to be a strong chemical reductant and was able to convert
many mobile oxidized oxyanions (e.g., CrO2�4 and TcO
�
4 ) and oxy-
cations (e.g. UO2þ2 ) into immobile forms (Blowes et al., 1995).
Colloidal ZVI of micro-nanometer particle size can be injected into
natural aquifers and this was advantageous than a treatment wall
filled with ZVI since no excavation of contaminated soil was
needed, human exposure to hazardous materials was minimum
and injection wells could be installed much deeper than trenches
(Cantrell and Kaplan, 1997; Gillham et al., 1993). Furthermore, the
treatment barrier created this way could be renewed with minimal
cost or disturbance to above-ground areas (Yin and Allen, 1999).
Manning et al. (2002) suggested that the As(III) removal was
mainly due to the spontaneous adsorption and coprecipitation of
As(III) with Fe(II) and Fe(III) oxides/hydroxides formed in-situ
during ZVI oxidation (corrosion). The oxidation of ZVI by water
and oxygen produces Fe(II) (Ponder et al., 2000):
Fe0 þ 2H2O/2Fe
2þ
þH2 þ 2OH
� (4)
Fe0 þ O2 þ 2H2O/2Fe
2þ
þ 4OH� (5)
Fe(II) further reacts to give magnetite (Fe3O4), ferrous hydroxide
(Fe(OH)2) and ferric hydroxide (Fe(OH)3) depending upon redox
conditions and pH:
6Fe2þ þ O2 þ 6H2O/2Fe3O4ðsÞ þ 12H
þ (6)
Fe2þ þ 2OH�/FeðOHÞ2ðsÞ (7)
6FeðOHÞ2ðsÞ þ O2/2Fe3O4ðsÞ þ 6H2O (8)
Fe3O4ðsÞ þ O2ðaqÞ þ 18H2O412FeðOHÞ3ðsÞ (9)
Recent research suggests that the formation of Fe2þ and H2O2 on
the corroding Fe0 surface in turn forms OH� radical (Joo et al.,
2004):
Fe0 þ O2 þ 2H
þ
/Fe2þ þH2O2 (10)
Fe2þ þH2O2/Fe
IIIOH2þ þ OH� (11)
The As (III) oxidation reaction then proceeds as:
2OH � þH3AsO3/H2AsO
�
4 þ H2Oþ H
þ (12)
Toxic intermediates may be generated as by-product from this
technique. Also, the barrier-integrity verification, effective
Fig. 4. In-situ gaseous treatment system, adapted from US DoE (1996).
1 Heavy metals are mentioned with their IUPAC symbols.
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emplacement of barriers and modelling were found to be quiet
difficult (Joo et al., 2004).
3.1.1.3.2. Removal of chromium by ferrous salts. Puls et al. (1999)
suggested the following reaction for chromate reduction and
immobilization by Fe:
Fe2þ þ CrO2�4 þ 4H2O/
�
FeX2Cr1�x
�
ðOHÞ3þ5OH
� (13)
The toxic or carcinogenic Cr(VI) was reduced to the less toxic
Cr(III) form, which readily precipitated as Cr(OH)3 or as the solid
solution FexCr1 � x(OH)3. CL:AIRE (2007) reported a case where at
the site of a former paper mill on the Delaware River, USA, the in
situ application of an acidified solution of ferrous sulphate hepta-
hydrate, via a combination of wells and trenches, reduced
concentrations of Cr(VI) in groundwater from 85,000 mg L�1 to
50 mg L�1 by reductive precipitation. Here, ferrous-ammonium
sulphate could also have been applied which would act relatively
rapidly over neutral to alkaline pHs, thus avoiding the need for
acidification.
Brown et al. (1998) suggested the following reaction of ferrous
sulphate to reduce Cr(VI) from the metal industry process effluents
as:
CrðVIÞðaqÞ þ 3FeðIIÞðaqÞ ¼ CrðIIIÞðaqÞ þ 3FeðIIIÞðaqÞ (14)
If the pH of the solution was near neutral, then the following
precipitates could be formed rapidly (Walker and Pucik-Ericksen,
2000):
CrðIIIÞ þ 3OH ¼ CrðOHÞ3 (15)
and if excess Fe was present, then the reaction will be:
ð1� xÞFeðIIIÞ þ xCrðIIIÞ þ 3OH ¼ CrxFe1�xðOHÞ3ðsolidÞ (16)
The mobile contaminants such as TcO�4 , UO
þ2
2 and MoO
�2
2 were
also thought to be suitable for precipitation by Fe0. Numerous
halogenated-hydrocarbon compounds and CrO2�4 had been repor-
ted to be removed effectively from groundwater by this mechanism
(Cantrell et al., 1995).
Seaman et al. (1999) also similarly used buffered and unbuffered
Fe(II) solutions to stabilize Cr(VI) by converting it to Cr(III) and
dichromate which were trapped in FeeAl system to prevent future
leaching. They concluded that this process of Cr(VI) binding might
be successful at lesser depth of soil.
3.1.2. Soil washing
This technique involves washing of contaminated soil by water
and other extracting agents, i.e. acid or chelating ligands added to
the water to leach out the reactive contaminants from the soil (Tuin
et al., 1987). According to Sikdar et al. (1998), two approaches are
taken for soil washing. In the first approach, soil washing as is
considered as a fractionating technique for isolating the finer
particles i.e. clay, silt, or humic substances which captivates the
contaminants in the soil. The washed oversize fraction can be used
for refill. The wash water, remain hazardous on account of the
presence of a fraction of the contaminants in it. The second
approach is based on washing the entire soil with a fluid that
extracts the contaminants from all size fractions. The in situ soil
washing and surfactant- or solvent-assisted soil washing tech-
niques use organic solvents, such as alcohols, polymers, poly-
electrolytes, chelants, inorganic acids, or surfactants depending on
site-specific circumstances. A comprehensive review on the use of
chelating agents for soil heavy metal remediation was undertaken
by Le�stan et al. (2008). Sikdar et al. (1998) stated that soil perme-
ability was an important determinant for in situ washing (soil
flushing) since low-permeability limited the transport of liquid or
vapor through the soil. An inherent part in contaminant removal by
soil washing is the use of membranes or some other technology to
segregate the contaminants from the wash liquids. Dermont et al.
(2008) reviewed the basic principles, applicability, advantages
and limitations, methods of predicting and improving the perfor-
mance of physical and chemical technologies for soil washing
practiced between the period of 1990e2007. The role of
membranes in the contaminant separationwill be discussed later in
section 3.3.2.2.
3.1.2.1. In-situ soil flushing. The flushing fluid (water or chemical
extractant solutions) is applied on the surface of the site or injected
into the contaminated zone. The resulting leachate can then be
recovered from the underlying groundwater by pump-and-treat
methods (DoD Environmental Technology Transfer Committee,
1994). In contaminated soil, metal ions remain sorbed to soil
particles of natural aquifer (Evanko and Dzombak, 1997; Yin and
Allen, 1999). The injection of dilute acids reduces the aquifer pH
to much lower values resulting in desorption of metal ions from
solid surfaces due to proton competition. Most commonly used
acids for soil washing are sulphuric acid, hydrochloric acid and
nitric acid (Smith et al., 1995). The fluids can be introduced by
surface flooding, surface sprinklers, basin infiltration systems, leach
fields, vertical or horizontal injection wells or trench infiltration
systems (USEPA, 1997).
Earlier, this technique was mostly followed for the treatment of
organic contaminants rather than metals. At two Superfund sites of
USA, Lipari Landfill in New Jersey and the United Chrome Products
site in Oregon, in-situ soil flushing was reported to be operational
(USEPA, 1995). At the United Chrome Products site, Cr levels in
groundwater was reduced frommore than 5000mg L�1 to less than
50 mg L�1 in areas of high concentration (McPhillips and Loren,
1991). Moore et al. (1993) suggested the use of solutions of
hydrochloric acid, EDTA and calcium chloride as soil flushing
agents. However, treating the washing solution for reuse can be
more difficult than the soil flushing itself (Mulligan et al., 2001).
Navarro andMartínez (2010) performed soilflushingexperiments
to dissolute metals by water from an old mining area of size 0.9 ha
contaminatedbyuncontrolleddumpingof base-metal smelting slags.
The results of the pilot-scale study showed the removal of Al
(43.1e81.1%), Co (24.5e82.4%), Cu (0e55%), Fe (0e84.7%), Mn
(66.2e85.8%), Mo (0e51.7%), Ni (0e46.4%) and Zn (0e83.4%). Few
othermetals such as As, Se, Sb, Cd and Pbweremobilized or removed
innegligible amounts from the groundwater. Geochemicalmodelling
of groundwater indicated the presence of ferrihydrite which may
have caused the mobilization of As, Sb and Se.
The technical options for soil cleanup resulting in soil wash
water and subsequent treatment options are shown in Fig. 5.
3.1.2.2. In-situ chelate flushing. Injecting chelating agents in
contaminated soil may give rise to very stable soluble metal-
echelate complexes pulling out the metals from solid phase to the
solution phase. The most frequently used chelating agents are
EDTA, citric acid and diethylene triamine pentaacetic acid (DTPA)
(Smith et al., 1995).
Peters (1999) did some detailed work on the treatability of
representative soils from a contaminated site for extracting Cu, Pb
and Zn by EDTA, citric acid, nitrilotriacetic acid (NTA), gluconate,
oxalate, Citranox�, ammonium acetate, phosphoric acid and pH-
adjusted water. NTA, being a class II carcinogen, was avoided while
EDTA and citric acid offered the greatest potential as chelating
agents for removing Cd, Cu, Pb, Zn, Fe, Cr, As, andHg simultaneously.
The overall removal of Cu, Pb and Zn after multiple-stage washing
were 98.9%, 98.9%, and 97.2%, respectively. Lim et al. (2004) assessed
the suitability of using EDTA, NTA and DTPA to cleanup Pb(II), Cd(II)
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andCr(III). Pb andCdwerequickly removedat lowdoseof the ligand.
However, Cr could not be extracted by any of the 3 ligands due to its
tendency to hydrolyze and slow ligand exchange kinetics of the
hydrolyzed Cr species. Hong et al. (2008) reported complete
extractionof Pb fromasoil contaminatedwith3300mgkg�1 of Pbby
using 100 mM EDTA within 10 min at 150 psi (10 atm) pressure.
Pociecha and Lestan (2010) also extracted 67.5% of Pb from the
contaminated soil by EDTA solution, yielding washing solutionwith
1535mgL�1 Pband33.4mMEDTA.Notably, theyusedanaluminium
anode to regenerate the EDTA by removing the extracted Pb from
EDTAsolutionat currentdensity96mAcm�2 andpH10. Thisprocess
removed 90% of Pb from the solution through the electrodeposition
on the stainless steel cathode. Di Palma et al. (2003) proposed a two-
step recovery of EDTA after washing soils contaminated with Pb or
Cu. Initial evaporation led to a reductionof extractant volumeby75%
and subsequent acidification resulting in precipitation of more than
90% of the EDTA complexes.
Blue et al. (2008) found out that among K2BDET (1,3-
benzenediamidoethanethiol), sodium dimethyldithiocarbamate
(SDTC), sodium sulfide nonahydrate, disodium thiocarbonate (STC)
and trisodium 2,4,6-trimercaptotriazine nonahydrate (TMT-55),
BDET was most effective in removing Hg. One gram of BDET could
treat 353 L of a sample containing 66 ppb Hg and 126 L of sample
containing 188 ppb of Hg (Blue et al., 2008). The reaction between
BDET and Hg is shown in equation 17.
Di Palma et al. (2005) extracted Cu by Na2-EDTA from an arti-
ficially contaminated soil and evaluated the influence of the speed
of percolation and chelating agent concentration on the removal
efficiency. At pH of 7.3, the flushing solution viz 500ml of Na2-EDTA
0.05 M solution and 100 ml of pure water at 0.792 cm h�1 extracted
up to 93.9% of the Cu. Under these operating conditions, the
competitive cations such as Ca and Fe did not get the chance to
form EDTA complexes. Recovery of EDTA up to 91.6% was achieved
by evaporating and acidifying the extracted solution after filtration.
Under alkaline conditions, about 99.5% of the extracted Cu was
recovered. Some researchers used soil washing to reduce the
amount of labile contaminant in soil before stabilizing the rest of
the contaminants by using soil stabilization techniques such as in
situ chemical fixation (Isoyama and Wada, 2007; Tokunaga et al.,
2005). It was reported by Zhang et al. (2010) that pre-washing of
contaminated soil fractions with EDTA facilitated the subsequent
chemical immobilization of Cu and Cr, while Pb and Zn were
mobilized, especially when Ca(OH)2 was added as the immobilizing
agent. The influence of soil washing by chelate on the subsequent
immobilization of heavy metals was found to be dominated by
three competitive processes viz. the removal of labile fractions, the
destabilization of less labile fractions, and chemical immobilization.
Heavy metals such as Cu, Pb and Zn were removed from few
contaminated soil samples pre-treated by conventional water
washing by using some chelating agents viz. [S,S]-
ethylenediaminedisuccinic acid (EDDS), methylglycinediacetic
acid (MGDA) and citric acid. EDDS and MGDAwere equally efficient
in removing Cu, Pb, and Zn after 10e60 min, reaching to maximum
efficiency after 10 days. After this, the biodegradable amino poly-
carboxylic acids were used as a second step resulting in the release
of most of the remaining metals (Cu, Pb and Zn) from the treated
soils after long leaching time. This indicated that a 2 step leaching
procedure with different wash liquids may be effective and time
consuming (Arwidsson et al., 2010).
The problem associated with using the chelates is that most
effective chelates such as EDTA and DTPA are not biodegradable
and may be hazardous to environment while NTA is a carcinogen.
Moreover, the chelates are expensive to use, but they can be
recovered by various separation processes (Kos and Lestan, 2004;
Fig. 5. A generic flow sheet for soil washing/flushing and treatment of the soil washing solution, adapted from Sikdar et al. (1998).
NH
O
S
NH
O
S
K
K
Hg
NH
O
S
NH
O
S
Hg
2+
(17)
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Lim et al., 2004; Luo et al., 2006; Means et al., 1980; Pociecha and
Lestan, 2010; Römkens et al., 2002).
3.1.2.3. In-situ remediation of heavy metals by selective ion exchange
methods. An important class of ion-exchange resins includes
solvent-impregnated resins (SIRs). These materials combine the
advantages of liquideliquid extraction and ion exchange involving
a separate solid phase. Korngold et al. (1996) suggested the idea to
use ion-exchange resins for tap water remediation. SIRs removed
very low concentration (>1mg L�1) of contaminants in the presence
of high concentration of microelements (e.g. calcium, magnesium,
sodium, potassium and chloride) present in water at nearly neutral
pH and the presence of other anions, all of which compete for
available sites on the SIRs. Vilensky et al. (2002) studied the feasi-
bility of commercial ion-exchange resins and synthetically prepared
type II SIRs (Duolite GT-73 and Amberlite IRC-748, produced by
Rohm andHaas) for groundwater remediation as amaterial for using
in Permeable Reactive Barriers (PRBs). In type II SIRs, extractant
molecules were bound to a functional matrix due to acidebase
interactions. A major advantage of ion-exchange resins over other
adsorbents is that they can be effectively regenerated up to 100%
efficiency (Korngold et al., 1996; Warshawsky et al., 2002).
3.1.3. In-situ chemical fixation
Lombi et al. (2002) investigated the ability of red mud (an iron
rich bauxite residue), lime and beringite (a modified aluminosili-
cate) to chemically stabilize heavy metals and metalloids in agri-
cultural soil. A 2% red mud performed as effectively as 5% beringite.
The red mud amendment decreased acid extractability of metals by
shifting metals to the iron-oxide fraction from the exchangeable
form and was much reliable.
Yang et al. (2007) tried out a new method of in-situ chemical
fixation of As through stabilizing it with FeSO4, CaCO3 and KMnO4.
The initial design of the remediation experiments was based on the
following possible reactions:
15AsO�33 þ 6MnO
�
4 þ 18H
þ
/2Mn3ðAsO4Þ2þ11AsO
3�
4 þ 9H2O
(18)
3Fe2þ þMnO�4 þ 4H
þ
/3Fe3þ þMnO2ðcÞ þ 2H2O (19)
Fe3þ þ 3H2O/FeðOHÞ3þ3H
þ (20)
4Fe3þ þ 2AsO3�4 þ 6H2O/2FeðOHÞ3þ2FeAsO4 þ 6H
þ (21)
hFeOH0 þ H3AsO4/FeH2AsO4 þ H2O (22)
hFeOH0 þ H3AsO3/FeH2AsO3 þ H2O (23)
FeSO4 was used as the major component of the fixation solutions
due to the close association of iron compounds with arsenic and the
low solubility of ferric arsenate. In two of the treatment solutions,
KMnO4 was used to oxidize any As(III) in the soil samples into the less
toxic and more stable As(V). They also tested different treatment
solutions,vizonlyFeSO4, FeSO4þKMnO4andFeSO4þCaCO3þKMnO4
onsample.Althoughsoils treatedwithKMnO4solutions showed lower
mobility of arsenic than those treated with only FeSO4 for aggressive
TCLP sequential leaching, KMnO4 treatments actually left large
portions of the soil arsenic vulnerable to environmental leaching
simulated using SPLP. Finally, treatment with the solution containing
only FeSO4 was considered optimal (Yang et al., 2007).
Pb and As in contaminated soil could be immobilized by the
addition of Ca(H2PO4)2 and FeSO4 as stabilizing agents. Singular
addition of phosphate decreased Pb leachability, but significant
mobilization and plant uptake of As was noticed. Mixtures of
Ca(H2PO4)2 and FeSO4 immobilized both Pb and As by reducing
their water solubilisation. However, the soil pH decreased from 7.8
to 5.6, mobilizing Zn and Cd (Xenidis et al., 2010).
3.2. Biological, biochemical and biosorptive treatment technologies
3.2.1. Biological activity in the sub-surface
Biological treatment methods exploit natural biological
processes that allow certain plants and micro-organisms to help in
the remediation of metals in soil and groundwater. Plant based
remediation methods for slurries of dredged material and metal
contaminated soils had been proposed since the mid-1970s
(Cunningham and Berti, 1993). A number of researchers (Barona
et al., 2001; Boopathy, 2000; Salt et al., 1995) were sceptical about
significant metal extraction capability of plants. However, Salt et al.
(1995) reported a research group in Liverpool, England, making
three grasses commercially available for the stabilization of Pb, Cu
and Znwastes. Recently, a reviewpaper focusing on the use of plants
and micro-organisms in the site restoration process have been
published (Kavamura and Esposito, 2010).The biological processes
for heavy metal remediation of groundwater or sub-surface soil
occur through a variety of mechanisms including adsorption,
oxidation and reduction reactions and methylation (Means and
Hinchee, 1994). According to Boopathy (2000), some of the exam-
ples of in-situ and ex-situ heavy metal bioremediation are land-
farming, composting, useof bioreactors, bioventingbyoxygen, using
biofilters, bioaugmentation by microbial cultures and bio-
stimulation by providing nutrients. Some of the other processes
include bioaccumulation, bioleaching and phytoremediation.
Potentiallyuseful phytoremediation technologies for remediation of
metals-contaminated sites include phytoextraction, phytostabili-
zation and rhizofiltration (Evanko and Dzombak, 1997; USDOE,
1996; Vangronsveld et al., 1995). A hyperaccumulator is defined as
a plant with the ability to yield 0.1% Cr, Co, Cu, Ni or 1% Zn, Mn in the
above-ground shoots on a dry weight basis (Evanko and Dzombak,
1997). Since metal hyperaccumulators generally produce small
quantities of biomass, they are not suitable agronomically for phy-
toremediation. Nevertheless, such plants are valuable stores of
genetic and physiologic material and data (Cunningham and Berti,
1993). In order to provide effective cleanup of contaminated soils,
it is essential to find, breed, or engineer plants that absorb, trans-
locate and tolerate levels of metals in the 0.1%e1.0% range and are
native to the area (Salt et al.,1995).Wang and Zhao (2009) evaluated
the feasibility of using biological methods for the remediation of As
contaminated soils and groundwater. Ex-situ bioleaching, bio-
stimulation such as addition of carbon sources and mineral nutri-
ents, ex-situ or in-situ biosorption, coprecipitation with biogenic
solids or sulphides and introduction of proper biosorbents ormicro-
organisms to produce active biosorbents inside the aquifer or soil
were found to be suitable techniques for this purpose. Salati et al.
(2010) reported a highly efficient technique of augmenting phytor-
emediation process by using organic fraction of municipal solid
wastes (OFMSW) to enhance heavy metal uptake from contami-
nated soil by maize shoots. High presence of dissolved organic
matter, 41.6 times greater than soil control, exhibiting ligand prop-
erties due to presence of large amount of carboxylic acids made the
process very much efficient (Salati et al., 2010).
3.2.1.1. Natural biological activity. In oxygen containing aquifers,
the aerobic bacteria was found to degrade a variety of organic
contaminants e.g. benzene, toluene and xylenes. When all the
oxygen got depleted, anaerobic bacteria, e.g. methanogens as well
as sulphate and nitrate respirating bacteria continued the
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degradation (Wilson et al., 1986). Baker (1995) observed that some
plants such as Urtica, Chenopodium, Thlaspi, Polygonum sachalase
and Alyssim possessed the capability of accumulating heavy metals
such as Cu, Pb, Cd, Ni and Zn. So these could be considered for
indirectly treating contaminated soils. To date, this field of study
used to identify the botanical population of contaminated sites and
selected some plants for either phytoextraction or phytostabiliza-
tion purposes (Li et al., 2007; Regvar et al., 2006). However, more
detailed genetic level study must be done to understand the metal
uptake capability of plants. Yong and Mulligan (2004) stated that
the natural attenuation level of different heavy metals varied from
moderate to low levels with passage of considerable amount of
time. A review paper discussing various mechanisms of natural
attenuation for both organic and inorganic contaminants in soil,
role of monitoring, use of models and protocols and case studies
had been published by Mulligan and Yong (2004).
Practically, when large tracts of land gets contaminated, then
natural vegetation such as heavy metal accumulating willow trees
can cleanup the area over a long period of time (6 to 10 years) and
also can generate economically valuable products (Bañuelos, 2006;
Sas-Nowosielska et al., 2004). Also, Brassica napus (canola) and
Raphanus sativus (radish) are shown to be effective in remediating
multi-metal contaminated soil (Marchiol et al., 2004). Hellerich
et al. (2008) evaluated the potential for natural attenuation of
Cr(VI) for sub-wetland ground water at a Cr-contaminated site in
Connecticut and concluded that the attenuation capacity could be
exceeded only with high Cr(VI) concentration and extremely long
Cr source dissolution timeframes. Based on the 1-D transport
modelling and incorporating input parameter uncertainty, they
calculated a very high probability that the Cr(VI) level will remain
under the regulatory limit after the natural attenuation process.
Both willow (Salix sp. ‘Tangoio’) and poplar (Populus sp. ‘Kawa’)
were shown to uptake B, Cr and Cu from contaminated soils (Mills
et al., 2006). Kim and Owens (2010) reviewed the potential of
phytoremediation by using biosolids in contaminated sites or
landfills. Moreno-Jiménez et al. (2011) performed a study on phy-
tostabilization of heavy metals such as As, Zn, Cu, Cd and Al in the
Guadiamar river valley at Southern Spain and identified a native
Mediterranean shrub Retama sphaerocarpa having promising
ability to phytostabilize the heavy metal contaminated soils
(Moreno-Jiménez et al., 2011). However, these processes are still
not applied for groundwater remediation.
3.2.2. Enhanced biorestoration
Many of the biorestoration processes first pumped out the
leaked contaminants as far as possible. Then the micro-flora pop-
ulation was enhanced by pumping down nutrients and oxygen
through injectionwells in the aquifer (Raymond, 1974). Sometimes,
micro-flora such as bacteria, fungi, plant growth promoting rhizo-
bacteria (PGPR) and pseudomonads were used for assisting the
plant in metal uptake (Leung et al., 2006; Wu et al., 2006). In recent
years, many more bioprocesses have been developed to remove
a variety of heavymetals through enhanced biorestoration. Some of
them have been discussed here.
3.2.2.1. Immobilization of radionuclides by micro-organ-
isms. Researchers working on U(VI) reduction in contaminated
aquifers had suggested the use of acetate as an electron donor to
stimulate the activity of dissimilatory metal-reducing micro-
organisms (Finneran et al., 2002a, 2002b). It was reported that the
Geobacteraceae species available in pure culture were capable of
U(VI) reduction (Lovley et al., 1991). In an experiment conducted by
Anderson et al. (2003), U(VI) was substantially removed within 50
days of initiation of acetate injection. All the Fe(III) got reduced to
Fe(II) thereby switching the terminal electron accepting process for
oxidation of the injected acetate from Fe(III) reduction to sulphate
reduction. Sulphate reducing species of bacteria replaced the Geo-
bacteraceae species resulting in the increase of U(VI) concentration
once again. So optimization of acetate application was suggested to
ensure long term activity of Geobacteraceae. Mouser et al. (2009)
suggested that ammonium influenced the composition of bacte-
rial community prior to acetate amendment. Rhodoferax species
predominated over Geobacter species at higher ammonium
concentration while Dechloromonas species dominated the sites
with lowest ammonium. However, Geobacter species became the
predominant at all locations once acetate was added and dissimi-
latory metal reduction was stimulated. A number of researchers
worked on the stimulation of dissimilatory metal reducing activity
by micro-organisms using carbon donor amendments to immobi-
lize U and Tc from the contaminated groundwater containing
nitrate (Cardenas et al., 2008; Istok et al., 2003; North et al., 2004).
Cardenas et al. (2008) reduced U(VI) concentration in a contami-
nated site of the U.S. Department of Energy in Oak Ridge, TN from
60 mg L�1 to 30 mg L�1 by conditioning the groundwater above
ground thereby stimulating in-situ growth of Fe(III)-reducing,
denitrifying and sulphate-reducing bacteria by injecting ethanol
every week into the sub-surface.
In-situ biobarrier was used by Michalsen et al. (2009) to
neutralize pH and remove nitrate and radionuclides from ground-
water contaminated with nitric acid, U, and Tc over a time period of
21 months. Addition of ethanol effectively promoted the growth of
a denitrifying community, increased pH from 4.7 to 6.9, promoting
the removal of 116 mM nitrate and immobilizing 94% of total U(VI).
Betaproteobacteria were found to be dominant (89%) near the
source of influent acidic groundwater, whereas members of
Gamma- and Alphaproteobacteria and Bacteroidetes increased along
the flow path with increase in pH and decrease in nitrate concen-
trations. Groudev et al. (2010) treated experimental plots heavily
contaminated with radionuclides (mainly U and Ra) and non-
ferrous heavy metals (mainly Cu, Zn, Cd and Pb) in-situ using the
indigenous soil micro-flora. The contaminants were solubilised and
removed from the top soil layers by the dual role played by the
acidophilic chemolithotrophic bacteria and diluted sulphuric acid
in the acidic soil, and various heterotrophs and soluble organics and
bicarbonate in the alkaline soil. The dissolved contaminants from
top layer either drained away as effluent or got transferred to the
deeper soil subhorizon. The sulphate-reducing bacteria inhabiting
this soil subhorizon precipitated the metals in their insoluble forms
such as uranium as uraninite, and the non-ferrous metals as the
relevant sulphides.
3.2.2.2. In-situ bioprecipitation process (ISBP). In situ bio-
precipitation (ISBP), involves immobilizing the heavy metals in
groundwater as precipitates (mainly sulphides) in the solid phase.
Carbon sources such as molasses, lactate, acetate and composts are
injected in the aquifer where they undergo fermentation and trap
the metal ions in an organic matrix. The ISBP process was investi-
gated for stabilizing heavymetals such as Cu, Zn, Cd, Ni, Co, Fe, Cr, As
and was shown to be feasible as a strategy for improving ground-
water quality (Geets et al., 2003). However, the stability of the heavy
metal precipitates in the ISBP remains to be a questionable issue.
Janssen and Temminghoff (2004) assessed the ISBP based on
bacterial sulphate reduction (BSR) with molasses as carbon source
for the immobilization of a Zn plume of concentration 180mg L�1 in
an aquifer with high Eh, lowpH, loworganicmatter content and low
sulphate concentrations. They used deep wells for substrate injec-
tion. Batch experiments revealed the necessity of adding a specific
growth medium to the groundwater to an optimal molasses
concentration range of 1e5 g L�1 without which, BSR could not be
triggered. In an in-situ pilot experiment, Zn concentrations was
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reduced from around 40 mg L�1 to below 0.01 mg L�1. The BSR
process continued for at least 5 weeks even after termination of
substrate supply (Janssen and Temminghoff, 2004).
Diels et al. (2005) observed that the interruption in the carbon
source delivery stopped the ISBP. Another noticeable point was that
while CdeZn formed stable precipitates, NieCo formed less stable
precipitates which can undergo leaching (Diels et al., 2005).
Satyawali et al. (2010) investigated the stability of Zn and Co
precipitates formed after ISBP in an artificial and a natural solid
liquid matrix. In the artificial matrix, the Zn precipitate was not
affected by redox changes, but 58% of it got mobilized with
sequential pH change. In case of the natural matrices, the stability
of metal precipitates, mainly sulphur compounds Zn and Co, was
largely affected by the applied carbon source.
3.2.2.3. Biological sulphate reduction (BSR). BSR is the process of
reduction of sulphate to sulphide, catalyzed by the activity of
sulphate-reducing bacteria (SRB) using sulphate as electron
acceptor (Gibson,1990). BSR was proved to be an effective means in
reducing heavy metal concentrations in contaminated water
(Suthersan, 1997). Moreover, metal sulphides due to their low
solubility precipitate with metal ions already present in the solu-
tion. BSR was investigated for treatment of AMD on-site in reactive
barriers (Benner et al., 1999; Blowes et al., 1995, 1998; Waybrant
et al., 1998) as well as off-site in anaerobic bioreactors (Greben
et al., 2000; Hammack and Edenborn, 1992). AMD is character-
ized by low Eh, low pH and high concentrations of sulphate, Fe and
heavy metals. The use of BSR was aimed at pH increase and
sulphate and metal removal. A wide range of electron donors such
as ethanol, lactate, hydrogen and economically favorable waste
products, pure substrates or inoculated with monocultures or
media (manure, sludge, soil) containing SRB had been reported to
be quite effective in the BSR process (Annachhatre and
Suktrakoolvait, 2001; Dvorak and Hedin, 1992; Hammack and
Edenborn, 1992; Lens et al., 2000; Prasad et al., 1999; van Houten
et al., 1994; Waybrant et al., 1998).
According to Gibert et al. (2002), biologically mediated reduc-
tion of sulphate to sulphide, accompanied with the formation of
metal sulphides occurred through the reaction sequence:
2CH2OðSÞþSO
2�
4 ðaqÞ þ2H
þ
ðaqÞ/H2SðaqÞ þ2CO2ðaqÞ þH2O (24)
M2þ
ðaqÞ þ H2SðaqÞ/MSðsÞ þ 2H
þ
ðaqÞ (25)
Here, CH2O is an organic carbon and M2þ is a divalent metal
cation. Other processes related to the pH increase and the redox
potential decrease could also precipitate metals as hydroxides and
carbonates (Gibert et al., 2002).
3.2.2.4. In-situ As removal from contaminated groundwater by
ferrous oxides and micro-organisms. High concentration of arsenic
in sub-surface aquifer may arise due to the presence of bacteria,
using As bearing minerals as a energy source, reducing insoluble
As(V) to soluble As(III). Das et al. (1994) reported As in groundwater
of West Bengal in massive scale while Camacho et al. (2011) did
a detailed study on the occurrence of As in groundwater of Mexico
and south western USA.
The micro-organisms Gallionella ferruginea and Leptothrix
ochracea were found to support biotic oxidation of iron by
Katsoyiannis and Zouboulis (2004), who performed some experi-
ments in laboratory where iron oxides and these micro-organisms
were deposited in the filter medium, offering a favorable environ-
ment for arsenic adsorption. These micro-organisms probably
oxidized As(III) to As(V), which got adsorbed in Fe(III) resulting in
overall arsenic removal of up to 95% even at high initial As
concentrations of 200 mg L�1 Leupin and Hug (2005) passed aerated
artificial ground water with high arsenic and iron concentration
through a mixture of 1.5 g iron fillings and 3e4 g quartz sand in
a vertical glass column. Fe(II) was oxidized to hydrous ferric oxides
(HFO) by dissolved oxygen while As(III) was partially oxidized and
As(V) adsorbed on the HFO. Four filtrations reduced total As below
50 mg L�1 from 500 mg L�1 without any added oxidant.
Sen Gupta et al. (2009) successfully applied this principle in the
field when he reversed the bacterial arsenic reduction process
without using any chemical, by recharging calculated volume of
aerated water (DO > 4 mg L�1) in the aquifer to create an oxidized
zone. This boosted the growth of iron oxidizing bacteria and sup-
pressed the growth of As reducing anaerobic bacteria and
promoted the growth of chemoautotrophic As oxidizing bacteria
(CAOs) over a period of six to eight weeks. Subterranean ground-
water treatment turned the underground aquifer into a natural
biochemical reactor and adsorber that oxidizes and removes As
along with Fe and Mn at an elevated redox value of groundwater
(Eh > 300 mV in the oxidation zone). The technical diagram of the
arsenic removal facility is shown in Fig. 6. The success of the process
depended on controlled precipitation of Fe on the aquifer sand so
that the precipitate could acquire a dense goethite or lepidocrocite
type structure. Controlled precipitation of Fe(III) also ensured that
it trapped As(V) as it got adsorbed on the aquifer sand and is
subsequently oxidized to form a dense and compact structure,
without affecting the permeability of the aquifer sand (Sen Gupta
et al., 2009). The method was very effective in reducing the
concentration of As below the regulatory standard of 10 mg L�1 from
initial concentrations of 250 mg L�1 and was tested extensively in
field conditions (www.insituarsenic.org). van Halem et al. (2010)
also tested a community-scale facility in Bangladesh for injection
of aerated water (w1 m3) into an anoxic aquifer with elevated iron
(0.27 mMol L�1) and arsenic (0.27 mMol L�1) concentrations with
successful outcomes. Saalfield and Bostick (2009) demonstrated
a process in laboratory, where biologically mediated redox
processes affected the mobility of As by binding it to iron oxide in
reducing aquifers through dissimilatory sulphate reduction and
secondary iron reduction processes. Incubation experiments were
conducted using As(III/V)-bearing ferrihydrite in carbonate-
Fig. 6. Technical diagram of SAR technology.
M.A. Hashim et al. / Journal of Environmental Management 92 (2011) 2355e2388 2365

Author's personal copy
buffered artificial groundwater enriched with sulphate
(0.08e10 mM) and lactate (10 mM) and inoculated with Desulfo-
vibrio vulgaris (ATCC 7757), which reduces only sulphate but not Fe
or As. Magnetite, elemental sulphur and trace Fe sulphides were
formed as the end products through sulphidization of ferrihydrite.
It was suggested that only As(III) species got released under
reducing conditions and bacterial reduction of As(V) was necessary
for As sequestration in sulphides.
3.2.3. Biosorption of heavy metals
This field of remediation technology is an emerging and ever
developing field but somewhat lacking in field application. Exper-
iments with various biosorbents showed promising results. There
are a number of advantages of biosorption over conventional
treatment methods such as low cost, minimization of chemical or
biological sludge, high efficiency, regeneration of biosorbents and
possibility of metal recovery..
3.2.3.1. Metal removal by biosurfactants. Surfactants lower the
surface tension of the liquid in which it is dissolved by virtue of its
hydrophilic and hydrophobic groups. Decrease in the surface
tension of water makes the heavy metals more available for
remediation from contaminated soils (Ron and Rosenberg, 2001).
Biosurfactants are biological surfactant compounds produced by
micro-organisms and other organisms while glycolipids or lip-
opeptides are low-molecular-weight biosurfactants.
Biologically produced surfactants e.g. surfactin, rhamnolipids
and sophorolipids could remove Cu, Zn, Cd and Ni from a heavy
metal contaminated soil (Mulligan et al., 1999a, b; Wang and
Mulligan, 2004). Mulligan and Wang (2006) used a rhamnolipid
for studying its metal removal capacity both in liquid and foam
forms. Rhamnolipid type I and type II, with surface tensions of
29 mN m�1, were found to be suitable for soil washing and heavy
metal removal (JENEIL Biosurfactant Co. LLC, 2001). The metals
were removed by complex formation with the surfactants on the
soil surface due to the lowering of the interfacial tension and hence
associating with surfactant micelles. The best removal rates, 73.2%
of the Cd and 68.1% of the Ni, were achieved by adjusting the initial
solution pH value to 10. A 11% and 15% increase in Cd and Ni
removal was observed by rhamnolipid foam than rhamnolipid
solution of same concentration (Mulligan and Wang, 2006).
As¸çı et al. (2010) also experimented with rhamnolipids for
extracting Cd(II) and Zn(II) from quartz. When 0.31 mMol kgL1
Cd(II) in quartz was treated with 25 mM rhamnolipid, 91.6% of the
sorbed Cd(II) was recovered. In case of Zn(II), approximately 87.2%
of the sorbed Zn(II) or 0.672 mMol kgL1 was extracted by using
25 mM rhamnolipid concentration. On an average, 66.5% of Zn(II)
and 30.3% of Cd(II) were released at high or saturation metal ion
loadings on quartz. This indicated that a fairly large portion of the
metal ions was irreversibly retained by quartz (AsçI et al., 2010).
3.2.3.2. Metal uptake by organisms. Prakasham et al. (1999)
demonstrated 85e90% removal of Cr by adsorption in non-living
Rhizopus arrhizus biomass at acidic pH of 2 in a stirred tank
reactor at 100 rpm and at 1:10 biomasseliquid ratio for 4 h contact
time. However, fluidized bed reactors was more efficient. At sol-
ideliquid ratio of 1:10, the Cr ion removal was observed to be 73%
for 1 h and 94% for 4 h of contact time. Braud et al. (2006) revealed
that Pseudomonas aeruginosa and Pseudomonas fluorescens could
extract Pb from its carbonates to an exchangeable fraction although
the Pb bound to FeeMn oxides, organic matter and in the residual
fractions remained stable. Abou-Shanab et al. (2006) reported a 15
times increase of extractable Ni with Microbacterium arabinoga-
lactanolyticum depending on the initial Ni concentration in the soil.
Rangsayatorn et al. (2002) studied a cyanobacteria Spirulina
(Arthrospira) platensis TISTR 8217 for removing low level Cd
(>100 mg L�1) from water. Metal sorption amounting to 78% was
completed within 5 min at pH 7 and in a temperature range of
0e50OC. Earlier, Spirulinawas used both for industrial and domestic
wastewater treatments.
Pandey et al. (2008) found that Calotropis procera, a wild
perennial plant had high uptake capacity of Cd(II) at pH 5.0 and 8.0.
The adsorption equilibrium of �90% removal was attained within
5 min, irrespective of the Cd ion concentration. Pandey et al. (2008)
deducted that at lower adsorbate concentration, monolayer
adsorption or Langmuir isotherm was followed while multilayer
adsorption or Freundlich isotherm was followed at higher
concentrations. Other cations, viz. Zn(II), As(III), Fe(II) and Ni(II) also
interfered in the adsorption process when their concentration was
higher than the equimolar ratio. The involvement of hydroxyl
(eOH), alkanes (eCH), nitrite (eNO2) and carboxyl group (eCOO)
chelates in metal binding was indicated by the FTIR analysis. The
presence of common ions viz. Ca2þ, Mg2þ, Cl�, SO¼4 , PO
3�
4 did not
significantly interfere with metal uptake properties even at higher
concentrations. The complete desorption of the Cd was achieved by
0.1 M H2SO4 and 0.1 M HCl.
In an innovative approach, Kim et al. (2009) used single-
stranded DNA aptamers to remove As from Vietnamese ground-
water. One of the As binding DNA aptamers, Ars-3, was found to
have the highest affinity to both As(V) and As(III) with dissociation
constants (Kd) of 4.95 � 0.31 nM and 7.05 � 0.91 nM, respectively.
Different As concentrations ranging from 28.1 to 739.2 mg L�1 were
completely removed after 5 min of incubation with Ars-3.
Srivastava et al. (2011) isolated fifteen fungal strains from soils
of West Bengal, India to test the biological removal of As. The Tri-
choderma sp., sterile mycelial strain, Neocosmospora sp. and
Rhizopus sp. fungal strains were found to be most effective in bio-
logical uptake of As from soil, the removal rate ranging between
10.92 and 65.81% depending on pH. More research can be done
with these strains to apply them in As contaminated aquifers after
properly creating aerobic environment.
Therefore, a number of living organisms, micro and macro,
either up took metals in their body or increased the extracted the
heavy metals from their bound condition. These organisms can be
applied in suitable condition for aquifer remediation but more
research is required to suit them to field conditions.
3.2.3.3. Biosorption of heavy metals by cellulosic materials and
agricultural wastes. Unmodified cellulose had been reported to
possess low heavy metal adsorption capacity and also variable
physical stability, forcing the researchers to carry out chemical
modification of cellulose to achieve adequate structural durability
and higher adsorption capacity for heavy metal ions (Kamel et al.,
2006). O’Connell et al. (2008) reviewed a range of modified cellu-
lose materials mainly produced by esterification, etherification,
halogenation and oxidation. Some modified cellulose materials
used by a number of researchers over years to remove various
heavy metals have been listed in Table 3. Sud et al. (2008) also
reviewed cellulosic agricultural wastematerials for their capacity of
significant metal biosorption. The functional groups such as acet-
amido, alcoholic, carbonyl, phenolic, amido, amino and sulphydryl
groups present in agricultural waste biomass formed metal
complexes or chelates with heavy metal ions. The biosorption
process occurred by chemisorption, complexation, adsorption on
surface, diffusion through pores and ion exchange mechanisms.
Han et al. (2009) described the use of electrospinning process to
fabricate oxidized cellulose (OC) by introducing a more porous
structure inside the OC matrix, thereby increasing its capacity to
chelate with metal ions. These OCs were shown to be uptaking Th
and U from groundwater. The optimum pH conditions for heavy
M.A. Hashim et al. / Journal of Environmental Management 92 (2011) 2355e23882366

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metal binding on modified cellulose materials were mostly
observed to occur in the pH range of 4.0e6.0. Most of the adsorp-
tion mechanisms between the modified cellulose adsorbents and
heavy metals were characterized either by the Langmuir model or
in a lesser number of cases by the Freundlich model of adsorp-
tion.Hasan et al. (2000) indicated rubber-wood ash to be a suitable
adsorbent for the Ni(II) cation from dilute solution. The adsorption
reaction could be described as first order reversible reaction and
the equilibrium was reached within 120 min. The optimum effi-
ciency of adsorbing 0.492 mMol g�1 of Ni(II) was obtained at pH 5
and 30 �C temperature.. Above pH value of 5.5, precipitate of
Ni(OH)2 was formed, thus reducing number of available free Ni
ions.
Tabakci et al. (2007) studied the sorption properties of adsor-
bents prepared from cellulose grafted with calix[4]arene polymers
(CGC[4]P-1 and CGC[4]P-2) for adsorbing some heavy metal cations
such as Co2þ, Ni2þ, Cu2þ, Cd2þ, Hg2þ and Pb2þ and dichromate
anions, Cr2O
2�
7 =HCr2O
�
7 . CGC[4]P-2 exhibited excellent sorption
properties for heavy metal ions and dichromate anions at pH 1.5
but CGC[4]P-1 was not much effective (Tabakci et al., 2007).
Grape stalk, a by-product of wine production, was utilized by
Martínez et al. (2006) for sorption of lead and cadmium from
aqueous solutions. Maximum sorption capacities was found to be
0.241 and 0.248 mMol g�1 for Pb(II) and Cd(II), respectively, at
a pH value of 5.5. However, HCl or EDTA solutions desorbed 100%
of Pb and 65% of Cd from the grape stalks. In addition to ion
exchange process, other mechanisms, such as surface complexa-
tion and electrostatic interactions, were thought to be involved in
the adsorption process. Amin et al. (2006) used untreated rice
husk for complete removal of both As(III) and As(V) from aqueous
solution at an initial As concentration of 100 mg L�1 in 6 g of rice
husk at a pH range of 6.5 to 6.0. Desorption in the range of 71e96%
was observed when rice husk was treated with 1 M of KOH. Sahu
et al. (2009b) used activated rice husk to achievemaximum Pb and
BOD reduction of 77.15% and 19.05%, respectively in a three phase
modified multi-stage bubble column reactor (MMBCR). Walnut
hull was studied for Cr(VI) adsorption from solution and was
found to be pH-dependent, reaching 97.3% at pH 1.0. The efficiency
increased with temperature and with increasing initial Cr(VI)
concentration up to 240e480 mg L�1, and decreased with
increasing adsorbent concentration ranging from 1.0 to 5.0 g L�1.
Sodium chloride, as supporting electrolyte in the medium,
induced a negative effect on the efficiency. So, besides the modi-
fied cellulose materials, the simple walnut hull demonstrated
good performance and can be tested out in field application (Wang
et al., 2009).
Munagapati et al. (2010) studied the kinetics, equilibrium and
thermodynamics of adsorption of Cu(II), Cd(II) and Pb(II) from
aqueous solution on Acacia leucocephala bark powder. The bio-
sorption capacity followed the order Pb(II) > Cd(II) > Cu(II) at
optimum conditions of pH 4.0, 5.0 and 6.0 at biosorbent dosage of
6 g L�1. The biosorption process best fitted the pseudo-second-
order kinetic model, was exothermic and spontaneous. It was
concluded that A. leucocephala bark powder could be used as a low
cost, effective biosorbent for the removal of Cu(II), Cd(II) and Pb(II)
ions from aqueous solution (Munagapati et al., 2010).The cellulose
materials and agricultural wastes therefore show promising
results depending on pH and polymerization reactions. Never-
theless, more extensive research is required for their field
application.
The biological, biochemical and biosorption treatment processes
are summarized in Table 4.
3.3. Physico-chemical treatment technologies
The techniques are dependent upon physical processes or
activities such as civil construction of barriers, physical adsorption
or absorption, mass transfer as well as harnessed chemical or
biochemical processes are discussed here. Most of the times, two or
more processes are coupled together to deal with the contamina-
tion problem. The physico-chemical treatment processes are
summarized in Table 5.
3.3.1. Permeable reactive barriers (PRB)
USEPA (1989) defined Permeable Reactive Barrier (PRB) as ‘an
emplacement of reactive media in the sub-surface designed to
intercept a contaminated plume, provide a flow path through the
reactive media and transform the contaminant(s) into environ-
mentally acceptable forms to attain remediation concentration
goals downgradient of the barrier’. The concept behind PRB is
that a permanent, semi permanent or replaceable reactive media
is placed in the sub-surface across the flow path of a plume of
contaminated groundwater which must move through it under
its natural gradient, thereby creating a passive treatment system.
Treatment walls remove contaminants from groundwater by
degrading, transforming, precipitating, adsorbing or adsorbing
the target solutes as the water flows through permeable reactive
trenches (Vidic and Pohland, 1996). PRBs are designed to be
more permeable than the surrounding aquifer materials so that
water can readily flow through it maintaining groundwater
hydrogeology while contaminants are treated (Yin and Allen,
1999). The reactive cell is generally constructed approximately
0.6 m above the water table and 0.3 m keyed into the aquitard,
deeper in case of the funnel-and-gate system. Such construction
would prevent contaminants from flowing either on top or
bottom of the reactive cell. Gavaskar et al. (1998) summarized
four possible arrangements for construction of the reactive cell
(Fig. 7). Adequate site characterization, bench-scale column
testing, and hydrogeologic modelling are essential for designing
and constructing PRBs (Gavaskar, 1999). Lee et al. (2009)
proposed a design-specific site exploration approach for PRB
designing called quantitatively directed exploration (QDE),
employing three spatially related matrices such as covariance of
input parameters, sensitivity of model outputs and covariance of
model outputs to identify the ideal location for the PRB (Lee
et al., 2009).
Table 3
Heavy metal removal by modified cellulose materials.
Modified cellulose
materials
Metals
removed
Operating
concentration
(mg g�1)
References
Peanut hulls Cu(II) 65.6 (Periasamy and
Namasivayam, 1996)
Tree bark Cu(II) 21.6 (Gaballah and
Kilbertus, 1998)
Sugar cane bagasse Pb(II) 133.6 (Peternele et al., 1999)
Orange peel Ni(II) 80.0 (Ajmal et al., 2000)
P. chrysosprium Cu(II) 26.5 (Say et al., 2001)
Pb(II) 85.9
Cd(II) 27.8
P. versicolor Ni(II) 57.0 (Dilek et al., 2002)
Hazelnut shell Ni(II) 10.1 (Demirbas et al., 2002)
Trametes versicolor Cu(II) 116.9 (Bayramoglu et al., 2003)
Pb(II) 229.9
Zn(II) 109.2
Sugar beet pulp Pb(II) 73.8 (Reddad et al., 2003)
Grape stalk waste Ni(II) 10.6 (Villaescusa et al., 2004)
Cu(II) 10.1
Pb(II) 49.9
M.A. Hashim et al. / Journal of Environmental Management 92 (2011) 2355e2388 2367

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Table4
Biological/biochemicaltreatmenttechnologies:comparativeoverview.
TechnologyScopeConditionsand
modesofapplication
AdvantageDisadvantageMechanismandprocessSelectedreferences
1.Biologicalactivityin
thesub-surface
Cr,Co,Cd,Ni,
Zn,Pb,Cu
In-situcultureofaerobic
bacteriaandplanting
oftrees.Onlyin
shallowsub-surface
Verylittlecost;Applicable
tolargetractoflandover
longtime
Notsuitableforaquifer
remediation;Veryslow
process;Nomodelling
canbedone
Oxidation,precipitation,
bioaccumulation
(Baker,1995;Salatietal.,2010;
Wilsonetal.,1986;Yong
andMulligan,2004)
2.Enhancedbiorestoration
2.1Immobilizationof
radionuclidesby
micro-organisms
U,Ra,TcInjectingcarbondonor
e.g.acetatetosupport
Geobactorspeciesofbacteria
inbiobarrier
Noharmfulbyproducts
areproduced
Acetateinjectiontobe
optimizedtopreventgrowth
ofSRB
Reduction,agglomeration,
absorptionofU(IV)into
sediments
(Andersonetal.,2003;Finneran
etal.,2002a;Finneranetal.,
2002b;Mouseretal.,2009)
2.2ISBPCu,Zn,Cd,Ni,Co,
Fe,Cr,As
InjectionofCarbonsource
(e.g.molasses)inaquifer
bydeepwells
Cheapcarbonsources
available
Heavymetalppts(e.g.Ni
andCo)mayremobilizewith
changingsoilpH
Fermentationofcarbonsources
insideaquiferandtrappingof
heavymetalsinorganicmatrix
(Dielsetal.,2005;Geets
etal.,2003;Janssenand
Temminghoff,2004;Satyawali
etal.,2010)
2.3BSRDivalentmetal
cations
Injectionofelectrondonors
andinoculatingthesoilor
aquiferwithbacterialcultures.
On-siteremediationof
AMD;Offsiteusein
bioreactors;Canbe
usedinPRBs
Reactionratelimitedand
requiressufficientresidencetime.
Reductionofsulphatetometal
sulphideppts,catalyzedbythe
activityofSRB.
(DvorakandHedin,1992;
Gibertetal.,2002;Hammack
andEdenborn,1992;Hammack
etal.,1994;Waybrantetal.,1998)
2.4In-situAsremovalby
ferrousoxidesand
micro-organisms
As,Fe,MnIn-situoxidationofFe(II)
andAs(II)byinjectingaerated
waterinaquiferbyboosting
aerobicAsoxidizingbacteria
Nochemicalsused;
Nowasteproduced;Low
operatingcost
Regularinjectionofaeratedwater
neededtomaintainoxidationzone
OxidationofFe(II)andAs(III)
byelevatingEhandboosting
microbialgrowthandthen
co-precipitatingAs,FeandMn
(Camachoetal.,2011;
KatsoyiannisandZouboulis,
2004;LeupinandHug,2005;
SenGuptaetal.,2009)
3.Biosorptionofheavymetals
3.1BiosurfactantsCd,Zn,NiExperimentaluseinlaboratory
withrhamnolipidssolution
andfoam
Highmetalretention
capacity
Nottestedinfield.Foamis
supposedtobemoresuitable,
buttransportationtodeep
aquiferscanbetough
Bioadsortionthroughmetal
complexformingwith
surfactantsduetolowering
ofinterfacialtension
(AsçIetal.,2010;Mulligan
andWang,2006;Ronand
Rosenberg,2001)
3.2UptakebyorganismsCd,Cr,Zn,As,Fe,NiLaboratoryexperimentsdone
withinpH2e8toremoveCd
andCrfromaqueoussolution.
Zn,As,Fe,Nicanalso
getadsorbed,anions
donotinterfere
DesorptionoftheHeavymetals
underhighacidiccondition
Bacteria,fungus,plantsand
DNAaptamersuptookmetals
incellcytoplasm,or
stabilizatedthem
(Kimetal.,2009;Pandey
etal.,2008;Prakasham
etal.,1999;Srivastava
etal.,2011)
3.3Cellulosicmaterialsand
agriculturalwastes
Pb,Ni,Cu,Cd,ZnLabexptwitharangeofmodified
cellulosematerials
Lowcostcellulose
materials.Largerange
ofheavymetalscan
betreated
NosignificantfieldstudydoneBioadsorptionofheavymetals
inmodifiedcellulosestructure
atpHrange4e6
(Hanetal.,2009;Hasan
etal.,2000;Kameletal.,
2006;Sahuetal.,2009a;
Sudetal.,2008;Tabakci
etal.,2007)
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