An overview of permeable reactive barriers for in situ sustainable groundwater remediation

Chemosphere 111 (2014) 243–259

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Chemosphere journal homepage: www.elsevier.com/locate/chemosphere

Review

An overview of permeable reactive barriers for in situ sustainable groundwater remediation Franklin Obiri-Nyarko a, S. Johana Grajales-Mesa b,⇑, Grzegorz Malina b a b

Hydrogeotechnika Sp z oo, Department of Environmental Protection and Cartography, ul. Sciegiennego 262A, 25-112, Kielce, Poland AGH University of Science and Technology, Department of Hydrogeology and Engineering Geology, Al. Mickiewicza 30, 30-059, Kraków, Poland

h i g h l i g h t s  Permeable reactive barriers (PRBs) are a technology for remediation of groundwater.  There is a wide spectrum of contaminant that can be treated with PRBs.  Zero valent iron still remains the most often applied material for PRBs.  A key aspect for the design of the technology is an adequate site characterization.  Long term performance of the barrier is still not well understood.

a r t i c l e

i n f o

Article history: Received 9 October 2013 Received in revised form 14 March 2014 Accepted 22 March 2014 Available online 8 May 2014 Handling Editor: Klaus Kümmerer Keywords: Permeable reactive barriers In situ remediation Groundwater contamination Reactive materials

a b s t r a c t Permeable reactive barriers (PRBs) are one of the innovative technologies widely accepted as an alternative to the ‘pump and treat’ (P&T) for sustainable in situ remediation of contaminated groundwater. The concept of the technology involves the emplacement of a permeable barrier containing reactive materials across the flow path of the contaminated groundwater to intercept and treat the contaminants as the plume flows through it under the influence of the natural hydraulic gradient. Since the invention of PRBs in the early 1990s, a variety of materials has been employed to remove contaminants including heavy metals, chlorinated solvents, aromatic hydrocarbons, and pesticides. Contaminant removal is usually accomplished via processes such as adsorption, precipitation, denitrification and biodegradation. Despite wide acknowledgment, there are still unresolved issues about long term-performance of PRBs, which have somewhat affected their acceptability and full-scale implementation. The current paper presents an overview of the PRB technology, which includes the state of art, the merits and limitations, the reactive media used so far, and the mechanisms employed to transform or immobilize contaminants. The paper also looks at the design, construction and the long-term performance of PRBs. Ó 2014 Elsevier Ltd. All rights reserved.

Contents 1. 2. 3.

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Advances in the PRB technology . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Reactive media used in PRBs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.1. Zero valent iron (ZVI) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abbreviations: PRBs, permeable reactive barriers; PRmBs, permeable reactive multi-barriers; P&T, pump and treat; ZVI, zero valent iron; PCE, tetrachlorocthylene (or perchloroethcne); TCE, trichloroethylcne (trichloroethene); DCE, dichloroethylcnc; VC, vinyl chloride; HAH, halogenated aliphatic hydrocarbons; BTEX, benzene, toluene, ethylbenzene, xylene; DDT, 1,1,1-trichloro-2,2-bis(p-chlorophenyl)ethane; DDD, 2,2-bis(p-chlorophenyl)-1,1-dichloroethylene; DDE, 1,1-dichloro-2,2-bis(p-chlorophenyl)ethane; ORC, oxygen releasing compound; TCA, 1,1,1-trichloroethane; PCB, polychlorinated biphenyl; PAHs, polycyclic aromatic hydrocarbons; COD, chemical oxygen demand; AOX, adsorbable organic halogens; DOC, dissolved organic carbon; AC, activated carbon; GAC, granular activated carbon; SMZ, ssurfactant modified zeolites; OC, organic carbon; AMD, acid mine drainage; TRM, transformed red mud; AFO, amorphous ferric oxide; BOF, basic oxygen furnace; MTBE, methyl tertiary-butyl ether; SRB, sulfate reducing bacterial; O&M, operation and maintenance; PV, present value; CBA, cost-benefit analysis. ⇑ Corresponding author. Tel.: +48 126174789; fax: +48 12 6172427. E-mail address: [email protected] (S.J. Grajales-Mesa). http://dx.doi.org/10.1016/j.chemosphere.2014.03.112 0045-6535/Ó 2014 Elsevier Ltd. All rights reserved.

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3.2. 3.3. 3.4. 3.5. 3.6. 3.7. 3.8. 3.9.

4.

5. 6.

Activated carbons (AC) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Zeolites. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Lime and other alkaline materials . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Apatite . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Sodium dithionite . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Transformed red mud (TRM) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Oxides . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Materials for biobarriers . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.9.1. Materials for aerobic biodegradation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.9.2. Materials for anaerobic biodegradation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.10. Miscellaneous . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.11. Combination of reactive materials . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . PRB design . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.1. PRB configuration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.2. Construction methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4.3. Costs. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . PRB longevity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Concluding remarks . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Acknowledgments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

1. Introduction The quality of groundwater resources globally has been under serious threat due to their exposure to a broad spectrum of contaminants emanating from a variety of sources including agricultural systems, industries and mines (Tase, 1992; Schipper et al., 2010; Wiafe et al., 2013; Rodak et al., 2014). The conventional technology used to remediate contaminated groundwater has been the ‘pump-and-treat’ (P&T) systems. However, clean-up goals have hardly been met with this technique. Thus the past three decades have seen a lot of research directed toward the development of novel sustainable groundwater remediation techniques (Henderson and Demond, 2007). Permeable reactive barriers (PRBs) are one of the innovative technologies being used for in situ remediation of contaminated groundwater (Tratnyek, 2002; USEPA, 2002). The PRB concept involves the emplacement of a reactive media perpendicular to the potential trajectory of the contaminated groundwater. As the contamination plume passively migrates through the media under the influence of the natural hydraulic gradient, the contaminants in the plume react with the media leading to either their transformation to less harmful compounds or fixation to the reactive materials (Powell et al., 1998; Carey et al., 2002; Skinner and Schutte, 2006). The decontamination of the groundwater, which usually occurs within and (or) downgradient of the barrier, depending on the type of reactive media used, is accomplished via destructive and/or non-destructive processes (Carey et al., 2002; Wilkin and Puls, 2003; Puls, 2006; Henderson and Demond, 2007; Chen et al., 2011a). Since the serendipitous invention of the PRB technology in the early 1990s, its ability to remove groundwater contaminants has been extensively investigated. The results of some of these investigations are phenomenal, thereby presenting the PRB technology as a suitable alternative to the conventional P&T method (Korte, 2001; Carey et al., 2002; Wilkin and Puls, 2003; Puls, 2006; Skinner and Schutte, 2006; Henderson and Demond, 2007; Chen et al., 2011a). Despite this, there is still a dearth of empirical evidence regarding the long-term performance of PRBs as most of the investigations are laboratory based (Warner and Sorel, 2002). There have also been reports on pollution swapping in some types of PRBs (Schipper et al., 2010), which have necessitated their improvements to enable the treatment of a broad spectrum of contaminants, and thereby expand their remit. To date, however,

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PRBs are still considered a promising technology in the field of contaminant remediation, with a record of over 200 field installations since its inception (ITRC, 2011). There are many published documents and reviews on PRBs; however, majority of them have focused on specific issues related to barriers with zero valent iron (ZVI) as a reactive material (Scherer et al., 2000; Korte, 2001; Henderson and Demond, 2007; Noubactep, 2010). Recently, Schipper et al. (2010) and Careghini et al. (2013) presented a review on bioreactors and biobarriers, respectively, which are a type of PRB. This paper is focusing on a contaminated groundwater/hydrogeology audience, although PRBs currently exist in many forms, e.g. denitrifying bioreactors which are used extensively in groundwater and tile drainage agricultural systems. It, therefore, presents an overview of PRBs including the current state of the technology; the merits and limitations; the reactive media used so far and the mechanisms employed to transform or immobilize contaminants. It also looks at the design, construction and the long-term performance of PRBs.

2. Advances in the PRB technology The first field PRB studies were conducted at the Canadian Forces Base, Borden (O’Hannesin and Gillham, 1998). This has since been followed by a spate of investigations. According to Bone (2012), a total of 624 publications on PRBs were made between 1999 and 2009. Approximately 40% of these were laboratory-based investigations, with field studies accounting for ca. 32%. A comparison of the latter with the 16% estimated by Scherer et al. (2000) indicates that the number of field publications doubled in 10 years. The PRB technology was first used to remediate groundwater contaminated with chlorinated solvents such as trichloroethylene (TCE), the three isomers of DCE (1,2-cis-, 1,2-trans- and 1,1-DCE) and vinyl chloride (VC) in the early stages. After proving to be effective in the treatment of these contaminants, its application was extended to include other contaminants. In Table 2, a list of the contaminants that have been treated with PRBs so far has been presented (Blowes et al., 1998; Conca et al., 2002; Köber et al., 2002; USEPA, 2002). The contaminants include halogenated aliphatic hydrocarbons, metals, metalloids, radionuclides, pesticides, petroleum hydrocarbons, and nutrients emanating from agricultural systems. Bone (2012), however, reported that 38% of the contaminants treated with PRBs at field scale from 1994 to 2009 were halogenated aliphatic hydrocarbons (HAH). Since the beginning of

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the technology, two main configuration types of PRBs for field applications have been used; they are: the funnel-and-gate and the continuous gate designs (Powell et al., 1998; Thiruvenkatachari et al., 2008). The former is made up of two structures: the funnel, which consists of cut-off walls meant to converge the plume to the treatment zone, and the reactive gate in which the treatment occurs. Even though this configuration type is expensive to construct compared to the latter, it allows for pockets of plumes widely distributed to be captured for treatment (Starr and Cherry, 1994). The continuous gate configuration, on the other hand, involves the placement of the treatment barrier across the entire contaminant plume path. The merits of this configuration type are that it is easy to construct, less expensive and has little effect on the groundwater flow. However, this design is only suitable for plumes with narrow widths (Gavaskar et al., 2000). Recent review of literature shows that several modifications have been made to the two original PRB designs. Notable reasons for such modifications are the desire to use a design that suits the geology of the site, the possibility to treat a wide range of contaminants, and the discovery of relatively cheap barrier materials (Yang et al., 1995; Korte, 2001; Phillips, 2010; Hosseini et al., 2011). In the early stages of the PRB technology, the majority of the barriers were filled with zero valent iron (ZVI), which could treat a limited number of contaminants. The term biobarrier was thus introduced which allowed the use of organic materials for the removal of contaminants that were not amenable to the most frequently applied reactive material, i.e. ZVI in PRBs (Yang et al., 1995). With the introduction of biobarriers, it became possible to use inexpensive organic materials/substrates to be used as filling materials to enhance the growth and activities of autochthonous or inoculated microorganisms to facilitate the degradation of contaminants (Vesela et al., 2006; Wilson et al., 2001; Yerushalmi et al., 1999). (See also section on biobarriers). Denitrifying bioreactors, which is a collective term for denitrification walls (Robertson and Cherry, 1995), denitrification beds (Robertson et al., 2009) and denitrification layers (Robertson and Cherry, 1995; Schipper and McGill, 2008) were also introduced for the removal of nitrate from groundwater and agricultural systems. Nitrogen is one of the major plant nutrients in addition to phosphorus and potassium. However, it is also a major environmental problem due to its ability to leach (mainly in the form of nitrate) into groundwater and surface waters, thereby affecting their quality. Denitrifying bioreactors also utilize solid carbon substrates to achieve contaminant removal, and initially they were used mainly for the removal of nitrate (conversion of nitrate to nitrogen gas). However, recent studies have shown that they can be used to treat other contaminants including pathogens, pharmaceutical compounds, pesticides and phosphates in agricultural drainage, and perchlorate (Schipper et al., 2010). The early stages of the technology also saw most PRB applications focusing on the use of a single barrier (usually filled with a single reactive material) (ITRC, 2011). These barriers were mainly used for contamination plumes containing one contaminant or contaminants of similar nature (e.g. heavy-metals). However, for most sites where the plume is a mixture of contaminants with different physical, chemical and thermodynamic properties (e.g. heavy-metals, BTEX and TCE) such barriers were shown to be ineffective (Köber et al., 2002). Apart from the inability of such barriers to attenuate multi-contaminant plumes, most researchers also had to deal with the issue of pollution swapping, which is defined as the inadvertent generation or release of new potentially hazardous contaminants during the removal of other contaminants (Stevens and Quinton, 2009; Healy et al., 2012). Lai et al. (2006) reported the generation of 1,2-dichloroethane and dichloromethane during the reductive dechlorination of chlorinated aliphatic hydrocarbons

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(CAHs) by ZVI. Borden (2008) also noted the accumulation of cisDCE and vinyl chloride (VC) during the removal of TCE and PCE. A number of studies have also reported on the pollution swapping phenomena during the use of denitrifying bioreactors for the removal of nitrate emanating from agricultural systems. Nitrate is usually swapped with contaminants such as ammonium, or greenhouse gases such as nitrous oxide (N2O), CO2 and CH4. The review of Schipper et al. (2010) is replete with such studies. Initially, the issue of pollution swapping was not considered in the design of PRBs. However, in the wake of the increased global environmental consciousness and desire for the use of sustainable remediation, pollution swapping has become a major subject for discussion. Consequently, the multi-barrier concept was introduced to make PRBs a more sustainable technology and to broaden their field of application. A multi-barrier system is broadly defined as PRBs consisting of two or more barriers filled with the same or different reactive materials (sequenced multi-barrier). However, it can also be interpreted to mean a single barrier filled with different reactive materials (a mixed or multi-treatment barrier). In the case of the former, contaminants are removed sequentially while in the latter their removal occurs simultaneously. Though the concept of multi-barriers is relatively recent, it has also received considerable attention (Köber et al., 2002; Birke et al., 2007). Some multi-barriers have been designed for the removal of mainly organic contaminants. Morkin et al. (2000) used two barriers for the treatment of a mixed organic plume of chlorinated ethenes (DCE and VC) and BTEX within a funnel-and-gate system. The first barrier consisted of granular iron for reductive dechlorination, while the second was a biosparged zone to facilitate biological degradation of the contaminants, mainly BTEX and the by-products of the chlorinated ethenes. The results showed well performance of the barriers in attenuating the contaminants. BTEX was shown to be strongly retarded by the granular iron, whereas DCE and VC were observed to be removed in both barriers. A total mass removal of about 99% for both DCE and VC of which biodegradation accounted for 65% and 30%, respectively, was observed. A multipermeable reactive barrier with iron and biologically active tire as a bio-barrier was also used by Lee et al. (2007) to remove PCE, TCE and organic matter. Köber et al. (2002) used a mixed-treatment barrier by combining Fe0 and granular activated carbon (GAC) as fillings for the removal of complex contaminant mixtures in groundwater. The contamination plume consisted of monochlorbenzene (MCB) and TCE. Three different systems were evaluated. The first system consisted of GAC, the second was a mixture of Fe0 and GAC, and the third consisted of a sequential combination of a Fe0 column followed by a GAC column. Best results were obtained when Fe0 and GAC were in sequence. The authors indicated that the separation minimized or prevented a decrease of the sorption capacity of the GAC. Conca et al. (2002) also used a four-component permeable reactive barrier to remediate groundwater contaminated with mainly inorganic contaminants (radionuclides, metals and nitrates). Van-Nooten et al. (2010) also used a sequential setup combining different reactive materials and removal processes for the treatment of a complex mixture of organic and inorganic contaminants consisting of ammonium, adsorbable organic halogens (AOX), chemical oxygen demand (COD) and the by-products of nitrification and denitrification. For a comprehensive work on sequenced reactive barriers for remediation of groundwater contaminated with a variety of contaminants, the reader is referred to Fiorenza et al. (2000). These studies generally demonstrate the effectiveness of multi-barriers in attenuating multi-contaminant plumes. Permeable reactive interceptors (PRI) which are a novel approach of multi-PRBs have also been introduced recently as a panacea to the problem of pollution swapping in denitrifying bio-

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reactors. There are not many published studies to verify the efficiency of this treatment system. However, the work of Fenton et al. (2014) suggests that, indeed, PRI can be a viable solution to pollution swapping in agricultural systems. Apart from costs, the complexities in the design of such multitreatment systems and difficulties in monitoring them holistically have, however, been acknowledged. Problems such as antagonistic effects imposed by one contaminant (or a treatment barrier) on another during the remediation of multi-contaminant plumes have been reported (Johnston et al., 1996; Chen et al., 2011b). A number of approaches, however, have and are being used to address such issues to make PRBs a more sustainable technology. Laboratory batch and column studies have frequently been performed to obtain information about various geochemical and microbiological phenomena in such treatment systems. Batch tests are quick to perform; nevertheless column experiments provide dynamic flow conditions which closely approximate those expected in a PRB system in field deployments. However, they both provide valuable information that can be used to forfend or alleviate, for example, the antagonistic effects in the multi-barrier systems (Gavaskar et al., 2000; Henderson and Demond, 2007). Geochemical modeling tools like MINTEQA2 (Allison et al., 1991) and PHREEQC (Parkhurst and Appelo, 1999) are also being used to study and predict the possible mineral phases that may be formed, and changes in geochemical parameters such as pH and Eh from the reactions between the reactive materials, the contaminants and the groundwater constituents, to inform on the choice of the reactive materials and sequencing of the barriers to preclude the antagonistic or inhibitory effects observed in multibarriers. The issue of pollution swapping has not been given much attention in this review; however, we note herein that addressing it will require the assessment of the total pollution swapping capacity, which will depend on the type of contaminants present. Moreover, it will require knowledge of the mechanisms as well as the factors/ parameters controlling the attenuation of the contaminants. Fenton et al. (2014) also noted that the design of a sustainable PRB system as a measure of dealing the pollution swapping will depend on national legislation as there are different emphases on soluble or gaseous losses in different geographical locations. Healy et al. (2012) highlighted certain parameters that need to be quantified for the assessment of the total pollution swapping and the associated risk in terms of GHG emissions and release of dissolved contaminants in denitrifying bioreactors. Such parameters have mostly been quantified using a combination of physicochemical analytical techniques. The field of microbial ecology is progressing rapidly with the development of new techniques that can be applied to advance understanding of various biotransforma-

tion processes within PRBs. Recently, Fenton et al. (2014) suggested the use of microbial and molecular fingerprinting as an in situ cost-effective tool to assess nutrient and gas balances in PRIs. Such tools are rapid and can allow for simultaneous analyses of multiple samples, and thus can be useful in elucidating the pollution swapping capacity of such systems (Rastogi and Sani, 2011). Generally, some studies have also noted the failure in terms of capturing contaminants or abatement of the effectiveness of PRBs with time due to some biogeochemical processes arising from reactive media-contaminant-aquifer interactions within or up-gradient of the barriers. Clogging of barrier due to accumulation of carbonate and sulfate precipitates, loss of reactivity and decreases in hydraulic residence time, gas production and the subsequent reduction of permeability, and competition for or loss of reactive sites due to corrosion, fouling or precipitation, are among the documented causes of PRB failure (Gavaskar, 1999; Mackenzie et al., 1999; Phillips et al., 2000; Korte, 2001; Gu et al., 2002; Roberts et al., 2002; Kamolpornwijit et al., 2003; Liang et al., 2003; Wilkin and Puls, 2003; Borden, 2007; Zolla et al., 2009). Henderson and Demond (2007) did a critical review on the longterm performance of PRBs with ZVI by using graphical and statistical methods to identify the most critical factor(s) affecting the longevity of barriers. Their study revealed that improper hydraulic characterization was the principal factor that led to most PRB failures. To ensure that the target contaminants are being captured and to minimize the possibility of failure, preliminary site assessment and detailed site characterization are required. These assessments help to identify possible constraints that may affect the installation of the barrier or its performance after installation, and thus indicate whether or not the PRB will be suitable for the particular site. Mostly, the assessment includes characterization of the hydrogeology of the site, the contaminants, and the groundwater geochemistry (Gavaskar et al., 2000). Geo-technical assessment is also important as it helps to indicate whether there are any structures that may obstruct the use of heavy equipment aboveground during barrier installation, or complex geological formations below ground surface that may pose difficulties in the construction and/ or subsequent use of the barrier. Contaminant characterization and groundwater chemistry are both needed for the selection of a suitable media, and the design of the PRB. They are also needed to assess or predict possible bio-geochemical changes that may affect the longevity of the barrier (Morrison and Spangler, 1992; Gavaskar et al., 2000). The advantages and limitations of the PRB technology are summarized in Table 1 (USEPA, 1997; Powell et al., 1998; National Technical University of Athens [NTUA], 2000; Carey et al., 2002;

Table 1 Advantages and limitations of the PRB technology. Advantages

Limitations

(a) Relatively cheap passive technology, i.e.: inexpensive but effective reactive materials can be used; low energy cost; little or no disposal costs for treated wastes and; relatively low maintenance and monitoring costs with the exception of initial cost of installation (b) Allows for treatment of multiple contamination plumes since more than one barrier can be used

(a) Only contaminants flowing in the direction of the barrier can be treated

(c) Ability to treat a wide range of contaminants (d) The aboveground of the contaminated site can be put to profitable use while treatment is on-going (e) No cross-media contamination since contaminants are not brought to the surface (f) Requires occasional monitoring to ensure that barriers are functioning properly (g) Obviates the handling and loss of large volumes of groundwater

(b) Requires proper characterization of the site, aquifer, hydrogeological conditions and accurate delineation of the contaminant plume prior to barrier installation (c) Restricted to plumes no deeper than 20 m beneath the ground surface (d) Limited field data concerning longevity of barriers (e) Below ground structures (e.g. services, foundations) may present problems in construction and performance (f) Reactive media may have to be removed or be replaced during operation (g) May require long-term monitoring, particularly in the case of persistent contaminants or very slow groundwater flow

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Warner and Sorel, 2002; Puls, 2006; Henderson and Demond, 2007; Jirasko, 2012).

3. Reactive media used in PRBs Contaminant removal in PRBs occurs mainly in the zone of reactive media, and to some extent down-gradient of the barrier, depending on the type of media used. Some of the reactive media remove contaminants through physical contact while others work by altering the biogeochemical processes in the treatment zone, thus providing conditions conducive for contaminant immobilization or (bio)degradation. Thus, the main objective of the PRB, irrespective of the design used, is to bring the contaminants into the reactive zone, where they can be destroyed or immobilized. There is currently a wide range of reactive materials available (Table 2). The most common of them is ZVI. Others, including activated carbon (AC), zeolites, peat, saw dust, oxygen releasing compounds (ORC), etc. have also been used and evaluated. The limitations with most of these materials, however are that they are expensive, difficult to access or effective for only certain groups of contaminants (Köber et al., 2002). Thus efforts are still being made to find more suitable and cost-effective materials to broaden the spectrum of contaminants that can be treated with PRBs (Golab et al., 2006;

247

Di-Nardo et al., 2010). The choice of the reactive material is generally influenced by: (i) the type of contaminants to be treated (i.e., organic and/or inorganic), their concentrations, and the mechanisms needed for their removal (e.g. biodegradation, sorption or precipitation) (McGovern et al., 2002); (ii) the hydrogeological and biogeochemical conditions of the aquifer; (iii) the environmental/health impacts; (iv) mechanical stability (capacity of the material to preserve its hydraulic conductivity and reactivity over time), and (v) the availability and cost of the material. It is of utmost importance that novel materials intended to be used as reactive media are assessed for their removal efficiency and kinetics (reactivity), longevity, hydraulic conductivity (Ahmad et al., 2007; Fallico et al., 2010), and their potential to release toxic byproducts when they interact with the contaminants (USEPA, 1998; Muegge, 2008). The hydraulic performance of the material is usually a function of the grain size. A suitable grain size ensures a trade-off between reactivity and permeability. Usually the materials are mixed with sand to achieve the suitable permeability (NTUA, 2000; Jirasko, 2012). These properties are often determined via laboratory batch and column studies (Muegge, 2008). The mechanisms/processes responsible for the removal of the contaminants, though diverse, can be put into two main categories: (i) destructive biotic or abiotic processes such as e.g. biodegradation or reductive dechlorination that transform contaminants

Table 2 Summary of major contaminants and reactive materials in PRB applications. Contaminants

Reactive materials

References

PCE, TCE, DCE, TCA, VC

ZVI, GAC, H2/Paladium, SMZ, Zn0, Mulch, Sand/wood chips, Tire rubber

AFCEE (2008), Benner et al. (2002), Chen et al. (2011b), Fennelly and Roberts (1998), Gavaskar et al. (2000), Henry et al. (2003), Lee et al. (2007), Li et al. (1999), NTUA (2000), Öztürk et al. (2012), Tobiszewski and Namies´nik (2012), Vogan et al. (1999)

BTEX

GAC, ORC, Compost, Peat, Saw dust, Ground rubber, Leaf litter, ZVI, , SMZ, H2/Paladium, Cyclophane I, II

Aivalioti et al. (2008), Chen et al. (2011b), Guerin et al. (2002), Kershaw et al. (1997), Kwon et al. (2011), NTUA (2000), Powell et al., 1998, Ranck et al. (2005), Simantiraki et al. (2012), USEPA (2002), Vidic and Pohland (1996), Yeh et al. (2010)

Phenol

GAC, SMZ

USEPA (2002), Vidic and Pohland (1996)

Nitrobenzene

ZVI

USEPA (1995, 2002), Vidic and Pohland (1996)

PCB, PAHs, DDE, DDT, DDD

GAC, ZVI

Katz et al. (2006), NTUA (2000), Sayles et al. (1997), USEPA (2002), Yang et al. (2010a)

Ni, Cu, Zn, Pb, Cd, Fe, As, Cr, Hg, etc.

Limestone, Zeolites, OC, ZVI, Bone char, Apatite (Clinoptilolite), Bauxite, Activated alumina, Fly ash, Atomized slag Peat moss, Compost Sodium-dithionite, Bentonite, Ferric oxyhydroxides, TRM, Chitosan

Benner et al. (1999, 2002), Cappai et al. (2012), Cheng et al. (1997), Chung et al. (2007), Conca et al. (2002), Gavaskar et al. (2000), Genç-Fuhrman et al. (2005), Kriegman-King and Reinhard (1992), Ludwig et al. (2009), Manios et al. (2003), NTUA (2000), Peric´ et al. (2004), Seelsaen et al. (2006), USEPA (1995, 2002), Vidic and Pohland (1996), Wilkin and McNeil (2003)

ZVI, Bone char phosphate, Hydroxyapatite, Limestone AFO, BOF, Zeolite Apatite II, Pecan shells, Lignite, Coal, Titanium oxide, Ferric chloride, Ferric nitrate

Arey et al. (1999), Chung et al. (2007), Farrell et al. (1999), Gavaskar et al. (2000), ITRC (2005), Misaelides et al. (1995), Morrison and Spangler (1992), Naftz et al. (2002), NTUA (2000), Skinner and Schutte (2006), USEPA (2002), Vidic and Pohland (1996)

+ COD, AOX, NO 3 , NH4

ZVI, Saw dust, Pecan shell O2/Clinoptilolite, Wood chips, Apatite II, Compost, Polystyrene, Wheat straw, Softwood and sand, Maize cobs, zeolite

Cameron and Schipper (2010), Cheng et al. (1997), Choe et al. (2004), Chung et al. (2007), Gavaskar et al. (2000), Gibert et al. (2008), ITRC (2005), Liao et al. (2003), Nguyen and Tanner (1998), Phillips and Love (1998), Robertson et al. (2000), Skinner and Schutte (2006), USEPA (2002), Vidic and Pohland (1996), Westerhoff and James (2003)

PO3 4

OC, ZVI, Iron oxide, Peat/sand, Limestone,Ocher

Fenton et al. (2009), Gavaskar et al. (2000), James et al. (1992), NTUA (2000), Van-Nooten et al. (2010), Vidic and Pohland (1996)

SO2 4

Limestone, SMZ OC, Mushroom-compost TRM, ZVI

Benner et al. (1999, 2002), Cappai et al. (2012), Conca et al. (2002), Gavaskar et al. (2000), Lapointe et al. (2006), NTUA (2000), Peric´ et al. (2004), Skinner and Schutte (2006), USEPA (2002), Vidic and Pohland (1996)

Cl

Geza rock, ZVI, Zeolite, AC

Fronczyk et al. (2010)

U, Tc, Mo, Se,

137

Cs,

90

Sr, Pu, Am

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into entirely new non-toxic products; (ii) non-destructive processes e.g. adsorption, cation exchange, surface complexation, and precipitation that sequester or immobilize contaminants and consequently reduce their concentrations in the groundwater. The occurrence of some of these mechanisms can be easily recognized. Others, such as anaerobic reductive dechlorination, anaerobic oxidation and abiotic reductive dechlorination can occur concurrently, thus making it difficult to clearly distinguish between them (Air Force Center for Environmental Excellence [AFCEE], 2008). Generally, materials that utilize adsorptive processes to achieve groundwater decontamination must demonstrate high hydrophobicity and insolubility. Materials that utilize precipitation to immobilize contaminants must be able to adjust geochemical parameters such as pH and redox of the aquifer to minimize the solubility of the contaminants (Panturu et al., 2009). The removal processes utilized in PRBs are generally dependent on the media used, the target contaminants as well as the biogeochemical conditions prevailing in the aquifer. Knowledge of these processes is necessary for enhancing the removal efficiencies of the materials and also for the engineering design of PRBs (USEPA, 1998; Gu et al., 2002). In this review, we have looked mainly at reactive media which have been extensively used in PRBs and touched tangentially on minor and novel ones. We have also attempted to include some of the mechanisms employed by these materials for the contaminant removal. However, it is not our goal here to provide in-depth elucidation of these mechanisms as some have been well treated by other authors (Tratnyek et al., 2003; Henderson and Demond, 2007; Noubactep, 2010; ITRC, 2005, 2011). 3.1. Zero valent iron (ZVI) ZVI is the most frequently utilized media both in laboratory studies and field applications. The report of ITRC (2005) indicated that by 2004 more than 60% of the PRBs installed worldwide were iron-based. It has been used mostly in the form of chips, jet blasting media, iron foams and pellets; particulates and powders, or as Fe-filler material for concrete. ZVI has a high reduction potential of -440 mV (Simon and Meggyes, 2000). As a result, it acts primarily as a reductant in most systems. The rate of contaminant removal mostly depends on the grain size and specific surface area of the iron and the prevailing geochemical conditions of the aquifer. Typically, ZVI with grain sizes ranging from 0.25 to 2 mm and surface area of 0.5 to 1.5 m2 g1 have been used (Powell et al., 1998; Mackenzie et al., 1999; Tratnyek et al., 2003; ITRC, 2011). Contaminants known to be treated by ZVI include chlorinated hydrocarbons (e.g. TCE, PCE, VC and DCE) (Gallinati et al., 1995; Orth and Gillham, 1996; O’Hannesin and Gillham, 1998; Vogan et al., 1999; ITRC, 2005; Da-Silva et al., 2007; Henderson and Demond, 2007), heavy metals, metalloids and radionuclides including, but not limited to, Cr, Cd, Pb, Cu, U, As (Blowes et al., 1998, 2000; Powell et al., 1998; Scherer et al., 2000; Su and Puls, 2001, 2004; Wilkin and McNeil, 2003; Sun et al., 2006; Yang et al., 2007; Li and Zhang, 2007; Wilkin et al., 2008; Ludwig et al., 2009), nutrients (e.g. nitrates, phosphates and sulfates) (Cheng et al., 1997; Mayer, 1999; Benner et al., 2002; Liao et al., 2003; Westerhoff and James, 2003; Da-Silva et al., 2007), Cl- (Fronczyk et al., 2010) and pesticides (e.g. DDT, DDE and DDD) (Sayles et al., 1997; Yang et al., 2010b). ZVI has been extensively used for the removal of chlorinated solvents such as TCE and PCE. These contaminants mostly act as oxidants and thus are readily reduced by ZVI via transfer of electrons in anaerobic conditions. The mechanism for their degradation has been dealt with elsewhere (Matheson and Tratnyek, 1994; Roberts et al., 1996; Vogan et al., 1999; Scherer et al., 2000; Phillips et al., 2010).

Metals, metalloids and radionuclides have also been removed through processes including, adsorption (Fiedor et al., 1998; Farrell et al., 1999), surface complexation, reductive precipitation (Morrison et al., 2001; Gu et al., 2002; Naftz et al., 2002; Zhang et al., 2005) and co-precipitation (Scherer et al., 2000; Lien and Wilkin, 2005; Kanel et al., 2006; Henderson and Demond, 2007; Noubactep, 2010). Generally, the removal of these contaminants is influenced by pH, dissolved organic carbon (DOC) and redox conditions. Wilkin and McNeil (2003) performed reversibility test and indicated that the ability of ZVI to retain metals is largely dependent on its acid neutralizing ability rather than the redox conditions. The presence of inorganic species in the aquifer may also compete with contaminants for ZVI reactive sites and affect contaminant removal. Further discussion on ZVI can be found under the section on PRB longevity. 3.2. Activated carbons (AC) Activated carbons are carbonaceous materials possessing chemically heterogeneous surfaces with high adsorption capacity. They contain a substantial amount of organic carbon with phenolic and carboxylic groups defining their surface chemistry. Owing to these properties, they have been applied widely for the removal of contaminants such as phenols, BTEX, PCE and TCE. A number of reports have indicated that heavy metals may also be removed by AC (USEPA, 1998; Scherer et al., 2000; Köber et al., 2002; Nakagawa et al., 2003; Panturu et al., 2009; Di-Nardo et al., 2010). AC, mostly in the granular form (GAC), was one of the materials commonly used in the early stages of the PRB technology (Bone, 2012). Contaminants removal occurs mainly through sorption, which is strongly influenced by the solution pH. Peng et al. (2003) noted that high pH causes ionization of the carboxylic and hydroxylic groups on the surface of the AC. This increases the interaction of water molecules with the AC surface consequently decreasing the adsorption of particularly hydrophobic contaminants. Liu and Pinto (1997) also reported a decrease in phenol adsorption when the pH was reduced from 6.3 to 3.07. The efficacy of AC may also be affected by groundwater constituents. Cornelissen et al. (2005) indicated that natural organic matter may compete with contaminants for the binding sites and consequently reduce the magnitude of their sorption. It will thus require adequate characterization of the aquifer and careful management if AC is considered as a barrier material. 3.3. Zeolites Zeolites are aluminosilicate minerals with cage-like structures that have high cation-exchange capacity (200–400 meq/100 g) and large surface area (up to 145 m2 g1) (ITRC, 2011). The surface of zeolites is characterized by different pore structures which allow for selective adsorption of contaminants. The high ion-exchange capacities of zeolites are attributed to their permanent negative charges, which develop from isomorphic substitution. These charges are not pH dependent and are usually balanced by alkali and alkaline earth metals such as Na+, Ca2+, K+ and Mg2+ (Peric´ et al., 2004). Natural zeolites generally have larger particle size which makes them suitable for use as a reactive media. However, their low organic carbon content makes them somewhat unsuitable for sorption of organic compounds. Examples of natural zeolites include: analcime, chabzite, clinoptilolite, heulandite, natrolite, philipsite, mordenite and stilbite (Coombs et al., 1997). Among these, clinoptilolite has been extensively used for the removal of mainly cationic contaminants such as Pb, Cu and Cd (Peric´ et al., 2004). The surface chemistry of zeolites can be modified with surfactants (e.g. hexadecyltrimethylammonium, HDTMA, (C16H33)(CH3)3N+))

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to create hydrophobic and hydrophilic segments on the zeolite (Scherer et al., 2000). However, this does not affect the cation exchange capacity, thus making possible the removal of a wide 3 range of contaminants including heavy metals, NH+4, NO 3 , PO4 , radionuclides, PCE and BTEX, with efficiencies ranging between 80% and 100% (Bowman et al., 1995; Vidic and Pohland, 1996; Li et al., 1999; Scherer et al., 2000; Park et al., 2002; USEPA, 2002; Lee et al., 2003; Ranck et al., 2005; Katz et al., 2006). Bowman et al. (1995) reported that surface modified zeolite (SMZ) demonstrates strong affinity for CrO2 4 and PCE as both contaminants were completely removed in pilot-scale investigations. This was explained by the positively charged bilayer and the increased organic carbon content that resulted from the surfactant modification. In the column investigations of Ranck et al. (2005), the applied SMZ removed BTEX from wastewater to below a detection limit. Further studies conducted in the field showed an even greater sorption of BTEX than predicted from the column studies. The mean Kd determined in these studies ranged from 18.3 L kg1 for benzene to 95.0 L kg1 for p- and m-xylene. It was also indicated that SMZ could be regenerated via air sparging without compromising its sorption capacity. Other reports have also revealed that sulfates and phenols can be removed with zeolites (Vidic and Pohland, 1996; USEPA, 2002). For natural zeolites, removal of cationic contaminants reportedly occurs via sorption directly to the residual negative charges on the zeolite or through ion-exchange with Na+, Ca2+, K+ and Mg2+ (Peric´ et al., 2004). However, Misaelides et al. (1995) reported that the impounding of contaminants from groundwater by natural zeolites (clinoptilolite, heulandite) could also be due to precipitation. The use of zeolite may be affected by the pH of the groundwa2 ter and its constituents such as Ca, Mg, Na, SO2 and 4 , CO3 dissolved organic matter. Groundwater constituents may compete with the contaminants for binding sites, and thereby affect the removal efficiency of the zeolite. Application of zeolites would require pH values between 3 and 11 (Bowman et al., 1995; Misaelides et al., 1995). 3.4. Lime and other alkaline materials Limestone (calcite, aragonite), hydrated or slaked lime (Ca(OH)2), dolomite (CaMg(CO3)2) and quick lime (CaO) are alkaline materials used particularly for the treatment of groundwater contaminated with acid mine drainage (AMD) (Dixon et al., 1989; Triay et al., 1989; Conca et al., 2002). AMD is characteristically acidic and it usually contains panoply of metals because most metals become more soluble when the pH is low. The application of these materials adjusts the pH of the groundwater (mostly raising the pH) to a point where the solubility of metals is reduced to allow their precipitation. These materials have mostly been applied to remove heavy metals; however, Baker et al. (1998) noted that a mixture of crushed lime stone and sand can be used to precipitate phosphate. Since different metals have different solubility product constants, it will require careful control of the pH to achieve optimal removal of a mixture of metals. A major issue with these materials is that, the precipitates formed can clog the barrier and affect its hydraulic performance. The application of materials such as limestone can also lead to environmental problems such as directly increasing the hardness of the groundwater and CO2 content, which is a greenhouse gas. McDonald and Grandt (1981) also pointed out that large doses of alkaline materials may be required to raise and maintain pH values above 6.5 for optimal metal removal. Pang et al. (2009) also indicated that increasing the pH beyond the minimum solubility point for metals can result in their remobilization. Therefore, since these materials work by pH adjustment, it is imperative to ensure that the ideal pH conditions are maintained in the barriers during their use.

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3.5. Apatite Apatites are phosphate minerals with partial solubility in water. Van Wazer (1974) noted that of all the known phosphate minerals, apatites are the most abundant. Natural apatite minerals contain mainly calcium and phosphorus, with varying fractions of hydroxide (hydroxyapatite [Ca5(PO4)3OH]), fluoride (fluorapatite [Ca5(PO4)3F]), and chloride (chlorapatite [Ca5(PO4)3Cl]). Among them, hydroxyapatite (HAP) has been extensively studied (Liao et al., 2010). Apatites are generally stable over a wide range of geological conditions, and are characterized by a net negative charge at neutral and alkaline pH (ITRC, 2005). Apatites may remove contaminants via mechanisms such as (i) direct sorption of cationic contaminants to their negative surface charges or via ion-exchange, (ii) precipitation as phosphates, carbonates, oxides, and hydroxides and (iii) surface adsorption or incorporation into the internal structure of the apatite (Takeuchi et al., 1988; Ma et al., 1993; Bostick, 2003; Fuller et al., 2003; Zhu et al., 2010). Lower et al. (1998) investigated the ability of hydroxyapatite to remove Pb and observed good performance by the hydroxyapatite. The loss of Pb from solution resulted partially from simultaneous dissolution of HAP and precipitation of Pb5(PO4)3OH, hydroxypyromorphite. Similar studies were performed by Xu and Schwartz (1994) and drew similar conclusions regarding the mechanism of Pb immobilization. Fuller et al. (2003) also investigated the feasibility of using a commercial hydroxyapatite (Ca5(PO4)3 (OH)) to immobilize dissolved U(VI). They noted that the interaction between the apatite and U resulted in the formation of autunite (Ca(UO2)2(PO4)210H2O)), which is of relatively low solubility. Ma et al. (1993) also studied the removal of Pb by hydroxyapatite. Pb concentrations were reduced from 100 lg L1 to <1 lg L within 10 min, demonstrating a removal efficiency of almost 100%. The report of ITRC (2011) showed that apatites can also be used for the removal of AsO3 4 . However, attempts to remove other  2 2 2 anions such as VO3 4 , MoO4 , SeO3 , CrO4 , TcO4 have been less successful (Xu and Schwartz, 1994; Lower et al., 1998; Bostick, 2003; Fuller et al., 2003). The removal of heavy metals by apatite in PRBs is affected by the pH. Low pH is necessary to dissolve the apatite to release the phosphate for the precipitation of metals. Ma et al. (1993) noted that the removal of Pb via precipitation was accelerated in aqueous solutions with pH values ranging from 3 to 7. However, at extremely low pH, the formation of hydroxypyromorphite may be inhibited. Scherer et al. (2000) also indicated that high carbonate concentrations will increase solution pH and subsequently inhibit metal precipitation by apatite. Reactions involving apatites are reversible; therefore, the contaminant can be released again into groundwater when geochemical conditions that favor such a process are present. 3.6. Sodium dithionite Sodium dithionite (NaS2O4) is an inexpensive, environmentally benign and a readily available material. It transforms naturally occurring ferric oxides (Fe3+) to the ferrous (Fe2+) form, thus creating a treatment zone that is capable of reducing or precipitating contaminants via transfer of electrons (Paul, 2002). Theoretically, it can be used to reduce contaminants such as halogenated hydrocarbons and immobilize various inorganic contaminants such as chromate and uranium. However, investigations so far have shown that it is only effective for easily reducible compounds such as Cr6+ and UO2+ 2 (Gavaskar et al., 2000; Ott, 2000; Paul, 2002). Paul (2002) reported a decrease in Cr(VI) concentrations from 2.2 to 5 mg L1 to non-detectable levels within 60 h of injecting NaS2O4. Eqs. (1) and (2) are mechanism that explain U immobilization by sodium dithionite in Clap et al. (2012)

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2 3þ 2þ S2 O2 þ 4Hþ 4 ðaqÞ þ 2FeðsÞ þ 2H2 O $ 2SO3 þ 2Fe

ð1Þ

2þ 3þ UO2þ 2 þ 2FeðsÞ $ UO2ðsÞ þ FeðsÞ

ð2Þ

Being the only known liquid material, its use will require special handling compared to other barriers with solid media (Ott, 2000). In addition, this treatment can only be applied in areas where ferric oxides abound. 3.7. Transformed red mud (TRM) The use of TRM for the removal of groundwater contaminants has also been investigated. TRM is alkaline in nature (pH of 8– 10.5), and it is formed when brine is mixed with waste materials (red mud) generated during the production of alumina from bauxite. TRM contains mainly impregnated hydrated alumina and oxides of iron. The impregnation is done with alkaline minerals such as calcium aluminosilicates, magnesium hydroxides, calcium hydroxides, and hydroxycarbonates. TRM has been applied as an alternative to lime for treating acid sulfate soils or AMD (Munro et al., 2004). Lapointe et al. (2006) recorded a removal efficiency of 99% for Fe, Cu, Zn, Ni, and Pb. Cappai et al. (2012) also demonstrated that TRM is capable of treating Cr. Due to their fine texture, they have to be mixed with sand or gravel to increase the permeability when used in PRBs. 3.8. Oxides The use of various naturally occurring oxides such as basic oxygen furnace (BOF) oxide, alumina (Al2O3), amorphous ferric oxide (AFO) (Fe(OH)3), goethite (a-FeOOH), magnetite (a-Fe2O3) and hematite (Fe3O4) and hydrous titanium oxide (Ti(OH)4) for groundwater remediation has also received much attention (Morrison and Spangler, 1992). Characteristically, they: (i) have large surface areas; (ii) demonstrate strong binding affinity for many dissolved metals; and (iii) show fast metal adsorption kinetics. The contaminant removal by these materials usually occurs via electrostatic attraction due to development of surfaces charges on the oxides, or by surface complexation mechanisms. At low pH, there is increased surface protonation, which tends to favor sorption of anions as shown, for example, in Eq. (3), whereas cation adsorption is favored at high pH due to deprotonation – Eq. (4) (Scherer et al., 2000).

FeOOH þ Hþ ! FeOOHþ2

ð3Þ

FeOOH þ OH ! FeOO þ H2 O

ð4Þ 3+

Contaminants treated with oxides include As and As5+ 2+ (Giménez et al., 2007), UO2 (Giammar and Hering, 1999), PO4 6+ (Baker et al., 1998; Herbert et al., 2000), MoO2 (Morrison 4 , Cr and Spangler, 1992) and other cationic metals. Giménez et al. (2007) evaluated the effectiveness of three sorbents (hematite, goethite and magnetite) in removing As3+ and As5+. Results indicated that sorption of both contaminants increased at lower pH and vice versa for all the three sorbents. However, goethite and magnetite tended to have higher sorptive affinity than hematite for both contaminants. Giammar and Hering (1999) also indicated that goethite is an effective sorbent for uranyl ions. The use of ochre to remove groundwater contaminants has also been investigated, though not extensively. Fenton et al. (2009) demonstrated that ochre from abandoned metal mine could be used to remove phosphorus from an untreated AMD with high removal efficiencies. However, they observed that heavy metals became more mobile alongside and

concluded that this may hinder the use of ocher in phosphorus sequestration technologies. Singh et al. (1999) noted that a number of factors including iron (Fe) mineralogy, degree and rate of oxidation, water content of sediments, age of deposits, pH, Fe supply, and concentrations of alkali and sulfate (SO4) associated cations influence the physical and chemical characteristics of ochre. The use of oxides is generally dependent on geochemical conditions such as pH, Eh and the groundwater constituents such as sulfates and carbonates, which compete for surface sites. Thus, it is imperative to keep these parameters and analytes under control (Scherer et al., 2000). 3.9. Materials for biobarriers Biobarriers are a type of PRBs filled with materials that stimulate or enhance microorganisms to degrade contaminants aerobically or anaerobically. Since biobarriers rely on microorganisms to accomplish contaminant degradation, it is important that the right microbial populations are present. Usually, the microbes needed to initiate such processes are often ubiquitous, particularly in the upper layers of the aquifer (Herbert et al., 2000; Di-Nardo et al., 2010; ITRC, 2011). 3.9.1. Materials for aerobic biodegradation There are a number of organic contaminants that are efficiently removed via aerobic biodegradation. Such contaminants including petroleum hydrocarbons such as BTEX and MTBE are mostly reduced in nature. An important requirement for the occurrence of such process is the presence of terminal electron acceptors (TEAs) to receive electrons transferred from the contaminants during their degradation. The TEAs commonly available in the envi2 ronment include O2, NO 3 , CO2, sulfate (SO4 ), manganese (Mn2+), and Ferric iron (Fe3+). Molecular oxygen is, however, usually preferred to other TEAs because it yields more energy to the microbes, thus accelerating the degradation rates. In the absence of O2, the other TEAs are utilized often in a sequential manner (Weelink et al., 2010). The need for external source of oxygen to be supplied for this metabolic process is necessitated by the fact that most aquifer conditions are anaerobic and the available anaerobic TEAs are usually not sufficient to achieve complete degradation within the desired time frame (Johnson et al., 2003). Several methods including oxygen diffusers, air sparging and application of solid oxygen-and nitrate-releasing compounds (ORC) such as calcium peroxide (CaO2), magnesium peroxide (MgO2), hydrogen peroxide (H2O2), NaNO3 and sodium percarbonate (Na2CO31.5H2O2) have been used to release oxygen or create an aerobically active zone in the subsurface. In PRBs, however, ORCs have mostly been applied due to their relatively low cost (Scherer et al., 2000; Simon and Meggyes, 2000; ITRC, 2011). Bianchi-Mosquera et al. (1994) used cement-impregnated ORC as a source of oxygen for the removal of benzene and toluene. Yeh et al. (2010) applied CaO2 for the removal of BTEX in column experiments. They concluded that the CaO2 adequately met the oxygen demand of bacteria and can thus be applied in PRBs. Mackay et al. (2002) also evaluated in situ aerobic biological treatment of MTBE using the PRB at two sites: Vandenberg Air Force Base (VAFB) and Port Hueneme Naval Construction Battalion Center (PHNCBC). Results of their investigation indicated that the introduction of oxygen stimulated the microbes native to both sites and resulted in complete degradation of MTBE. A major problem, however, with ORCs is that, they are usually encapsulated with cement of high pH to control their release rates. Bianchi-Mosquera et al. (1994) noted that such cement encapsulated materials can be effective for over hundreds of days. However, the high pH of the cement, can also affect the activities of the microorganisms that degrade the contaminants. The use of

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ORCs may also not be practical at sites with high levels of native organic carbon and reduced dissolved iron as these may consume the oxygen (Borden et al., 1997). ORCs applied in the form of nitrates can also be problematic when their dissolution/release rate is not synchronized with their utilization rate. It can result in a plethora of nitrate in the groundwater (Kao and Borden, 1997). 3.9.2. Materials for anaerobic biodegradation Unlike the contaminants that are biodegraded aerobically, the contaminants that are biodegraded anaerobically are oxidized in nature and thus are preferably degraded by their reduction. The materials applied in PRBs therefore serve as the source of electrons whereas the contaminants serve as the electron acceptors for the metabolic process. A number of contaminants are known to be biodegraded anaerobically; however, we focus mainly on nitrates, sulfates and chlorinated solvents and the materials used for their biodegradation in PRBs. 3.9.2.1. NO 3 . Several low-cost organic materials including alfalfa, leaves, peat, sewage sludge, manure, sawdust, wood waste and compost have been applied for the removal of nitrate from groundwater (Schipper et al., 2010; ITRC, 2011). Vogan (1993) used alfalfa in denitrifying layers to promote nitrate removal from septic effluent. Robertson and Cherry (1995) reported a reduction of nitrate concentration from 90 mg L1 to below standard level of 10 mg L1 when saw dust was used as a source of carbon. Robertson et al. (2008) also investigated the long-term removal rates of nitrate in a pilot scale PRB consisting of sand and wood particle (sawdust) mixtures. The media retained 50–100% NO3-N during the experimental period of 15 years. These findings demonstrated the longterm effectiveness of denitrifying PRBs, which is in contrary to the widely held opinion that biobarriers require periodic replacement. The study, however, confirmed the effects of ancillary reactions on the effectiveness of reactive media in PRBs. Nitrate is removed by denitrifying bacteria such as Paracoccus denitrificans and other microorganisms belonging to the Pseudomonas group in the presence of organic substrates as shown in Eqs. (5) and (6) (Robertson and Cherry, 1995; Simon and Meggyes, 2000):

5CH2 OðsÞ þ 4NO3 ! 2N2 þ 5HCO3 þ 2H2 O þ Hþ

ð5Þ

5CH2 OðsÞ þ 4NO3 ! 2N2 þ 5CO2 þ 3H2 O þ OH

ð6Þ

The preferred end-product of nitrate reduction is N2 due to its inert nature resulting from the strong triple covalent bond that binds the molecule. However, reports of nitrous oxide, NH+4 (dissimilatory nitrate reduction to NH+4), and NO 2 have also been made (Burgin and Hamilton, 2007). Intermediate products such as nitrous oxide, CH4, CO2 are greenhouse gases, which are of a major environmental concern. On the other hand, the complete reduction of nitrate to N2 reportedly can have some consequences on the hydraulic performance of the barriers. Henderson and Demond (2011), for instance, observed that column containing 100 mg L1 nitrate experienced the high porosity loss, of approximately two orders of magnitude over the course of 200 pore volumes due to gas production. In the study of Istok et al. (2007), however, the effects of gas production on water saturation and hydraulic conductivity were considered to be relatively minor. A possible way of reducing the effect of N2 is to design the barrier in a manner not to retain the N2 in the PRB system, while providing other possibilities to trap the greenhouse gases. As suggested before, multi-barrier systems such as PRIs can be an effective way of completely removing these swapped pollutants. 3.9.2.2. SO2 4 and heavy metals. Sulfate reducing bacteria (SRB) utilize the organic substrates under anaerobic conditions to reduce

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sulfates to sufides (Benner et al., 1999; Waybrant et al., 2002). The process increases the pH and releases other products such as dissolved phosphates, which may be utilized by the microbes as nutrients. Under such relatively high pH conditions, the sulfides can subsequently react with heavy metals resulting in their precipitation as metal sulfides. Waybrant et al. (2002) performed column experiments using model mine-drainage water to assess the performance of organic carbon-based reactive mixtures (wood chips, sawdust, composted municipal sewage sludge and leaf) under controlled groundwater flow conditions. The reactive media were sandwiched between top and bottom layers of silica sand and pyrite in the columns. The contaminants investigated, among others, were SO4, Zn and Ni. Over a 14-month period of experimentation, minimum rates of SO4 removal averaging between 500 and 800 mmold1 m3 were recorded. The removal efficiency of Zn at different initial concentrations ranged from 88% to 98% and that of Ni was >98%. However, the authors pointed out that sulfate reduction rates decreased in the course of the experiment. This was anticipated and attributed to the rapid decomposition of the easily decomposable fractions (e.g. cellulose) of the organic substrate in the initial stage of experiment leaving the lignin, which is somewhat recalcitrant to microbial degradation. Ludwig et al. (2002) also applied a compost-based sulfate reducing bacteria to remove sulfate and precipitate heavy metals including Cu, Co, Ni, Cd and Zn. The results after 21 months indicated significant reduction of the metals. The investigations of Benner et al. (1997, 1999, and 2002) buttress these findings. It is worthy of note, however, that the immobilization of heavy metals by sulfide precipitation can be reversed. Thus, it is of utmost importance to ensure that conditions that favor reversibility of these processes are repressed. 3.9.2.3. Chlorinated solvents. Among the known organic waste materials, compost and mulch have recently been frequently used as materials in biobarriers for the degradation of chlorinated solvents. Öztürk et al. (2012) investigated the anaerobic degradation of TCE in groundwater using commercial compost and eucalyptus mulch as organic media. Results indicated that a strong bioactive zone was created by the organic media which enabled the anaerobic biodegradation of TCE. Henry et al. (2003) documented that TCE concentrations decreased by an average of 98% within the mulch bio-wall after 3 months of installation. Wilson et al. (2010) also observed a 99% reduction of TCE in a mulch biowall. Lee et al. (2007) used tyre rubber for the removal of PCE and TCE. The tyre rubber was shown to be an effective sorbent and a good biowall material as altogether, 93% of TCE, 77% of PCE and 80% of organic matter were removed from groundwater. 3.10. Miscellaneous There are other materials including tin (Sn), Ni/Fe, Zn0/Fe0 and Fe0/Pd and Zn0/hydroxyapatite, Pd/H2 which have also been evaluated for their use in PRBs (Cheng et al., 1997; Fennelly and Roberts, 1998; USEPA, 1998; Ma and Wu, 2007; Song et al., 2008). However, they have not been frequently used due to insufficient understanding of their removal mechanisms, possible generation of other toxic products, long-term effectiveness (Scherer et al., 2000), costs and inability to treat a wide range of contaminants. Some attempts have recently been made to use immobilised membranes and resins in PRBs to remove toxic metals from groundwater and landfill leachates. Zawierucha et al. (2012) applied the PRB with polymer inclusion membranes (PIMs) for the removal of Cd2+ and Zn2+ ions from contaminated groundwater and reported the recovery factor values of 98% and 91%, respectively. They discussed also the performance of a lab-scale two-step treatment system for landfill leachate with PIMs followed by

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sorption on impregnated resins (Zawierucha et al., 2013). The results showed well performance of this system with the overall removal efficiency of 99%, 88% and 55% for Pb2+, Cd2+ and Zn2+ions, respectively. 3.11. Combination of reactive materials Single or individual materials were frequently applied in the early stages of the PRB technology. In recent times, however, combinations of materials are frequently being applied mainly to eliminate the drawbacks of the single/pure materials by increasing the long-term performance of the barriers, improving the permeability, reducing costs of using single/pure materials, providing multiple mechanisms for contaminants removal, as well as enhancing and accelerating the removal rates. Reactive materials have also been combined to allow treatment of multi-contaminant plumes (See Section 2 on multi-barriers). The combination of materials can be biotic–biotic, abiotic–biotic or abiotic–abiotic. Ma and Wu (2008) used two abiotic materials: zero-valent zinc (Zn0) and ZVI for the removal of TCE, and observed the rate of TCE degradation to be three times faster for the mixture than for ZVI alone. Saberi (2012) combined Ni/Fe nanoparticles and observed good performance in the removal of Pb2+. Moraci and Calabrò (2010) noted that a mixture of iron and pumice was effective in the Cu and Ni removal as well as in maintaining the long-term hydraulic conductivity of the PRB. Two biotic materials: compost and mulch combined for the degradation of chlorinated solvents (Henry et al., 2003; Öztürk et al., 2012) gave better results than when the materials are used individually (Ahmad et al., 2007; AFCEE, 2008). Mulch is more stable and not readily decomposed due to its high lignin content, and thus serves as a long-term supply of organic C, while compost easily decomposes and makes nutrients readily available for the microbes which participate in the anaerobic degradation of the chlorinated solvents. A combination of biotic and abiotic materials such as ZVI and organic substrates for the degradation of chlorinated solvents and immobilization of inorganic contaminants has also been reported (Mueller et al., 2004). Such combinations allow for the treatment of different contaminants as they provide different removal mechanisms. They have also been shown to improve the permeability of the barrier (ITRC, 2005). In some cases, the material combinations do not produce the synergistic (stimulatory) effect, showing rather an antagonistic (inhibitory) or no effect on contaminants removal or the overall performance of the PRBs. Köber et al. (2002) noted that a combination of GAC and Fe0 did not improve the removal of TCE and MCB due to the reduction in adsorption capacity of the GAC by the Fe0. Scherer et al. (2000) also mentioned that coupling of metals can introduce a catalytic or synergistic effect; however, it can also result in the formation of oxide layers, which may affect the performance of the barrier. The effect of combining materials on the overall performance of the PRBs is dependent on a number of factors including the ratio of materials in the mixture. As a general rule, the type and ratio of the materials in combination will depend on the objective to be achieved. The following may, however, be considered when selecting material combinations for PRBs: (i) contaminant(s) to be treated, (ii) removal mechanisms required, (iii) aquifer geology and geochemistry, (iv) availability and costs of the materials, and (v) the reactivity and long-term interactive effects of the materials. 4. PRB design In general, the design of a PRB comprises some consecutive steps that include: a preliminary technical and economic

assessment, characterization of the site where the barrier is to be constructed, selection of the reactive media, treatability studies (batch and column tests), engineering design, choice of the construction method, formulation of the monitoring plan, and economic analysis (Gavaskar et al., 2000). A key aspect of the PRB design is a good understanding of the site and aquifer characteristics, which includes the site geology, aquifer hydrogeology (Powell et al., 1998; McMahon et al., 1999; Klammler and Hatfield, 2008; Erto et al., 2011), geochemistry (Puls, 2006), microbial activity and the contaminated plume delineation (Powell et al., 1998; Erto et al., 2011). Assessment of the horizontal and vertical distribution of the contaminant, directions and rates of groundwater flow, and preferential flow paths are equally important. To avoid bypassing or overflow of the contaminant, temporal and depth variations in flow velocity and direction need to be understood (Federal Remediation Technologies Roundtable [FRTR], 2002; Henry et al., 2008; ITRC, 2011). In general, PRBs are site specific and unique to site characteristics (USEPA, 1998; Ott, 2000; Henry et al., 2008). Contaminants to be treated are also to be considered (USEPA, 1998). Table 3 presents some reactive materials and the optimal conditions for the removal of contaminants. The reactive material selection depends on the factors which have been described in Section 3. In addition, it is important to ensure the hydraulic permeability of the PRB, including screens and reactive media are at least twice that of the aquifer. Higher permeability within the barrier is recommended to avoid problems due to permeability changes with time due to precipitation of iron oxides/hydroxides, carbonates and or other metal precipitates in the treatment media or filter layers (USEPA, 1998). After the reactive material has been selected, the dimension, location and orientation of the barrier have to be defined. Two important parameters mutually dependent are width of the capture zone and the residence time. The capture zone refers to the width of the barrier necessary to capture the plume without bypassing. The residence time is defined as the time required for contaminated groundwater and the reactive material within the PRB to be in contact to achieve treatment goals (Ott, 2000; FRTR, 2002; Puls, 2006; Gillham et al., 2010). The residence time tres, of the first order reaction, can be calculated using Eq. (7) (Ott, 2000; Carey et al., 2002; Gillham et al., 2010):

tres ¼ ½ lnðC T =C 0 Þ=k

ð7Þ

where CT is the target concentration down-gradient of the PRB; C0 is the concentration of the contaminant entering the PRB; and k is the rate of reaction. The PRB thickness b, for the first order reaction, can be calculated using Eq. (8):

b ¼ v  tres  SF

ð8Þ

where: v is the velocity of groundwater through reactive media; tres is the residence time; and SF is a safety factor considered for all uncertainties in reaction kinetics and groundwater flow. The velocity of groundwater v, can be calculated using Eq. (9):

v ¼ KI=np

ð9Þ

where K is the reactive media hydraulic conductivity; I is the hydraulic gradient across PRB; and np is the reactive medium porosity. The design of the barrier must ensure sufficient residence time to treat the contaminant of interest. A maximum concentration of the contaminant, a maximum velocity and a conservative decay (degradation) rate are to be considered for residence time

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F. Obiri-Nyarko et al. / Chemosphere 111 (2014) 243–259 Table 3 Reactive materials, contaminants and optimal conditions for their removal. Reactive material

Contaminant

Optimal conditions

Reference

Zero valent IronZVI

Copper

Higher initial copper concentration and pore water velocity accelerates iron corrosion and slows down the rate of copper removal due to the formation of mineral precipitates Degradation of benzene and toluene is adversely affected by elevated pH Low pH is favorable to remove arsenic compounds in aerobic condition, while in relative anaerobic condition, acidic and alkaline condition seems to be favorable for arsenate and arsenite removal, respectively Removal rates dependent on pH. Mo removal occurs at high pH values Reaction rate of TCE degradation and iron corrosion decreases with pH

Komnitsas et al. (2007)

High influent pH, high influent concentrations of nitrate, chloride, and alkalinity, are problematic for PRBs Nitrobenzene reduction affected by pH and ZVI particle size Nitrobenzene reduction rate decreases when pH is increased A faster nitrobenzene reduction rate is achieved when a finer ZVI particle is used Increasing the initial nitrate concentration and pore velocity has inhibitor effect on the process of nitrate removal

Henderson and Demond (2007)

Benzene and toluene Arsenate and arsenite compounds

Molybdenum (Mo) Trichloroethylene (TCE) Organics, metals, and radionuclides Nitrobenzene

Nitrate

Chen et al. (2011) Sun et al. (2006)

Morrison et al. (2006) Chen et al. (2001)

Dong et al. (2010)

Hosseini et al. (2011)

Zeolites

Lead and copper

Favorable pH region from 5 to 6 higher removal efficiencies at a grain size (0.42–0.85 mm)

Lee et al. (2003), Park et al. (2002)

Surfactant modified zeolite (SMZ)

Hydrocarbons

Adsorption efficiency affected by low temperature (4 °C)

Hornig et al. (2008)

Granular activated carbon (GAC)

Hydrocarbons

Sorption capacity temperature dependent. Reduced absorption efficiency between 10% and 30% can be evidenced at 4 °C TCE adsorption capacity is promoted by a high B.E.T. surface area, micropore volume and carbon content and it is significantly affected by the presence of a non-ionic compound of similar structure (PCE) Trichloroethylene (TCE) adsorption decreases with increasing concentrations of monovalent ions (NaCl), calcium or dissolved oxygen Adsorption capacity decreases with increasing temperature, although the rate of adsorption may increase

Arora et al. (2011), Hornig et al. (2008)

Higher sorption capacity appears to be attributed to a higher organic carbon content and the lower polarity Concentrations of chlorinated solvents in excess of 10–100 milligrams per liter (mg L1) may not be completely degraded, resulting in the production and persistence of intermediate dechlorination products Groundwater seepage velocities of less than 0.3 m d1 are generally suitable for biowall systems, while seepage velocities greater than 0.3 m d1 will likely require multiple biowall trenches to effectively remediate the contaminant plume A pH close to neutral (i.e., 6.5–7.5) is the most conducive to the proliferation of healthy, diverse microbial populations The addition of a buffering material should be considered at sites with low alkalinity (less than 300 mg L1) or acidic pH (less than 6.5) In general, concentrations of DO less than 0.5 mg L1 and ORP levels less than 0 millivolts are desired to stimulate anaerobic degradation processes

Wei and Seo (2010), AFCEE (2008), Morris (2009), Henry et al. (2005), AFCEE (2005)

Trichloroethylene (TCE)

Organic and heavy metal contaminants Mulch and compost

TCE

estimates. However, for the final design to be cost-effective minimum barrier dimensions are to be achieved (Henry et al., 2008; Erto et al., 2011; Kacimov et al., 2011). After the geometrical parameters of the barrier together with the reactive material have been selected the performance of the PRB over time needs to be addressed. The PRB performance over time can be predicted by simulation of longevity scenarios with the aid of numerical models (Muegge, 2008). General speaking, numerical models are commonly used to support experimental data and to gain insight into the processes involved in the material degradation as well as the formation of secondary minerals. However, most of them do not take into consideration the changes in reactivity of the material over time, raising the question of the ability of these models to precisely estimate the long term performance. More recently, models are able to incorporate the declining reactivity and permeability of the material in order to adequately represent long- term performance (Kouznetsova et al.,

Erto et al. (2010), Kilduff and Karanfil (2002)

Thiruvenkatachari et al. (2008), Testa and Winegardner (2000)

2007; Jeen et al., 2011). Nevertheless, there are still only few models that are applied to real sites. 4.1. PRB configuration There are other PRB configurations in addition to the two main PRB configurations described in Section 2. These include the trench-and-gate system, in which contaminated groundwater is directed to the reactive material (gate) through permeable corridors, and reaction vessels in which reactive material is contained in a vessel, including reactors of sub-surface vertical flow (Carey et al., 2002). When there is not much confidence on the contaminant flux, it is recommended to place a funnel-and-gate system. When there are some uncertainties regarding the performance of the treatment system, a sequential gates system can be constructed in order to get some in-built redundancy. Nevertheless, the advantage of the

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continuous PRB is represented by the fact that is generally easier to design (Carey et al., 2002). 4.2. Construction methods There are several different construction methods for PRBs. These include deep soil mixing, vibrated beam, continuous deep trenching machines, vertical hydraulic fracturing (Puls, 2006), sheet pile walls, backhoe excavation, jetting and caissons. The choice of construction method depends mainly on the PRB dimensions, depth to the aquitard, and presence of highly consolidated sediments or rocks (Gavaskar et al., 2000). For instance, several PRBs have been installed with the method of backhoes, for which not too much skill is needed. It allows a rapid excavation rate, and a depth of 30 m is reachable. However, when it is necessary to dig depths greater than 70 m, the method of clamshell excavators is required for PRB installation, although it demands more skill (Sethi et al., 2011). Full scale iron barriers installed using trenching, braced excavation and hydraulic fracturing have shown satisfactory results allowing contaminant reductions greater than 95% (Hocking and Ospina, 1999). 4.3. Costs The capital costs associated with the implementation of a PRB system include: site characterization, design, installation, monitoring, replacement of the reactive materials (if required) and their recovery or disposal after the completion of the treatment (Blowes et al., 1998; Gavaskar et al., 2002; Ladwig, 2006). The design of the technology comprises treatability desk and lab studies to select the reactive material and determine its hydraulic properties and hydrologic modeling to select the barrier location, configuration and orientation. The largest capital costs are associated with the installation of the PRB. They are site-specific; however, strongly depend on the geometry (the length and depth) of the wall. Depth is a crucial factor since the deeper the aquifer the greater the installation costs (Gavaskar et al. 2002). The trenching accounts for 70% whereas the reactive material for 10–15% of the total costs (AFCEE, 2008). For some materials like mulch the only costs are related to handling and transport to the site. For ZVI the costs depend on the grain size, the fine-grain is usually more expensive than coarse-grain materials (ITRC, 2011). The main operational and maintenance costs include the long-term performance, monitoring and the replacement of the reactive media. The replacement costs are difficult to quantify, since little is known about the long-term performance of most of reactive materials used in PRBs. However, once the material is depleted or clogged with precipitates, replacement is needed resulting in an increase of the maintenance costs (Ladwig, 2006). 5. PRB longevity The time at which a PRB continues to treat contaminants at designed levels is finite and is defined as longevity of the barrier (Robertson et al., 2000; Henderson and Demond, 2007; ITRC, 2011). In principle, when designing a PRB sufficient amounts of reactive materials might be placed within the barrier to reduce contaminant concentrations to target values. However, when the groundwater dissolved constituents come in contact with the reactive material, numerous reactions take place that may cause the formation of mineral precipitates and compromise the barrier performance (Mackenzie et al., 1999; Phillips et al., 2000; Furukawa et al., 2002; Moon et al., 2008). Little is known about the performance of different reactive materials besides ZVI (Geranio, 2007), the most common material

used up to date. Numerous authors have reported a decrease in ZVI performance with time as a result of the accumulation of secondary mineral precipitates and the subsequent impact on barrier hydraulics due to pore filling (Eykholt et al., 1999; Mackenzie et al., 1999; Liang et al., 2005). A decrease in iron reactivity has also an important influence in the barrier performance and is attributed to the consumption of the reactive material and the decline of reactive surface of the material due to mineral precipitates (O’Hannesin and Gillham, 1998; Vogan et al., 1999; Liang et al., 2000; Kamolpornwijit et al., 2003; Wilkin and Puls, 2003; Li et al., 2006; Jeen et al., 2008). In the case of ZVI injection, reduction of longevity of the barrier is produced when insufficient or non-uniform mass of the material is introduced (ITRC, 2011). The process in which these precipitates are produced is called iron corrosion and starts when the ZVI gets in contact with the groundwater and its dissolved constituents. This reaction results in a decrease in Eh, an increase in pH and a subsequent generation of mineral precipitates (Blowes et al., 2000; Geranio, 2007). In some cases the reactions are not related to the contaminant treated but to the concentrations of dissolved constituents. Dissolved inorganic constituents that react with ZVI in PRBs are: oxygen, magnesium, carbonates, calcium and sulfates (Puls et al., 1999; Puls, 2006; Jeen et al., 2008). Under anaerobic conditions, reduction of water occurs (Eq. (10)) (Zolla et al., 2007), while under aerobic conditions, dissolved oxygen acts as the oxidant and can lead to the production of ferrous and ferric iron (Eqs. (11) and (12)) (Wilkin et al., 2000):

Fe0 þ 2H2 O ! Fe2þ þ H2 þ 2OH

ð10Þ

2Fe0 þ O2 þ 2H2 O ! 2Fe2þ þ 4OH

ð11Þ

4Fe2þ þ O2 þ 2Hþ ! 4Fe3þ þ 2OH

ð12Þ

Under aerobic conditions, ferric oxides and oxy-hydroxides precipitate while under anaerobic conditions and at high pH, ferrous hydroxide or green-rust minerals are expected to form (Wilkin et al., 2000). Consequences of iron corrosion are the loss of iron mass and the formation of a passivating layer on the iron surface reducing the reactivity of the remaining iron mass (Su and Puls, 2004; Puls, 2006). Nonetheless, some iron corrosion products including iron hydroxides, oxy-hydroxides, and mixed valance Fe(II)–Fe(III) green rusts are beneficial for As removal from water (Furukawa et al., 2002; Wilkin et al., 2008). Dissolved hydrogen produced during iron corrosion can be utilized as an electron donor by various anaerobic bacteria for nitrate and sulfate reduction. This process produces however, the proliferation of bacteria and consequent pore filling by microorganisms and metabolic by-products (i.e. nitrogen and methane) (Wilkin et al., 2000). There is a consensus among researchers indicating that although iron corrosion products might cause a reduction in hydraulic properties and reactivity of the material, field studies have demonstrated that barriers performance has not deteriorated in some cases, therefore the degree and speed of affectation has been varied. However, it is also the case that many installed PRBs are relatively new and definitive conclusions cannot be drawn yet (Kouznetsova et al., 2007). Researchers are still approaching the prediction of long-term performance by the aid of modeling tools. Nonetheless, there is still a lack of field data to sustain these studies (Henderson and Demond, 2007), and the information regarding recently employed reactive materials and their long-term performance is yet to be tested when sufficient field data has been generated. Efforts to predict long-term performance include geochemical modeling and laboratory column studies. However, difficulties comparing both

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types of studies are common. On one hand, by modeling there is an attempt to foresee the geochemical behavior of parameters that affect performance; on the other hand, short term accelerated columns studies may not represent the real material aging process (Farrell et al., 2000; Henderson and Demond, 2007). To overcome the iron pore filling and permeability reduction, some strategies have been evaluated. The simplest one is to mix iron with pumice in different weight ratios, as well as ZVI granular mixtures and sand, in order to preserve the designed removal efficiency (Moraci and Calabrò, 2010). Blowes et al. (2000) suggest that an increase of the barrier thickness together with a more even distribution of the materials can improve the treatment. Li and Benson (2010) evaluated five strategies to limit the effect of a PRB fouling and the reduction in its hydraulic performance. These strategies include equalization zones up- and down-gradient the barrier by addition of pea gravel, placement of a pre-treatment zone up gradient, pH adjustment, utilization of ZVI larger grains and mechanical mixing. However, none of the strategies eliminated porosity reduction or prevented an increase in residence time during 30 years period. It was concluded that the most costeffective strategy would be the placement of a pre-treatment zone. 6. Concluding remarks The use of PRBs for sustainable groundwater treatment has come a long way since it was invented. Clearly, our review has shown that even though the technology is relatively young, a substantial amount of research has been done, leading to several PRB design modifications, discovery of new reactive materials, improved PRBs performance, and slowly transitioning it from an innovative into a developed (proven) technology. Clearly, the spectrum of contaminants that can be treated with PRBs has been broadened, owing to the discovery of suitable materials and improved understanding of the removal mechanisms. Though new materials have and are being discovered, ZVI still remains the most often applied material both in laboratory and field investigations. There is also evidence that lots of efforts have been made to elucidate the mechanisms of contaminants removal by diverse reactive media and the factors controlling these; however, more research is still required as understanding of these mechanisms is still lacking, particularly for the newly discovered materials. Regular rather than occasional monitoring of biogeochemical changes occurring within and down-gradient of a barrier is required to ensure the success of the technology. The oldest PRB is close to two decades; however, this is still not enough to provide sufficient information to help in adequately understanding and predicting their long-term performance. Key aspects for a proper and effective design of the technology and to prevent any failures that may occur are an adequate site characterization, the understanding of groundwater flow conditions, and numerical groundwater flow and contaminant transport modeling. Acknowledgments This research was completed within the framework of the Marie Curie Initial Training Network ADVOCATE – ‘Advancing sustainable in situ remediation for contaminated land and groundwater’, funded by the European Commission, Marie Curie Actions Project No. 265063. References AFCEE, Air Force Center for Environmental Excellence, 2005. Treatability Study Work Plan for Bioremediation of Chlorinated Solvents Using a Permeable Reactive Biowall at the BG05 Site. Texas.

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AFCEE, Air Force Center for Environmental Excellence, 2008. Technical Protocol for Enhanced Anaerobic Bioremediation Using Permeable Mulch Biowalls and Bioreactors. Denver CO. Ahmad, F., McGuire, T.M., Lee, R.S., Becvar, E., 2007. Considerations for the design of organic mulch permeable reactive barriers. Remed. J. 18 (1), 59–72. Aivalioti, M., Mavrakis, N., Gidarakos, E., 2008. Municipal Solid Waste Compost Utilization for Groundwater Remediation in Biobarriers. In: Proceedings of the 1st International Conference Hazardous Waste Management, Chania-Crete, Greece, p. 19. Allison, J.D., Brown, D.S., Novo-Gradac, K.J., 1991. MINTEQA2/PRODEFA2, A geochemical Assessment Model for Environmental Systems. Report EPA/600/ 3-91/021, U.S. Environmental Protection Agency, Athens, Georgia. 107pp. Arey, J.S., Seaman, J.C., Bertsch, P.M., 1999. Immobilization of uranium in contaminated sediments by hydroxyapatite addition. Environ. Sci. Technol. 33 (2), 337–342. Arora, M., Snape, I., Stevens, G.W., 2011. Toluene sorption by granular activated carbon and its use in cold regions permeable reactive barrier: Fixed bed studies. Cold Reg. Sci. Technol. 69, 59–63. Baker, M.J., Blowes, D.W., Ptacek, C.J., 1998. Laboratory development of permeable reactive mixtures for the removal of phosphorus from onsite wastewater disposal systems. Environ. Sci. Technol. 32 (15), 2308–2316. Benner, S.G., Blowes, D.W., Ptacek, C.J., 1997. A full-scale porous reactive wall for prevention of acid mine drainage. Ground Water Monit. R. 17 (4), 99–107. Benner, S.G., Blowes, D.W., Gould, W.D., Herbert, R.B., Ptacek, C.J., Herbert, R.B., 1999. Geochemistry of a permeable reactive barrier for metals and acid mine drainage. Environ. Sci. Technol. 33 (16), 2793–2799. Benner, S.G., Blowes, D.W., Ptacek, C.J., Mayer, K.U., 2002. Rates of sulfate reduction and metal sulfide precipitation in a permeable reactive barrier. Appl. Geochem. 17 (3), 301–320. Bianchi-Mosquera, G.C., Allen-King, R.M., Mackay, D.M., 1994. Enhanced degradation of dissolved benzene and toluene using a solid oxygen-releasing compound. Ground Water Monit. R. 14 (1), 120–128. Birke, V., Burmeier, H., Jefferis, S., Gaboriau, H., Touze, S., Romain, C., 2007. Permeable reactive barriers (PRBs) in Europe: potentials and expectations. Ital. J. Eng. Geol. Environ. 1, 1–8. Blowes, D.W., Ptacek, C.J., Benner, S.G., McRae, C.W.T., 1998. Treatment of dissolved metals using permeable reactive barriers. In: Proceedings of the Groundwater Quality Conference 1998, Tubingen, Germany, pp. 483–490. Blowes, D.W., Ptacek, C.J., Benner, S.G., McRae, C.W., Bennett, T.A., Puls, R.W., 2000. Treatment of inorganic contaminants using permeable reactive barriers. J. Contam. Hydrol. 45 (1–2), 123–137. Bone, B.D., 2012. Review of UK guidance on permeable reactive barriers. In: 2012 Taipei International Conference on Remediation and Management of Soil and Groundwater Contaminated Sites, Taipei, Taiwan, pp. 611–768. Borden, R.C., 2007. Concurrent bioremediation of perchlorate and 1,1,1trichloroethane in an emulsified oil barrier. J. Contam. Hydrol. 94 (1–2), 13–33. Borden, R.C., 2008. Development of Permeable Reactive Barriers (PRB) Using Edible Oils. Final Report Strategic Research and Development Program SERDP Project ER-1205, North Carolina State University, 143pp. Borden, R.C., Goin, R.T., Kao, C.-M., 1997. Control of BTEX migration using a biologically enhanced permeable barrier. Ground Water Monit. Remediat. 17 (1), 70–80. Bostick, W., 2003. Use of Apatite for Chemical Stabilization of Subsurface Contaminants. Final Report Contract No. DE-AC26-01NT41306, U.S. Department of Energy, Tennessee, 186pp. Bowman, R.S., Haggerty, G.M., Huddleston, R.G., Neel, D., Flynn, M.M., 1995. Sorption of nonpolar organic compounds, inorganic cations, and inorganic oxyanions by surfactant-modified zeolites. ACS Symp. Ser. 594, 54–64. Burgin, A.J., Hamilton, S.K., 2007. Have we overemphasized the role of denitrification in aquatic ecosystems? A review of nitrate removal pathways. Front. Ecol. Environ. 5 (2), 89–96. Cameron, S.G., Schipper, L.A., 2010. Nitrate removal and hydraulic performance of organic carbon for use in denitrification beds. Ecol. Eng. 36 (11), 1588–1595. Cappai, G., De Gioannis, G., Muntoni, A., Spiga, D., Zijlstra, J.J.P., 2012. Combined use of a transformed red mud reactive barrier and electrokinetics for remediation of Cr/As contaminated soil. Chemosphere 86 (4), 400–408. Careghini, A., Saponaro, S., Sezenna, E., 2013. Biobarriers for groundwater treatment: a review. Water Sci. Technol. 67 (3), 453–468. Carey, M.A., Fretwell, B.A., Mosley, N.G., Smith, J.W.N., 2002. Guidance on the Use of Permeable Reactive Barriers for Remediating Contaminated Groundwater. National Groundwater and Contaminated Land Centre Report NC/01/51, UK Environment Agency, Bristol. 140pp. Chen, J.L., Al-Abed, S.R., Ryan, J.A., Li, Z., 2001. Effects of pH on dechlorination of trichloroethylene by zero-valent iron. J. Hazard. Mater. 83, 243–254. Chen, Y., Li, J., Lei, C., Shim, H., 2011a. Interactions between BTEX, TPH, and TCE during their bio-removal from the artificially contaminated water. In: Bionature 2011: The Second International Conference on Bioenvironment, Biodiversity and Renewable Energies, Venice, Italy, pp. 33–37. Chen, L., Liu, F., Liu, Y., Dong, H., Colberg, P.J.S., 2011b. Benzene and toluene biodegradation down gradient of a zero-valent iron permeable reactive barrier. J. Hazard. Mater. 188 (1–3), 110–115. Cheng, I.F., Muftikian, R., Fernando, Q., Korte, N., 1997. Reduction of nitrate to ammonia by zero-valent iron. Chemosphere 35 (11), 2689–2695. Choe, S., Liljestrand, H.M., Khim, J., 2004. Nitrate reduction by zero-valent iron under different pH regimes. Appl. Geochem. 19 (3), 335–342.

256

F. Obiri-Nyarko et al. / Chemosphere 111 (2014) 243–259

Chung, H.I., Kim, S.K., Lee, Y.S., Yu, J., 2007. Permeable reactive barrier using atomized slag material for treatment of contaminants from landfills. Geosci. J. 11 (2), 137–145. Clap, L., Abdelrahman, M., Pana, P.E., 2012. Using either hydrogen or dithionite as reductant in uranium contaminated groundwater at post-leach uranium mining sites. In: American Society for Engineering Education Annual Conference and Exposition, San Antonio, Texas, pp. 1707–1720. Conca, J., Strietelmeier, E., Lu, N., Ware, S.D., Taylor, T.P., Wright, J., 2002. Treatability study of reactive materials to remediate ground water contaminated with radionuclides, metals and nitrates in a four-component permeable reactive barrier. In: Naftz, S., Morrison, J.A., Fuller, C.C. (Eds.), Groundwater Remediation of Trace Metals, Radionuclides, and Nutrients, with Permeable Reactive Barriers. Academic Press, San Diego, CA, pp. 221–252. Coombs, D., Alberti, A., Armbruster, T., Artioli, G., Colella, C., Galli, E., Grice, J.D., Liebau, F., Mandarino, J., Minato, H., Nickel, E.H., Passaglia, E., Peacor, D.R., Quartieri, S., Rinaldi, R., Ross, M., Sheppard, R.A., Tillmanns, E., Vezzalini, G., 1997. Recommended nomenclature for zeolite minerals: report of the subcommittee on zeolites of the international mineralogical association, commission on new minerals and mineral names. Can. Mineral. 35, 1571–1606. Cornelissen, G., Gustafsson, Ö., Bucheli, T.D., Jonker, M.T.O., Koelmans, A.A., Van Noort, P.C.M., 2005. Extensive sorption of organic compounds to black carbon, coal, and kerogen in sediments and soils: mechanisms and consequences for distribution, bioaccumulation, and biodegradation. Environ. Sci. Technol. 39 (18), 6881–6895. Da-Silva, M.L.B., Johnson, R.L., Alvarez, P.J.J., 2007. Microbial characterization of groundwater undergoing treatment with a permeable reactive iron barrier. Environ. Eng. Sci. 24 (8), 1122–1127. Di-Nardo, A., Di Natale, M., Erto, A., Musmarra, D., Bortone, I., 2010. Permeable reactive barrier for groundwater PCE remediation: the case study of a solid waste landfill pollution. In: Pierucci, S., Ferraris, G.B. (Eds.), Proceedings of the 20th European Symposium on Computer Aided Process Engineering-ESCAPE20. Elsevier B.V, Naples, Italy. Dixon, J.B., Weed, S.B., Dinauer, R.C., 1989. Minerals in soil environments. In: Dixon, J.B., Weed S.B. (Eds.), second ed. Soil Science Society of America Book Series, Madison, Wisconsin. Dong, J., Zhao, Y., Zhao, R., Zhou, R., 2010. Effects of pH and particle size on kinetics of nitrobenzene reduction by zero-valent iron. J. Environ. Sci. 22, 1741–1747. Erto, A., Andreozzi, R., Lancia, A., Musmarra, D., 2010. Factors affecting the adsorption of trichloroethylene onto activated carbons. Appl. Surf. Sci. 256, 5237–5242. Erto, A., Lancia, A., Bortone, I., Di Nardo, A., Di Natale, M., Musmarra, D., 2011. A procedure to design a permeable adsorptive barrier (PAB) for contaminated groundwater remediation. J. Environ. Manage. 92 (1), 23–30. Eykholt, G.R., Elder, C.R., Benson, C.H., 1999. Effects of aquifer heterogeneity and reaction mechanism uncertainty on a reactive barrier. J. Hazard. Mater. 68 (1– 2), 73–96. Fallico, C., Troisi, S., Molinari, A., Rivera, M.F., 2010. Characterization of broom fibers for PRB in the remediation of aquifers contaminated by heavy metals. Biogeosciences 7 (8), 2545–2556. Farrell, J., Bostick, W.D., Jarabek, R.J., Fiedor, J.N., 1999. Uranium removal from ground water using zero valent iron media. Ground Water 37 (4), 618–624. Farrell, J., Kason, M., Melitas, N., Li, T., 2000. Investigation of the long-term performance of zero-valent iron for reductive dechlorination of trichloroethylene. Environ. Sci. Technol. 34 (3), 514–521. Fennelly, J.P., Roberts, A.L., 1998. Reaction of 1,1,1-trichloroethane with zero-valent metals and bimetallic reductants. Environ. Sci. Technol. 32 (13), 1980–1988. Fenton, O., Healy, M.G., Rodgers, M., 2009. Use of ochre from an abandoned metal mine in the south east of Ireland for phosphorus sequestration from dairy dirty water. J. Environ. Qual. 38 (3), 1120–1125. Fenton, O., Healy, M.G., Brennan, F., Jahangir, M.M.R., Lanigan, G.J., Richards, K.G., Thornton, S.F., Ibrahim, T.G., 2014. Permeable reactive interceptors: blocking diffuse nutrient and greenhouse gases losses in key areas of the farming landscape. J. Agric. Sci. 1, 1–11. Fiedor, J.N., Bostick, W.D., Jarabek, R.J., Farrell, J., 1998. Understanding the mechanism of uranium removal from groundwater by zero-valent iron using X-ray photoelectron spectroscopy. Environ. Sci. Technol. 32 (10), 1466–1473. Fiorenza, S., Oubre, C.L., Ward, C.H., 2000. Sequenced reactive barriers for groundwater remediation. In: Fiorenza, C.H., Oubre, S., Ward, C.L. (Eds.), CRC Press, Boca Raton, FL (US). Fronczyk, J., Pawluk, K., Michniak, M., 2010. Application of permeable reactive barriers near roads for chloride ions removal. Ann. Warsaw Univ. of Life Sci.SGGW, Land Reclam. 42(2), 249–259. FRTR, Federal Remediation Technologies Roundtable, 2002. Evaluation of Permeable Reactive Barrier Performance. Washington DC, 45pp. Fuller, C.C., Bargar, J.R., Davis, J.A., 2003. Remediation of uranium-contaminated ground water at fry canyon, Utah. Science Highlight (November), 2–5. Furukawa, Y., Kim, J.-W., Watkins, J., Wilkin, R.T., 2002. Formation of ferrihydrite and associated iron corrosion products in permeable reactive barriers of zerovalent iron. Environ. Sci. Technol. 36 (24), 5469–5475. Gallinati, J.D., Warner, S.D., Yamane, C.L., Szerdy, F.S., Hankins, D.A., Major, D.W., 1995. Design and evaluation of an in situ ground water treatment wall composed of zero-valent iron. Ground Water 33 (5), 834–835. Gavaskar, A.R., 1999. Design and construction techniques for permeable reactive barriers. J. Hazard. Mater. 68 (1–2), 41–71.

Gavaskar, A., Gupta, N., Sass, B., Janosy, R., Hicks, J., 2000. Design Guidance for Application of Permeable Reactive Barriers for Groundwater Remediation. Columbus, OH, 157pp. Genç-Fuhrman, H., Bregnhøj, H., McConchie, D., 2005. Arsenate removal from water using sand–red mud columns. Water Res. 39 (13), 2944–2954. Geranio, L., 2007. Review of Zero Valent Iron and Apatite as reactive materials for Permeable Reactive Barrier. Department of Environmental Sciences, ETH Zurich, Zurich, Switzerland. Giammar, D.E., Hering, J.G., 1999. Kinetics of uranium sorption on goethite. In: Symposium on colloidal and interfacial phenomena in aquatic environments. Anaheim, California, pp. 420–422. Gibert, O., Pomierny, S., Rowe, I., Kalin, R.M., 2008. Selection of organic substrates as potential reactive materials for use in a denitrification permeable reactive barrier (PRB). Bioresour. Technol. 99 (16), 7587–7596. Gillham, R.W., Vogan, J., Gui, L., Duchene, M., Son, J., Mccarty, P.L., Henry, B.M., 2010. Iron barrier walls for chlorinated solvent remediation. In: Stroo, H.F., Ward, C.H. (Eds.), In Situ Remediation of Chlorinated Solvent Plumes. Springer, New York, NY, pp. 537–571. Giménez, J., Martínez, M., De Pablo, J., Rovira, M., Duro, L., 2007. Arsenic sorption onto natural hematite, magnetite, and goethite. J. Hazard. Mater. 141 (3), 575– 580. Golab, A.N., Peterson, M.A., Indraratna, B., 2006. Selection of potential reactive materials for a permeable reactive barrier for remediating acidic groundwater in acid sulphate soil terrains. Q. J. Eng. Geol. Hydroge. 39 (2), 209–223. Gu, B., Watson, D.B., Wu, L., Phillips, D.H., White, D.C., Zhou, J., 2002. Microbiological characteristics in a zero-valent iron reactive barrier. Environ. Monit. Assess. 77 (3), 293–309. Guerin, T.F., Horner, S., McGovern, T., Davey, B., 2002. An application of permeable reactive barrier technology to petroleum hydrocarbon contaminated groundwater. Water Res. 36 (1), 15–24. Healy, M.G., Ibrahim, T.G., Lanigan, G.J., Serrenho, A.J., Fenton, O., 2012. Nitrate removal rate, efficiency and pollution swapping potential of different organic carbon media in laboratory denitrification bioreactors. Ecol. Eng. 40, 198–209. Henderson, A.D., Demond, A.H., 2007. Long-term performance of zero-valent iron permeable reactive barriers: a critical review. Environ. Eng. Sci. 24 (4), 401–423. Henderson, A., Demond, A., 2011. Impact of solids formation and gas production on the permeability of ZVI PRBs. J. Environ. Eng. 137 (8), 689–696. Henry, B.M., Goodspeed, M., Gonzales, J.R., Haas, P.E., Oakley, D., 2003. Permeable mulch biowall for enhanced bioremediation of chlorinated ethenes. In: Magar, V.S., Kelley, M.E. (Eds.), Proceedings of the Seventh International In Situ and OnSite Bioremediation Symposium. Orlando, FL, p. K03. Henry, B.M., Perlmutter, M.W., Downey, D.C., 2008. Permeable organic biowalls for remediation of perchlorate in groundwater. In: Stroo, H.F., Ward, C.H. (Eds.), In Situ Bioremediation of Perchlorate in Groundwater. Springer, New York, NY, pp. 177–198. Henry, B.M., Downey, D.C., Griffiths, D.R., Krumholz, M., Gonzales, J.R., Becvar, E.S.K., 2005. Design of Mulch Biowalls for Enhanced In Situ Anaerobic Bioremediation. In: Alleman, B.C., Kelley, M.E. (Eds.), Proceedings of the Eighth International In Situ and On-Site Bioremediation Symposium. Battelle Press, Baltimore, Maryland. pp. L–05. Herbert, R.B., Benner, S.G., Blowes, D.W., 2000. Solid phase iron–sulfur geochemistry of a reactive barrier for treatment of mine drainage. Appl. Geochem. 15 (9), 1331–1343. Hocking, G., Ospina, R.I., 1999. Construction and performance monitoring of in situ reactive barriers. In: IBC’s 2nd Annual Environmental Technology Symposium. Scottsdale, AZ. Hornig, G., Northcott, K., Snape, I., Stevens, G., 2008. Assessment of sorbent materials for treatment of hydrocarbon contaminated ground water in cold regions. Cold Reg. Sci. Technol. 53, 83–91. Hosseini, S.M., Ataie-Ashtiani, B., Kholghi, M., 2011. Bench-scaled nano-Fe0 permeable reactive barrier for nitrate removal. Ground Water Monit. R. 31 (4), 82–94. Istok, J.D., Park, M.M., Peacock, A.D., Oostrom, M., Wietsma, T.W., 2007. An experimental investigation of nitrogen gas produced during denitrification. Ground Water 45 (4), 461–467. ITRC, Interstate Technology & Regulatory Council, 2005. Permeable Reactive Barriers: Lessons Learned/New Directions. PRB: Technology Update Team, Washington, DC, 101pp. ITRC, Interstate Technology & Regulatory Council, 2011. Permeable Reactive Barrier: Technology Update. PRB: Technology Update Team, Washington, D.C., 156 pp. James, B.R., Rabenhorst, M.C., Frigon, G.A., 1992. Phosphorus sorption by peat and sand amended with iron oxides or steel wool. Water Environ. Res. 64 (5), 7. Jeen, S.-W., Blowes, D.W., Gillham, R.W., 2008. Performance evaluation of granular iron for removing hexavalent chromium under different geochemical conditions. J. Contam. Hydrol. 95 (1–2), 76–91. Jeen, S.-W., Gillham, R.W., Przepiora, A., 2011. Predictions of long-term performance of granular iron permeable reactive barriers: field-scale evaluation. J. Contam. Hydrol. 123 (1–2), 50–64. Jirasko, D., 2012. Problems connected with use of permeable Reactive Barriers for groundwater treatment. Construction on brownfields. Czech Technical University in Prague, Prague, Czech Republic. pp. 145–154. Johnson, S.J., Woolhouse, K.J., Prommer, H., Barry, D.A., Christofi, N., 2003. Contribution of anaerobic microbial activity to natural attenuation of benzene in groundwater. Eng. Geol. 70, 343–349.

F. Obiri-Nyarko et al. / Chemosphere 111 (2014) 243–259 Johnston, J.J., Borden, R.C., Barlaz, M.A., 1996. Anaerobic biodegradation of hazardous organics in groundwater down gradient of a sanitary landfill. J. Contam. Hydrol. 23 (4), 263–283. Kacimov, A.R., Klammler, H., Il’yinskii, N., Hatfield, K., 2011. Constructal design of permeable reactive barriers: groundwater-hydraulics criteria. J. Eng. Math. 71 (4), 319–338. Kamolpornwijit, W., Liang, L., West, O.R., Moline, G.R., Sullivan, A.B., 2003. Preferential flow path development and its influence on long-term PRB performance: column study. J. Contam. Hydrol. 66 (3–4), 161–178. Kanel, S.R., Greneche, J.-M., Choi, H., 2006. Arsenic(V) removal from groundwater using nano scale zero-valent iron as a colloidal reactive barrier material. Environ. Sci. Technol. 40 (6), 2045–2050. Kao, C., Borden, R., 1997. Enhanced TEX biodegradation in nutrient briquet-peat barrier system. J. Environ. Eng. 123 (1), 18–24. Katz, L.E., Kinney, K.A., Bowman, R.S., Sullivan, E.J., Kwon, S., Darby, E.B., Chen, L.-J., Altare, C.R., 2006. Treatment of Produced Waters Using a Surfactant Modified Zeolite/Vapor Phase Bioreactor System. Final Technical Report DE-FC2602NT15461, U.S. Department of Energy, National Energy Technology Laboratory, New Mexico, 133pp. Kershaw, D.S., Kulik, B.C., Pamukcu, S., 1997. Ground rubber: sorption media for ground water containing benzene and O-Xylene. J. Geotech. Geoenviron. Eng. 123 (4), 324–334. Kilduff, J.E., Karanfil, T., 2002. Trichloroethylene adsorption by activated carbon preloaded with humic substances: effects of solution chemistry. Water Res. 36, 1685–1698. Klammler, H., Hatfield, K., 2008. Analytical solutions for flow fields near continuous wall reactive barriers. J. Contam. Hydrol. 98 (1–2), 1–14. Köber, R., Schäfer, D., Ebert, M., Dahmke, A., 2002. Coupled in situ reactors using Fe0 and activated carbon for the remediation of complex contaminant mixtures in groundwater. In: Thornton, S.F., Oswald, S.E. (Eds.), Proceedings of the Groundwater Quality 2001 Conference. IAHS, Sheffield, UK, pp. 18–21. Komnitsas, K., Bartzas, G., Fytas, K., Paspaliaris, I., 2007. Long-term efficiency and kinetic evaluation of ZVI barriers during clean-up of copper containing solutions. Miner. Eng. 20, 1200–1209. Korte, N.E., 2001. Zero-Valent Iron Permeable Reactive Barriers: A Review of Performance. Environmental Sciences Division Publication No. 5056, U.S. Department of Energy, Washington DC. Kouznetsova, I., Bayer, P., Ebert, M., Finkel, M., 2007. Modelling the long-term performance of zero-valent iron using a spatio-temporal approach for iron aging. J. Contam. Hydrol. 90 (1–2), 58–80. Kriegman-King, M.R., Reinhard, M., 1992. Transformation of carbon tetrachloride in the presence of sulfide, biotite, and vermiculite. Environ. Sci. Technol. 26 (11), 2198–2206. Kwon, S., Sullivan, E.J., Katz, L.E., Bowman, R.S., Kinney, K.A., 2011. Laboratory and field evaluation of a pretreatment system for removing organics from produced water. Water Environ. Res. 83 (9), 843–854. Lai, K., Lo, I., Birkelund, V., Kjeldsen, P., 2006. Field monitoring of a permeable reactive barrier for removal of chlorinated organics. J. Environ. Eng. 132 (2), 199–210. Lapointe, F., Fytas, K., McConchie, D., 2006. Efficiency of Bauxsol™ in permeable reactive barriers to treat acid rock drainage. Mine Water Environ. 25 (1), 37–44. Lee, M.-G., Cheon, J.-K., Kam, S.-K., 2003. Heavy metal adsorption characteristics of zeolite synthesized from fly ash. J. Ind. Eng. Chem. 9 (2), 174–180. Lee, J.-Y., Moon, C.-H., Kim, J.-H., Oh, B.-T., 2007. Feasibility study of the bio-barrier with biologically-active tire rubbers for treating chlorinated hydrocarbons. Geosci. J. 11 (2), 131–136. Li, L., Benson, C.H., 2010. Evaluation of five strategies to limit the impact of fouling in permeable reactive barriers. J. Hazard. Mater. 181 (1–3), 170–180. Li, X., Zhang, W., 2007. Sequestration of metal cations with zerovalent iron nanoparticles – a study with high resolution X-ray photoelectron spectroscopy (HR-XPS). J. Phys. Chem. C 111 (19), 6939–6946. Li, Z., Jones, H.K., Bowman, R.S., Helferich, R., 1999. Enhanced reduction of chromate and PCE by pelletized surfactant-modified zeolite/zerovalent iron. Environ. Sci. Technol. 33 (23), 4326–4330. Li, L., Benson, C.H., Lawson, E.M., 2006. Modeling porosity reductions caused by mineral fouling in continuous-wall permeable reactive barriers. J. Contam. Hydrol. 83 (1–2), 89–121. Liang, L., Korte, N., Gu, B., Puls, R., Reeter, C., 2000. Geochemical and microbial reactions affecting the long-term performance of in situ ‘‘iron barriers’’. Adv. Environ. Res. 4 (4), 273–286. Liang, L., Sullivan, A.B., West, O.R., Moline, G.R., Kamolpornwijit, W., 2003. Predicting the precipitation of mineral phases in permeable reactive barriers. Environ. Eng. Sci. 20 (6), 635–653. Liang, L., Moline, G.R., Kamolpornwijit, W., West, O.R., 2005. Influence of hydrogeochemical processes on zero-valent iron reactive barrier performance: a field investigation. J. Contam. Hydrol. 78 (4), 291–312. Liao, C.-H., Kang, S.-F., Hsu, Y.-W., 2003. Zero-valent iron reduction of nitrate in the presence of ultraviolet light, organic matter and hydrogen peroxide. Water Res. 37 (17), 4109–4118. Liao, D., Zheng, W., Li, X., Yang, Q., Yue, X., Guo, L., Zeng, G., 2010. Removal of lead(II) from aqueous solutions using carbonate hydroxyapatite extracted from eggshell waste. J. Hazard. Mater. 177 (1–3), 126–130. Lien, H.-L., Wilkin, R.T., 2005. High-level arsenite removal from groundwater by zero-valent iron. Chemosphere 59 (3), 377–386. Liu, X., Pinto, N.G., 1997. Ideal adsorbed phase model for adsorption of phenolic compounds on activated carbon. Carbon 35 (9), 1387–1397.

257

Lower, S.K., Maurice, P.A., Traina, S.J., Carlson, E.H., 1998. Aqueous Pb sorption by hydroxylapatite: applications of atomic force microscopy to dissolution, nucleation, and growth studies. Am. Mineral. 83, 147–158. Ludwig, R.D., McGregor, R.G., Blowes, D.W., Benner, S.G., Mountjoy, K., 2002. A permeable reactive barrier for treatment of heavy metals. Ground Water 40 (1), 59–66. Ludwig, R.D., Smyth, D.J.A., Blowes, D.W., Spink, L.E., Wilkin, R.T., Jewett, D.G., Weisener, C.J., 2009. Treatment of arsenic, heavy metals, and acidity using a mixed ZVI-compost PRB. Environ. Sci. Technol. 43 (6), 1970–1976. Ma, C., Wu, Y., 2007. Dechlorination of perchloroethylene using zero-valent metal and microbial community. Environ. Geol. 55 (1), 47–54. Ma, C., Wu, Y., 2008. Dechlorination of perchloroethylene using zero-valent metal and microbial community. Environ. Geol. 55 (1), 47–54. Ma, Q.I.Y., Tralna, S.J., Logan, T.J., Ryan, J.A., 1993. In situ lead immobilization by apatite. Environ. Sci. Technol. 27 (9), 1803–1810. Mackay, D.M., Wilson, R.D., Naas, C., Wood, I., 2002. In Situ treatment of MTBEcontaminated groundwater at two sites in California. In: Thornton, S.F., Oswald, S.E. (Eds.), Groundwater Quality: Natural and Enhanced Restoration of Groundwater Pollution, Sheffield, p. 333–39. Mackenzie, P.D., Horney, D.P., Sivavec, T.M., 1999. Mineral precipitation and porosity losses in granular iron columns. J. Hazard. Mater. 68 (1–2), 1–17. Manios, T., Stentiford, E.I., Millner, P., 2003. Removal of heavy metals from a metaliferous water solution by Typha latifolia plants and sewage sludge compost. Chemosphere 53 (5), 487–494. Matheson, L.J., Tratnyek, P.G., 1994. Reductive dehalogenation of chlorinated methanes by iron metal. Environ. Sci. Technol. 28 (12), 2045–2053. Mayer, K.U., 1999. A Numerical Model for Multicomponent Reactive Transport in Variably Saturated Porous Media. PhD Thesis, Department of Earth Sciences, University of Waterloo, Waterloo, Ont., Canada. 282pp. McDonald, D.G., Grandt, A.F., 1981. Limestone-Lime Treatment of Acid Mine Drainage- Full Scale. EPA Project Summary, EPA-600/S7-81-033, Us Environmental Protection Agency Office of Research and Development, 5pp. McGovern, T., Guerin, T.F., Horner, S., Davey, B., 2002. Design, construction and operation of a funnel and gate in situ permeable reactive barrier for remediation of petroleum hydrocarbons in groundwater. Water Air Soil Pollut. 136, 11–31. McMahon, P.B., Dennehy, K.F., Sandstrom, M.W., 1999. Hydraulic and geochemical performance of a permeable reactive barrier containing zero-valent iron. Denver Federal Center. Ground Water 37 (3), 396–404. Misaelides, P., Godelitsas, A., Filippidis, A., Charistos, D., Anousis, I., 1995. Thorium and uranium uptake by natural zeolitic materials. Sci. Total Environ. 173–174 (95), 237–246. Moon, H.S., Shin, D.Y., Nam, K., Kim, J.Y., 2008. A long-term performance test on an autotrophic denitrification column for application as a permeable reactive barrier. Chemosphere 73 (5), 723–728. Moraci, N., Calabrò, P.S., 2010. Heavy metals removal and hydraulic performance in zero-valent iron/pumice permeable reactive barriers. J. Environ. Manage. 91 (11), 2336–2341. Morkin, M., Devlin, J., Barker, J., Butler, B., 2000. In situ sequential treatment of a mixed contaminant plume. J. Contam. Hydrol. 45 (3–4), 283–302. Morris, K.A., 2009. Applications of mulch biowalls-three case studies. Remediat. J. 20, 59–68. Morrison, S.J., Spangler, R.R., 1992. Extraction of uranium and molybdenum from aqueous solutions: a survey of industrial materials for use in chemical barriers for uranium mill tailings remediation. Environ. Sci. Technol. 26 (10), 1922–1931. Morrison, S.J., Metzler, D.R., Carpenter, C.E., 2001. Uranium precipitation in a permeable reactive barrier by progressive irreversible dissolution of zerovalent iron. Environ. Sci. Technol. 35 (2), 385–390. Morrison, S.J., Mushovic, P.S., Niesen, P.L., 2006. Early breakthrough of molybdenum and uranium in a permeable reactive barrier. Environ. Sci. Technol. 40, 2018– 2024. Muegge, J., 2008. An Assessment of Zero Valence Iron Permeable-Reactive Barrier Projects In California. Office of Pollution Prevention and Green Technology, Document No. 1219, California, 147pp. Mueller, J., MacFabe, S., Vogan, J., Duchene, M., Hill, D., Bolanos-Shaw, K., Seech, A., 2004. Reductive dechlorination of solvents in groundwater using controlled release carbon with microscale ZVI. In: Proceedings Battelle’s Fourth International Conference Remediation of Chlorinated and Recalcitrant Organics. Battelle Press, Monterey, CA. p. 8. Munro, L., Clark, M., McConchie, D., 2004. A Bauxsol™-based permeable reactive barrier for the treatment of acid rock drainage. Mine Water Environ. 23 (4), 183–194. Naftz, D.L., Fuller, C.C., Davis, J.A., Morrison, S.J., Feltcorn, E.M., Freethey, G.W., Rowland, R.C., Wilkowske, C., Piana, M.J., 2002. Field demonstration of three permeable reactive barriers to control uranium contamination in groundwater, fry canyon, Utah. In: Naftz, D.L., Morrison, S.J., Davis, J.A., Fuller, C.C. (Eds.), Handbook of Groundwater Remediation Using Permeable Reactive Barriers: Applications to Radionuclides, Trace Metals, and Nutrients. Elsevier Science, San Diego, CA, pp. 402–435. Nakagawa, K., Mukai, S.R., Suzuki, T., Tamon, H., 2003. Gas adsorption on activated carbons from PET mixtures with a metal salt. Carbon 41 (4), 823–831. Nguyen, M.L., Tanner, C.C., 1998. Ammonium removal from wastewaters using natural New Zealand zeolites. New Zeal. J. Agric. Res. 41 (3), 427–446. Noubactep, C., 2010. The fundamental mechanism of aqueous contaminant removal by metallic iron. Water SA 36 (5), 663–670.

258

F. Obiri-Nyarko et al. / Chemosphere 111 (2014) 243–259

NTUA, National Technical University of Athens, 2000. Literature Review: Reactive Materials and Attenuation Processes for Permeable Reactive Barriers. Long-term Performance of Permeable Reactive Barriers used for the Remediation of Contaminated Groundwater. 5th Framework Programme Research and Technological Development Project, Athens, Greece. O’Hannesin, S.F., Gillham, R.W., 1998. Long-term performance of an in situ ‘‘Iron Wall’’ for remediation of VOCs. Ground Water 36 (1), 164–170. Orth, W.S., Gillham, R.W., 1996. Dechlorination of trichloroethene in aqueous solution using Fe0. Environ. Sci. Technol. 30 (1), 66–71. Ott, N., 2000. Permeable Reactive Barriers for Inorganics. U.S. Environmental Protection Agency, Washington, DC. 55pp. Öztürk, Z., Tansel, B., Katsenovich, Y., Sukop, M., Laha, S., 2012. Highly organic natural media as permeable reactive barriers: TCE partitioning and anaerobic degradation profile in eucalyptus mulch and compost. Chemosphere 89 (6), 665–671. Pang, F.M., Teng, S.P., Teng, T.T., Mohd Omar, A.K., 2009. Heavy metals removal by hydroxide precipitation and coagulation–flocculation methods from aqueous solutions. Water Qual. Res. J. Canada 44 (2), 174–182. Panturu, E., Filcencu-Olteanu, A., Groza, N., Panturu, R.I., Ciobanu, L., 2009. Uranium immobilization on reactive material using RPB. Chem. Bull. ‘‘POLITEHNICA’’ 54(68), 50–53. Park, J.-B., Lee, S.-H., Lee, J.-W., Lee, C.-Y., 2002. Lab scale experiments for permeable reactive barriers against contaminated groundwater with ammonium and heavy metals using clinoptilolite (01-29B). J. Hazard. Mater. 95 (1–2), 65–79. Parkhurst, D.L., Appelo, C.A.J., 1999. User’s guide to PHREEQC (Version 2): a computer program for speciation, reaction-path, advective-transport, and inverse geochemical calculations. U.S. Geological Survey, Denver, CO. 309pp. Paul, C.J., 2002. Sodium Dithionite Injections used for Chromium Reduction. U.S. Environmental Protection Agency, Washington, DC. Peng, X., Li, Y., Luan, Z., Di, Z., Wang, H., Tian, B., Jia, Z., 2003. Adsorption of 1,2dichlorobenzene from water to carbon nanotubes. Chem. Phys. Lett. 376 (1–2), 154–158. Peric´, J., Trgo, M., Vukojevic´ Medvidovic´, N., 2004. Removal of zinc, copper and lead by natural zeolite-a comparison of adsorption isotherms. Water Res. 38 (7), 1893–1899. Phillips, D.H., 2010. Permeable reactive barriers: a sustainable technology for cleaning contaminated groundwater in developing countries. Desalination 251, 352–359. Phillips, J.B., Love, N.G., 1998. Biological denitrification using upflow biofiltration in recirculating aquaculture systems: pilot-scale experience and implications for full-scale. In: Proceedings of the 2nd International Conference on Recirculating Aquaculture, Roanoke, Virginia, pp. 171–178. Phillips, D.H., Gu, B., Watson, D.B., Roh, Y., Liang, L., Lee, S.Y., 2000. Performance evaluation of a zerovalent iron reactive barrier: mineralogical characteristics. Environ. Sci. Technol. 34 (19), 4169–4176. Phillips, D.H., Van Nooten, T., Bastiaens, L., Russell, M.I., Dickson, K., Plant, S., Ahad, J.M.E., Newton, T., Elliot, T., Kalin, R.M., 2010. Ten year performance evaluation of a field-scale zero-valent iron permeable reactive barrier installed to remediate trichloroethene contaminated groundwater. Environ. Sci. Technol. 44 (10), 3861–3869. Powell, R.M., Blowes, D.W., Gillham, R.W., Schultz, D., Sivavec, T., Puls, R.W., Vogan, J.L., Powell, P.D., Landis, R., 1998. Permeable Reactive Barrier Technologies for Contaminant Remediation. Report EPA/600/R-98/125, U. S. Environmental Protection Agency, Washington, DC, 51pp. Puls, R.W., 2006. Long-term performance of permeable reactive barriers: lessons learned on design, contaminant treatment, longevity, performance monitoring and cost – an overview. In: Twardowska, I. et al. (Eds.), Soil and Water Pollution Monitoring. Protection and Remediation, Springer, Dordrecht, The Netherlands, pp. 221–229. Puls, R.W., Blowes, D.W., Gillham, R.W., 1999. Long-term performance monitoring for a permeable reactive barrier at the U.S. Coast Guard Support Center, Elizabeth City, North Carolina. J. Hazard. Mater. 68 (5), 109–124. Ranck, J.M., Bowman, R.S., Weeber, J.L., Katz, L.E., Sullivan, E.J., 2005. BTEX removal from produced water using surfactant-modified zeolite. J. Environ. Eng. 131 (3), 434–442. Rastogi, G., Sani, R., 2011. Molecular techniques to assess microbial community structure, function, and dynamics in the environment. In: Ahmad, I., Ahmad, F., Pichtel, J. (Eds.), Microbes and Microbial Technology SE – 2. Springer, New York, NY, pp. 29–57. Roberts, A.L., Totten, L.A., Arnold, W.A., Burris, D.R., Campbell, T.J., 1996. Reductive elimination of chlorinated ethylenes by zero-valent metals. Environ. Sci. Technol. 30 (8), 2654–2659. Roberts, L.A., Ball, W.P., Searson, P., Fairbrother, H., Vikesland, P.J., Klausen, J., Kohn, T., Kamath, R., Zimmermann, H.J., 2002. Influence of Groundwater Constituents on Longevity of Iron-Based Permeable Barriers. Final Report for the SERDP Project CU-1125, Air Force Research Laboratory, Tyndall AFB, FL. Robertson, W.D., Cherry, J.A., 1995. In situ denitrification of septic-system nitrate using reactive porous media barriers: field trials. Ground Water 33 (1), 99–111. Robertson, W.D., Blowes, D.W., Ptacek, C.J., Cherry, J.A., 2000. Long-term performance of in situ reactive barriers for nitrate remediation. Ground Water 38 (5), 689–695. Robertson, W.D., Vogan, J.L., Lombardo, P.S., 2008. Nitrate removal rates in a 15year-old permeable reactive barrier treating septic system nitrate. Ground Water Monit. Remediat. 28 (3), 65–72.

Robertson, W.D., Ptacek, C.J., Brown, S.J., 2009. Rates of nitrate and perchlorate removal in a 5-Year-old wood particle reactor treating agricultural drainage. Ground Water Monit. Remediat. 29 (2), 87–94. Rodak, C., Silliman, S.E., Bolster, D., 2014. Time-dependent health risk from contaminated groundwater including use of reliability, resilience, and vulnerability as measures. JAWRA J. Am. Water Resour. Assoc. 50 (1), 14–28. Saberi, A., 2012. Comparison of Pb removal efficiency by zero valent iron 2+ nanoparticles and Ni/Fe bimetallic nanoparticles. J. Energy Environ. 3 (2), 189– 196. Sayles, G.D., You, G., Wang, M., Kupferle, M.J., 1997. DDT, DDD, and DDE dechlorination by zero-valent iron. Environ. Eng. Sci. 31 (12), 3448–3454. Scherer, M.M., Richter, S., Valentine, R.L., Alvarez, P.J., 2000. Chemistry and microbiology of permeable reactive barriers for in situ groundwater clean up. Crit. Rev. Microbiol. 26 (4), 221–264. Schipper, L.A., McGill, A., 2008. Nitrogen transformation in a denitrification layer irrigated with dairy factory effluent. Water Res. 42 (10–11), 2457–2464. Schipper, L.A., Robertson, W.D., Gold, A.J., Jaynes, D.B., Cameron, S.C., 2010. Denitrifying bioreactors – an approach for reducing nitrate loads to receiving waters. Ecol. Eng. 36 (11), 1532–1543. Seelsaen, N., McLaughlan, R., Moore, S., Ball, J.E., Stuetz, R.M., 2006. Pollutant removal efficiency of alternative filtration media in stormwater treatment. Water Sci. Technol. 54 (6–7), 299–305. Sethi, R., Day, S., Di Molfetta, A., 2011. Clamshell vs. backhoe excavation of permeable reactive barriers. Am. J. Environ. Sci. 7 (5), 463–467. Simantiraki, F., Maratos, D., Anastasiadou, K., Gidarakos, E., 2012. Characterization of municipal solid waste compost and use for groundwater remediation of petroleum hydrocarbons. In: Protection and Restoration of the Environment XI Conference, Thessaloniki, Greece, pp. 413–422. Simon, F., Meggyes, T., 2000. Removal of organic and inorganic pollutants from groundwater using permeable reactive barriers Part 1. Treatment processes for pollutants. Land Cont. Recl. 8 (2), 103–117. Singh, B., Wilson, M.J., McHardy, W.J., Fraser, A.R., Merrington, G., 1999. Mineralogy and chemistry of ochre sediments from an acid mine drainage near a disused mine in Cornwall, UK. Clay Miner. 34 (2), 301–317. Skinner, S.J., Schutte, C.F., 2006. The feasibility of a permeable reactive barrier to treat acidic sulphate- and nitrate-contaminated groundwater. Water SA 32 (2), 129–136. Song, H., Carraway, E.R., Kim, Y.H., Batchelor, B., Jeon, B.-H., Kim, J., 2008. Amendment of hydroxyapatite in reduction of tetrachloroethylene by zerovalent zinc: its rate enhancing effect and removal of Zn(II). Chemosphere 73 (9), 1420–1427. Starr, R.C., Cherry, J.A., 1994. In situ remediation of contaminated ground water: the funnel-and-gate system. Ground Water 32 (3), 465–476. Stevens, C.J., Quinton, J.N., 2009. Diffuse pollution swapping in arable agricultural systems. Crit. Rev. Environ. Sci. Technol. 39 (6), 478–520. Su, C., Puls, R.W., 2001. Arsenate and arsenite removal by zerovalent iron: kinetics, redox transformation, and implications for in situ groundwater remediation. Environ. Sci. Technol. 35 (7), 1487–1492. Su, C., Puls, R.W., 2004. Significance of iron(II, III) hydroxycarbonate green rust in arsenic remediation using zerovalent iron in laboratory column tests. Environ. Sci. Technol. 38 (19), 5224–5231. Sun, H., Wang, L., Zhang, R., Sui, J., Xu, G., 2006. Treatment of groundwater polluted by arsenic compounds by zero valent iron. J. Hazard. Mater. 129 (1–3), 297–303. Takeuchi, Y., Suzuki, T., Arai, H., 1988. A study of equilibrium and mass transfer in processes for removal of heavy-metal ions by hydroxyapatite. J. Chem. Eng. Jpn. 21 (1), 98–100. Tase, N., 1992. Groundwater contamination in Japan. Environ. Geol. Water Sci. 20 (1), 15–20. Testa, S.M., Winegardner, D.L., 2000. Restoration of Contaminated Aquifers: Petroleum Hydrocarbons and Organic Compounds, 2nd Edition. Thiruvenkatachari, R., Vigneswaran, S., Naidu, R., 2008. Permeable reactive barrier for groundwater remediation. J. Ind. Eng. Chem. 14 (2), 145–156. Tobiszewski, M., Namies´nik, J., 2012. Abiotic degradation of chlorinated ethanes and ethenes in water. Environ. Sci. Pollut. Res. 19 (6), 1994–2006. Tratnyek, P.G., 2002. Keeping up with all that literature: the ironrefs database turns 500. Ground Water Monit. R. 22 (3), 92–94. Tratnyek, P.G., Scherer, M.M., Johnson, T.L., Matheson, L.J., 2003. Permeable reactive barriers of iron and other zero-valent metals. In: Tarr, M.A. (Ed.), Chemical Degradation Methods for Wastes and Pollutants: Environmental and Industrial Applications. CRC Press, New York, pp. 371–421. Triay, I.R., Meijer, A., Cisneros, M.R., Miller, G.G., Mitchell, A.J., Ott, M.A., Hobart, D.E., Palmer, P.D., Perrin, R.E., Aguilar, R.D., 1989. Sorption of Americium in Tuff and Pure Minerals Using Synthetic and Natural Ground waters. In: Proceedings Migration ‘89: 2nd International Conference on Chemistry and Migration Behavior of Actinides and Fission Products in the Geosphere, Monterrey, CA, p. 13. U.S. EPA, 1995. In Situ Remediation Technology Status Report: Treatment Walls. EPA542-K-94-004, U.S. Environmental Protection Agency, Office of Solid Waste and Emergency Response, Washington, DC, 26pp. U.S. EPA, 1997. Permeable Reactive Subsurface Barriers for the Interception and Remediation of Chlorinated Hydrocarbon and Chromium (VI) Plumes in Ground Water. EPA/6ok00/F-97/008, U.S. Environmental Protection Agency, Office of Research and Development, Washington, DC, 4pp. U.S. EPA, 1998. Evaluation of Demonstrated and Emerging Technologies for the Treatment of Contaminated Land and Groundwater (Phase III) Treatment Walls

F. Obiri-Nyarko et al. / Chemosphere 111 (2014) 243–259 and Permeable Reactive Barriers. EPA 542-R-98-003. U.S. Environmental Protection Agency, Vienna, Austria. 101pp.. U.S. EPA, 2002. Field Applications of In Situ Remediation Technologies: Permeable Reactive Barriers. U.S. Environmental Protection Agency, Office of Solid Waste and Emergency Response, Washington, DC. Van Wazer, J.R., 1974. Phosphorus and Its Compounds. John Wiley & Sons Canada, Limited. van-Nooten, T., Diels, L., Bastiaens, L., 2010. Design of a multibarrier for the treatment of landfill leachate contamination: laboratory column evaluation. In: Bastiaens, L. (Ed.), 4th International Symposium: Permeable reactive Barriers and Reactive Zones, Vito, Antwerp, Belgium, p. 34. Vesela, L., Nemecek, J., Siglova, M., Kubal, M., 2006. The biofiltration permeable reactive barrier: practical experience from Synthesia. Int. Biodeterior. Biodegr. 58 (3–4), 224–230. Vidic, R.D., Pohland, F.G., 1996. Treatment Walls. Technology Evaluation Report TE96-01, Ground-Water Remediation Technologies Analysis Center, Pittsburg, PA., 34pp. Vogan, J.L., 1993. The Use of Emplaced Denitrifying Layers to Promote Nitrate Removal from Septic Effluent. University of Waterloo, Waterloo, Canada. Vogan, J.L., Focht, R.M., Clark, D.K., Graham, S.L., 1999. Performance evaluation of a permeable reactive barrier for remediation of dissolved chlorinated solvents in groundwater. J. Hazard. Mater. 68 (1–2), 97–108. Warner, S.D., Sorel, D., 2002. Chlorinated solvent and DNAPL remediation. In: Henry, S.M., Warner, S.D. (Eds.), American Chemical Society, Washington, DC. Waybrant, K.R., Ptacek, C.J., Blowes, D.W., 2002. Treatment of mine drainage using permeable reactive barriers: column experiments. Environ. Sci. Technol. 36 (6), 1349–1356. Weelink, S.B., Eekert, M.A., Stams, A.M., 2010. Degradation of BTEX by anaerobic bacteria: physiology and application. Rev. Environ. Sci. Bio/Technol. 9, 359–385. Wei, Z., Seo, Y., 2010. Trichloroethylene (TCE) adsorption using sustainable organic mulch. J. Hazard. Mater. 181, 147–153. Westerhoff, P., James, J., 2003. Nitrate removal in zero-valent iron packed columns. Water Res. 37 (8), 1818–1830. Wiafe, S., Ofosu, S.A., Ahima, J., 2013. The quality of groundwater resources around auto-mechanic workshop enclaves in Ghana. Eng. Sci. Technol. 1 (1), 38–49. Wilkin, R.T., McNeil, M.S., 2003. Laboratory evaluation of zero-valent iron to treat water impacted by acid mine drainage. Chemosphere 53 (7), 715–725. Wilkin, R.T., Puls, R.W., 2003. Capstone Report on the Application, Monitoring, and Performance of Permeable Reactive Barriers for Ground-Water Remediation: Volume 1 Performance Evaluations at Two Sites. Report EPA/600/R-03/045a, U.S. Environmental Protection Agency, National Risk Management Research Laboratory, Cincinnati, OH, 135pp. Wilkin, R.T., Puls, R.W., Sewell, G.W., 2000. Long-term Performance of Permeable Reactive Barriers Using Zero-valent Iron: An Evaluation at Two Sites. U.S. Environmental Protection Agency, Washington, DC. 19pp. Wilkin, R.T., Acree, S.D., Beak, D.G., Ross, R.R., Lee, T.R., Paul, C.J., 2008. Field Application of a Permeable Reactive Barrier for Treatment of Arsenic in Ground

259

Water. Report EPA 600/R-08/093, U.S. Environmental Protection Agency, Office of Research and Development, Washington, DC, 65pp. Wilson, R.D., Mackay, D.M., Scow, K.M., 2001. In situ MTBE biodegradation supported by diffusive oxygen release. Environ. Sci. Technol. 36 (2), 190–199. Wilson, J.T., He, Y., Henry, B., Evans, P., 2010. Abiotic and biotic transformation of TCE under sulfate reducing conditions: the role of spatial heterogeneity. In: Presented at 7th International Battelle Conference, Monterey, CA, May 24–27. Xu, Y., Schwartz, F.W., 1994. Lead immobilization by hydroxyapatite in aqueous solutions. J. Contam. Hydrol. 15 (3). 206-187. Yang, X., Fan, L.T., Erickson, L.E., 1995. A conceptual study on the bio-wall technology: feasibility and process design. Remed. J. 6 (1), 55–67. Yang, J.E., Kim, J.S., Ok, Y.S., Yoo, K.R., 2007. Mechanistic evidence and efficiency of the Cr(VI) reduction in water by different sources of zerovalent irons. Water Sci. Technol. 55 (1–2), 197–202. Yang, J., Cao, L., Guo, R., Jia, J., 2010a. Permeable reactive barrier of surface hydrophobic granular activated carbon coupled with elemental iron for the removal of 2,4-dichlorophenol in water. J. Hazard. Mater. 184 (1–3), 782–787. Yang, S.-C., Lei, M., Chen, T.-B., Li, X.-Y., Liang, Q., Ma, C., 2010b. Application of zerovalent iron (Fe(0)) to enhance degradation of HCHs and DDX in soil from a former organochlorine pesticides manufacturing plant. Chemosphere 79 (7), 727–732. Yeh, C.-H., Lin, C.-W., Wu, C.-H., 2010. A permeable reactive barrier for the bioremediation of BTEX-contaminated groundwater: microbial community distribution and removal efficiencies. J. Hazard. Mater. 178 (1–3), 74–80. Yerushalmi, L., Manuel, M., Guiot, S., 1999. Biodegradation of gasoline and BTEX in a microaerophilic biobarrier. Biodegradation 10 (5), 341–352. Zawierucha, I., Kozlowski, C., Malina, G., 2012. Heavy metals removal from contaminated groundwater using permeable reactive barriers with immobilized membranes. In: Małoszewski, P., Witczak, S., Malina, G. (Eds.), Groundwater Quality Sustainability. CRC Press, London, pp. 79–88. Zawierucha, I., Kozlowski, C., Malina, G., 2013. Removal of toxic metal ions from landfill leachate by complementary sorption and transport across polymer inclusion membranes. Waste Manage. 33, 2129–2136. Zhang, Y., Amrhein, C., Frankenberger, W.T., 2005. Effect of arsenate and molybdate on removal of selenate from an aqueous solution by zero-valent iron. Sci. Total Environ. 350 (1–3), 1–11. Zhu, K., Qiu, J., Ji, H., Yanagisawa, K., Shimanouchi, R., Onda, A., Kajiyoshi, K., 2010. Crystallographic study of lead-substituted hydroxyapatite synthesized by hightemperature mixing method under hydrothermal conditions. Inorganica Chim. Acta 363 (8), 1785–1790. Zolla, V., Sethi, R., Molfetta, A.Di., 2007. Performance assessment and monitoring of a permeable reactive barrier for the remediation of a contaminated site. Am. J. Environ. Sci. 3 (3), 158–165. Zolla, V., Freyria, F.S., Sethi, R., Di Molfetta, A., 2009. Hydrogeochemical and biological processes affecting the long-term performance of an iron-based permeable reactive barrier. J. Environ. Qual. 38 (3), 897–908.