Scholarly article on topic 'Opportunities for groundwater microbial electro-remediation'

Opportunities for groundwater microbial electro-remediation Academic research paper on "Environmental biotechnology"

Share paper
Academic journal
Microbial Biotechnology
OECD Field of science

Academic research paper on topic "Opportunities for groundwater microbial electro-remediation"

Opportunities for groundwater microbial electro-remediation

Narcís Pous, ^ Maria Dolors Balaguer, Jesús Colprim ^ and Sebastia Puig* 0

Laboratory of Chemical and Environmental Engineering (LEQUiA), Institute of the Environment, University of Girona, Campus Montilivi, Carrer Maria Aurelia Capmany, 69, E-17003 Girona, Spain.


Groundwater pollution is a serious worldwide concern. Aromatic compounds, chlorinated hydrocarbons, metals and nutrients among others can be widely found in different aquifers all over the world. However, there is a lack of sustainable technologies able to treat these kinds of compounds. Microbial electro-remediation, by the means of microbial electrochemical technologies (MET), can become a promising alternative in the near future. MET can be applied for groundwater treatment in situ or ex situ, as well as for monitoring the chemical state or the microbiological activity. This document reviews the current knowledge achieved on microbial electro-remediation of groundwater and its applications.

Opportunities for microbial electrochemical technologies in groundwater treatment

Groundwater is one of the main sources of drinking water all over the world. However, its usage as drinking water is threatened by the presence of different pollutants that have reached the aquifer due to anthropogenic or geologic sources (Katsoyiannis et al., 2007; Bohlke et al., 2009; Van Halem et al., 2009; Sprague et al., 2011). The pollutants can be accumulated in the aquifer by the lack of a suitable electron donor/acceptor. But

Received 30 May, 2017; revised 4 September, 2017; accepted 5 September, 2017.

*For correspondence. E-mail; Tel. +34972418182; Fax +34972418150. Microbial Biotechnology (2017) 0(0), 000-000 doi:10.1111/1751-7915.12866 Funding Information

This research was financially supported by the Spanish Government (CTQ2014-53718-R and CTM2015-71982-REDT) and the University of Girona (MPCUdG2016/137). LEQUIA has been recognized as a consolidated research group by the Catalan Government with code 2014-SGR-1168.

they need to be removed because of a further usage of drinking water or by means of environmental sustainabil-ity. The most used strategies are based on pollutant separation (membrane technologies or ion exchange) or external addition of chemicals for abiotic or biologic catalysis [e.g. organic matter for treating nitrates (McA-dam and Judd, 2006) or oxygen/nitrate for treating hydrocarbons (Bamforth and Singleton, 2005)]. However, these technologies present some drawbacks. On the one hand, separation-based technologies have a high energy cost and they concentrate the pollutant into a waste brine of difficult disposal (Twomey et al., 2010). On the other hand, the application of traditional remediating strategies that requires external chemical addition in situ or ex situ is limited by (i) undesired side reactions, (ii) poor chemical distribution (in situ strategies) and (iii) the addition of some chemicals can have collateral damages (e.g. organic matter addition can generate sludge that needs to be removed). Therefore, new sustainable strategies can have a golden opportunity on groundwater bioremediation if they have (i) low cost; (ii) no/low chemical consumption; and (iii) non-invasive and selective electron donor/acceptor dosing. These characteristics can be met in microbial electrochemical technologies (MET), which are an emerging technology platform where microbiology meets electrochemistry (Schröder et al., 2015). In this technological approach, electroactive bacteria are able to use a solid electrode as electron donor or electron acceptor (Rabaey et al., 2009). The electrode is the alternative to oxygen/nitrate as electron acceptor (Bamforth and Singleton, 2005; Oremland and Stolz, 2005), or organic matter/hydrogen as electron donor (McAdam and Judd, 2006; Karanasios et al., 2010). Depending on the pollutant and groundwater's characteristics, a MET system can be operated as a microbial fuel cell (MFC) or as a microbial electrolysis cell (MEC) (Schröder et al., 2015). MFC is an autonomous device from where energy can be extracted, while a MEC is a device where energy is supplied to allow/enhance a bioelectrochemical process.

Different commercial opportunities can be found for microbial electro-remediation of contaminated groundwa-ter (Fig. 1). The most studied application is the ex situ treatment. Through this strategy, groundwater has to be pumped to the treatment plant (either a permanent/offsite or a movable/on-site plant), where an intensive treatment is applied for a fast contaminant removal. The

© 2017 The Authors. Microbial Biotechnology published by John Wiley & Sons Ltd and Society for Applied Microbiology.

This is an open access article under the terms of the Creative Commons Attribution License, which permits use, distribution and

reproduction in any medium, provided the original work is properly cited.

faster the treatment is, the smaller the plant volume is needed (with the corresponding decrease in the capital cost). The treated water can be either used for human purposes (i.e. drinking water) or re-injected into the aquifer (i.e. to avoid salinity intrusion or to control the phrea-tic level). However, these ex situ strategies might not be recommended for some applications. In some scenarios, an in situ MET that allows the treatment and immobilization of the contaminant in the subsurface might be more suitable. For example, in aquifers with geochemical U (VI) solubilization (Williams et al., 2013), its in situ conversion into an insoluble form U(IV) and consequent immobilization in the subsoil might be preferred. For an in situ MET implementation, a less intensive treatment can be applied. Electrodes can be directly introduced in the aquifer to stimulate the native microorganisms and, in consequence, to accelerate the aquifer bioremediation (Gregory and Lovley, 2005). An approach can be followed similar to already available technologies like elec-trokinetic remediation (Acar et al., 1995) or vitrification (Mulligan et al., 2001).

Last but not least, small MET systems can also be used as biosensors to monitor the microbial activity in the aquifer (Williams et al., 2010; Wardman et al., 2014) or to evaluate its chemical state (Feng et al., 2013; Webster et al., 2014; Velasquez-Orta et al., 2017).

Considering these different MET configurations, the treatment of different groundwater pollutants has been investigated using a solid electrode either as electron sink [e.g. for oxidation of aromatic hydrocarbons (Zhang et al., 2010; Friman et al., 2013) or dissolved metals (Pous et al., 2015a)] or as electron donor (e.g. for reduction of

nitrates (Pous et al., 2013; Zhang and Angelidaki, 2013), metals (Gregory and Lovley, 2005) or chlorinated hydrocarbons (Aulenta et al., 2007). This review explores the MET platform for groundwater bioremediation.

Organic contaminants

Aromatic compounds

The presence of aromatic compounds in groundwater is mainly attributed to anthropogenic contamination, mostly derived from petrochemical activities (Turney and Goerl-itz, 1990; Teuten et al., 2009). With a lack of electron acceptors, these substances can remain in the environment for a long time. Although the presence of aromatic hydrocarbons is usually found at low concentrations [ig r1 level (Rakoczy et al., 2013)], they are already toxic at these levels. For example, the guideline value for nitrobenzene in drinking water is 17 ig l_1 in the United States (Environmental Protection Agency (EPA), 2004). Therefore, a highly specific and effective treatment for this kind of compounds is needed, which can be difficult to achieve by conventional biologic treatments. The versatility of MET in terms of operational mode [microbial fuel cell (MFC), microbial electrolysis cell with 2- or 3-elec-trode configuration (2-MEC, 3-MEC)], working electrode potential or active microbiome allows a plethora of aromatic contaminants to be treated (Table 1 and Fig. 2). Both MFC and MEC have proved to be an effective strategy to catalyse either polycyclic [e.g. phenantrene or naphthalene (Zhang et al., 2010; Yan and Reible, 2015)] or monocyclic aromatic compounds [e.g. benzene, phenol or nitrobenzene (Mu et al., 2009; Zhang et al., 2010;

Fig. 1. Framework of opportunities for microbial electrochemical technologies in groundwater.

tiaHE §w

ed po O E

-Q O --:

R al , t

R t ,

-a ®TO 0 tO^ a r-

ahuo ■ N > cc

bc I OO I

-M -M -M

2- 3- 3-

oo = ?? ulf

eeS nn

® a I

nhd epo haz Q-Z< I

-^a-z <

ed § §

&& ■P E

S o O o

£ «5

£ E Do

CT> ® g> g ®

„ioo^S ^

™ TOr TOr ^ ®

., t t ae

al t t nu CO C

at -G z-

t nn b

E ' eo I Q

OOi-i-0000 2222

t t t t tl., t tOtOtOtO — TO t0

__r!,JC\IC\l to at 2

O = cci

.jo .S ti

ib ac o

®to o^o

® a.c o

I LUO (3 co O

tasil ta

■S ^ o 'S aa

■Q 'C

o a <B ST

, to p. o

co p i ., on r

y !± P ! "> o ! = C

O <D s ¡to« E

' -Q& O 'O.O -Q

a alee nt ooh iffi iffi O

5 -(5 ® '<5 & ALLLS

¡n "> ar

C CD £ "


c c lo al

oi p cs

<« 5=

■c o

I Q 0Q

to -c ce


o S ■9 S

to to ha en Q <c

00/7 09 44

72 92 43

00 00 35

ocooooooRRo ommmmmmfnojO



-M -M -M -M -M -M -M -M -M -M -M -M -M

2- 3- 3- M M 3- 3- M 3- 3- 3- 3- 3- 3- 3- 3-

3 3 3 .3 .3

'(A « «


.3 .3 3 3

Uj Uj Uj UU

3 .3 .3 .3

Uj Uu Uj Uj


Cl C +

+ e O e

e oi e

th O th

Et C Et

-D -D -D

'CJ 1,

.a c o ®

orE ^ y is- ci

Friman et al., 2013; Rakoczy et al., 2013)]. Although electricity can be harvested from an MFC, it requires of stable and reliable counterelectrode reaction, which implies extramaintenance and surveillance. Thus, MEC operation might be preferred for bioremediation, as it allows focusing on aromatic compounds removal. Moreover, if an electrode control strategy is chosen (3-MEC), a better control of the reaction and the remediation rates could be reached.

Like in other MET applications, this field of research became wider after it was found that Geobacter was able to oxidize aromatic compounds using an electrode as electron acceptor (Zhang et al., 2010). The ability of Geobacter sp. on dealing with aromatics can accelerate MET application in this field, as it is a well-known and well-studied model electroactive genus (Lovley et al., 2011). In fact, Geobacter species had already been detected in aquifers contaminated with aromatic compounds (Rooney-Varga et al., 1999), thus suggesting their capacity to anaerobically oxidize aromatic hydrocarbons. Some years later, anaerobic benzene oxidation by Geobacter was successfully proven (Zhang et al., 2012). And even the genes for anaerobic benzene oxidation were identified for Geobacter metallire-ducens (Zhang et al., 2014). Thus, it was not surprising that one of the first MET experiences on aromatics removal evaluated the ability of Geobacter metallire-ducens to oxidize toluene, benzene or naphthalene to carbon dioxide using a graphite electrode as electron sink (Zhang et al., 2010). Nevertheless, oxidation of anaerobic aromatic hydrocarbons using MET is not an easy task. Only one pure microorganism not belonging to Geobacter genus has been reported to be able to oxidize phenol using an electrode as electron acceptor, Cupriavidis basilensis (Friman et al., 2013). The study of pure cultures is important for their understanding, but for real groundwater bioremediation applications, the usage of pure cultures might not be feasible. Then, the usage of mixed cultures gains interest. One of the most successful examples is the treatment of benzene. Benzene remediation has been successfully applied for either ex situ (Rakoczy et al., 2013; Wei et al., 2015a, b) or in situ experiences using mixed communities (Chang et al., 2016). An interesting finding was described by Rakoczy et al. (2013). The authors proved the simultaneous oxidation of sulfide and benzene in an anode mostly dominated by S-proteobacteria (31%). Iso-topic analyses revealed that small amounts of oxygen might be

required to activate the benzene oxidation in their system (Rakoczy et al., 2013). Thus, in real practical applications for aromatic hydrocarbon treatment, a positive coexistence of different microbial metabolisms is expected to happen.

The complexity of aromatics can increase with the presence of N- or S-functional groups, leading to the need for developing different strategies for their treatment. One example of functionalized aromatics treatment in MET is nitrobenzene degradation. In the ideal case, nitrobenzene would be converted into CO2 and NH^. However, nitrobenzene complexity makes this task hard, and its solely reduction into aniline can be already seen as a success (Mu et al., 2009; Wang et al., 2011; Yun et al., 2017). In fact, nitrobenzene reduction to aniline already reduces the water toxicity. Following a similar strategy, toxicity reduction instead of full oxidation, METs have been used for azo dye orange 7 reduction into sul-fanilic acid (Yun et al., 2017) and the toxicity of waters containing dibenzothiophene or atrazine has also been decreased (Rodrigo et al., 2014; Dominguez-Garay et al., 2016). Atrazine is an interesting example, as it has been successfully mineralized (Dominguez-Garay et al., 2017). This example shows the potential of METs over the treatment of complex aromatic compounds.

In conclusion, a big window of opportunities can be opened for microbial electro-remediation of aromatic hydrocarbons, as METs are capable to treat not only homoaromatic hydrocarbons, but also those containing N- or S-functional groups or heteroaromatic hydrocarbons. Nevertheless, there are still relevant challenges to be addressed. As a general overview of microbial electro-remediation, there is a lack of experiences at pilot-scale level, which is also occurring in the field of aromatic compounds removal. In this case, as aromatics contamination is mostly derived from petrochemical activities (Turney and Goerlitz, 1990; Teuten et al., 2009), the most appropriate strategy would be in situ bioremediation, but field testing is still scarce (Daghio et al., 2017). This lack of experience is relevant for aro-matics bioremediation, as more hurdles are expected to be found when moving to the field. For example, in a real petrochemical spill, there are several polyaromatic hydrocarbons species, some of which might not be bioavailable for bacteria due to its high hydrophobicity, and some others might also be toxic (Bamforth and Singleton, 2005). Nevertheless, laboratory testing is still needed to find the catalytic routes. Toluene, benzene or naphthalene has been already successfully converted into carbon dioxide (Zhang et al., 2010), but when treating more complex compounds such as nitrobenzene, azo dye or dibenzothiophene, the conversion to carbon dioxide could not be reached.

Chlorinated hydrocarbons

Chlorinated hydrocarbons can be found in groundwater at ppb level due to solvent spills that have leaked into the aquifer (Squillace et al., 2002; Moran et al., 2007).

Chlorinated hydrocarbons have been conventionally removed from groundwater by means of separation technologies (i.e. ion exchange, reverse osmosis or nanofil-tration) (Altalyan et al., 2016) or through permeable reactive barriers (Obiri-Nyarko et al., 2014). But there is a biologic alternative to deal with these compounds: reductive dechlorination (Holliger and Schraa, 1994; Holliger et al., 1998). In the ideal scenario, it allows turning the chlorinated hydrocarbons into ethene and chloride. Following this approach, the removal of chlorinated compounds using MET platform has been widely investigated by operating the system as a MEC (Aulenta et al., 2008; Strycharz et al., 2008). Bioelectrochemical dechlorination of some aromatic hydrocarbons, like chlorophenol (Strycharz et al., 2010; Wen et al., 2013), has also been reported. However, most of the studies have been focused on the removal of chlorinated aliphatic hydrocarbons (CAHs), the occurrence of which is high in groundwater.

Tetrachloroethene/perchloroethene (PCE) reduction using a polarized cathode as electron donor has been demonstrated by either mixed cultures (Yu et al., 2016) or a pure culture (Geobacter lovley!) (Strycharz et al., 2008). The main objective is to reduce PCE into ethene (Chambon et al., 2013). However, PCE is reduced through a sequence of reactions where trichloroethene (TCE), cis-dichloroethene (cis-DCE) and vinyl chloride (VC) are stable intermediates that can be accumulated (Chambon et al., 2013). When using Geobacter lovleyi at a poised cathode potential of -300 mV versus standard hydrogen electrode (SHE), PCE was reduced at a maximum rate of around 25 imol day-1, which was similar to the values observed when using acetate as electron donor (Strycharz et al., 2008). However, PCE was only converted into cis-DCE, which is still a toxic compound and needs further degradation. Positively, when using a mixed culture at -500 mV versus SHE, PCE could be finally degraded into ethene in batch mode (Yu et al., 2016). However, a minimum of 50% of initial PCE was accumulated as vinyl chloride, indicating that further process optimization is needed.

The most studied chlorinated aliphatic hydrocarbon using MET is trichloroethene (TCE) (Aulenta et al., 2007, 2008, 2010, 2011; Verdini et al, 2015; Lai et al, 2017). From the initial proof-of-concept (Aulenta et al., 2007), research has evolved towards the understanding of the whole process [electron transfer mechanism (Aulenta et al., 2007, 2010), cis-DCE role as intermediate (Aulenta et al., 2010, 2013; Lai et al., 2015) or electron competitors such as methane generation (Aulenta et al., 2008, 2011) and nitrate/sulfate presence (Lai et al., 2015)]. Process optimization through cathode potential, mass transport or continuous-flow operation has also been evaluated (Aulenta et al., 2011; Verdini et al.,

2015; Lai et al., 2017), and it has allowed to increase the bioelectrochemical dechlorination rates from 14.222.4 ieq l-1 day-1 (Aulenta et al., 2010) to 121.8 ieq l-1 day-1 (Lai et al., 2017) in the last years. These rates are similar to values obtained in conventional reductive dechlorination (Shukla et al., 2014), which highlights the competitiveness of bioelectrochemical reductive dechlorination. However, despite ethene is the desired product of reductive dechlorination, VC has been commonly observed as the main final product (Aulenta et al., 2007, 2008, 2010, 2011). In order to solve this issue, an interesting approach where TCE is reduced to VC in the biocathode and VC is further aerobically oxidized to carbon dioxide in the anode has been successfully implemented and demonstrated (Lai et al., 2017).

The list of chlorinated aliphatic compounds treated in MET can be further extended to the successful treatment of 1,2-dichloroethane (1,2-DCA) (Leitao et al., 2015, 2016, 2017). Initially, the 1,2-DCA conversion to ethene was evaluated at different cathode potentials from -300 to -900 mV versus SHE using a Dehalococcoides-enriched microbial culture. The authors observed 1,2-DCA conversion to ethene at -300 mV versus SHE, a potential at which it was deduced that direct electron uptake was the mechanism driving this process (Leitao et al., 2015). The work was further extended by investigating the effect of supplementing an external mediator [Anthraquinone-2,6-disulfonate (AQDS)] in a biocathode polarized at -300 mV versus SHE (Leitao et al., 2016). Through AQDS addition, the 1,2-DCA dechlorination rate increased from 20 imol l-1 day-1 in the first work (Leitao et al., 2015) to 65 imol l-1 day-1 in the last one (Leitao et al., 2016). AQDS could even be immobilized on the electrode surface for an easier application (Leitaao etal., 2017).

In conclusion, the experience on bioelectrochemical reductive dechlorination is already broad in MET field. In the recent years, a positive evolution took place that allowed increasing the removal rates up to values similar to conventional reductive dechlorination and a better understanding of the underlying fundamentals of bioelec-trochemical dechlorination was obtained (i.e. thermodynamics or the reductive pathway). Although important challenges still need to be addressed for becoming a market product, such as more studies at pilot-scale level or a higher specificity to ethene as final product, micro-bial electro-remediation is a promising approach for treating chlorinated hydrocarbons in groundwater.

Inorganic contaminants

Metallic compounds

Metals can be present in groundwater mainly because of the aquifer's geochemistry, but also due to leakages

Aromatic ! hydrocarbons

Chlorinated hydrocarbons!


Inorganic Non-metals


Fig. 2. Summary of electrochemical reactions for the different pollutants treated in groundwater.

from industrial contamination. METs have been used as a technological approach to deal with different metals such as hexavalent uranium (Gregory and Lovley, 2005), hexavalent chromium (Huang et al., 2010), arsenite (Pous et al., 2015a) or selenite (Catal et al., 2009) (Table 2 and Fig. 2). In these cases, the objective is to change the metal oxidation state to one that presents lower toxicity and/or lower solubility. Different strategies can be explored depending on the metal that needs to be treated. In geologic-associated contamination [such as U(VI), As(III), V(V) or Se(IV)], in situ microbial electro-remediation might be the best strategy with the aim to immobilize the chemical species in their natural habitat. While in anthropogenic contamination [such as Cr(VI), Cd(II) or Cu(II)], the ex situ operation can be more appropriate to decontaminate the aquifer or, in the case of copper, to further recover it (Ter Heijne et al., 2010).

One of the most studied applications is the microbial electro-remediation of uranium-contaminated sites (Gregory and Lovley, 2005). In these sites, uranium is present in form of U(VI) and the most desirable strategy for

its bioremediation is the in situ conversion of U(VI) to U (IV), which is relatively insoluble and allows uranium immobilization in the aquifer (Gavrilescu et al., 2009). One of the most common strategies to promote uranium immobilization is to spike acetate or ethanol into the aquifer to stimulate native microbial U(VI) reduction (Gavrilescu et al., 2009). The interesting finding for MET applications was that Geobacter genus had been abundantly detected and enriched in sites where uranium bioremediation was implemented (Anderson et al., 2003; Shelobolina et al., 2008; Holmes et al., 2015). Bioreme-diation of U(VI) using MET instead of dosing acetate could decrease the ecological impact of the treatment as well as their cost, as it would only require the implementation of electrodes to stimulate bacterial activity. For this reason, the bioelectrochemical reduction of U(VI) using Geobacter has been proved in controlled laboratory experiments, as well as in real contaminated aquifers (in situ experiences) (Gregory and Lovley, 2005). The results obtained were promising, as 87% of uranium was recovered on the electrode surface (Gregory and Lovley, 2005). Moreover, bioelectrochemical U(VI) reduction represented a breaking point for the MET field. Until that moment, MET research had been focused on developing systems that relied on microbes able to deliver electrons to an electrode (microbial bioanodes). But the finding that the well-known Geobacter was also able to get electrons from an electrode to perform bioelectrochemical reduction of U(VI), fumarate or nitrate opened a new field of research: microbial biocathodes (Gregory et al., 2004; Gregory and Lovley, 2005). Although the understanding of microbial electron transfer fundamentals in bioanodes is abundant, the knowledge for biocathodes is still scarce (Rosenbaum et al., 2011). For this reason, investigations over how Geobacter is able to get electrons from an electrode can be seen as a lighthouse for biocathodes in general. For example, the finding that Geobacter sulfurreducens requires outer-surface c-type cytochromes, but not conductive pili (microbial nano-wires), for the reduction of U(VI) is a relevant contribution to the understanding of microbial reduction of soluble extracellular electron acceptors (Orellana et al., 2013). Moreover, the Geobacter versatility can also be used to hypothesize future pollutants to be evaluated using MET-based bioremediation. For example, as Geobacter is also able to reduce the soluble V(V) to the more insoluble V(IV), MET could also become an alternative process for bioremediating vanadium-contaminated sites (Ortiz-Bernad et al., 2004). However, until now, only one experience of biocathodic V(V) reduction has been reported so far, getting a removal efficiency of 76.8% (Zhang et al., 2015).

MET is also contributing on the bioremediation of one of the most harmful and abundant metallic contaminants,

arsenic, which is found in groundwater as arsenite [As (III)]. Its chemistry is different from the two metals discussed above, uranium and vanadium, where the highest oxidation state (U(VI) and V(V), respectively), were mobile, and thus, a reduction was needed for immobilization. In the case of arsenic, As(III) is highly mobile, while As(V) (arsenate) is more insoluble. Thus, the purpose is to use a bioanode able to oxidize arsenite to arsenate using a solid electrode as electron acceptor. The first study on arsenite oxidation using MET did not rely on arsenite-oxidizing microorganisms. It was focused on coupling a MFC with zero valent iron to produce H2O2, which was further used to oxidize As(III) to As(V) (Xue et al., 2013). In 2014, Webster et al. (2014) engineered Shewanella oneidensis to develop an arsen-ite-specific biosensor (Webster et al., 2014). One year later, the biologic arsenite oxidation using an electrode as electron acceptor was evaluated and proved (Pous et al., 2015a). A biofilm predominantly covered by y- and S-proteobacteria was able to perform the As(III) conversion at a poised anode potential of +497 mV versus SHE. From there on, the arsenite bioanode oxidation has been further investigated. The As(III) oxidation performance has been improved, and a maximum As(III) oxidation rate of 29.6 mgAs r1 day-1 has been achieved (Nguyen et al., 2016d). Moreover, it has been obtained additional knowledge about the microbial ecology responsible of microbial As(III) electro-remediation, and arsenite oxidation has been successfully coupled to cathodic nitrate reduction (Nguyen et al., 2016d, 2017).

Microorganisms able to catalyse arsenic oxidation are usually considered together with selenium players (Stolz et al., 2006). However, a different approach for dealing with Se, which is commonly found as selenite [Se(IV)], has been tested in METs. In this case, selenite was successfully reduced to elemental selenium in microbial biocathodes, which allowed its immobilization (Catal et al., 2009; Nguyen et al., 2016c). Moreover, the finding that the well-known electroactive Shewanella oneidensis MR-1 has the ability to convert Se(IV) into Se(0) opens the door for more investigations on selenium-contaminated groundwater treatment (Li et al., 2014).

Shewanella sp. has also been associated to chromium electro-remediation (Hsu et al., 2012; Xafenias et al., 2013, 2015). Chromium is commonly used in different industries, and it can finally be released in their effluent streams as Cr(VI). As a result, it can be found in some groundwater bodies. In microbial biocathodes, Cr(VI) can be converted into Cr(III) using either a MFC (Huang et al., 2010; Hsu et al., 2012; Xafenias et al., 2013, 2015; Wu et al., 2015; Song et al., 2016) or a MEC configuration (Xafenias et al., 2013; Huang et al., 2015). The basis of the process is to convert the soluble Cr(VI) into a less soluble form, Cr(III). However, chromium can

precipitate on the Shewanella surface (Kim et al., 2014), which could be seen as a limiting factor at long-term operation. Nevertheless, the ability of MET to convert and anchor Cr(VI) can allow effluent concentrations below 5 ppb, which is below the guideline values for drinking water (Hsu et al., 2012).

In conclusion, microbial electro-remediation is a versatile technology that allows the treatment of different metal contaminants, and it can be applied in situ or ex situ depending on the contaminant.

Non-metallic inorganic contaminants - nutrients

The presence of inorganic non-metallic contaminants can be found in different groundwater bodies. MET has been proposed as an alternative method for nitrates (Pous et al., 2013; Zhang and Angelidaki, 2013), ammonium (Wei et al., 2015a), sulfates (Coma et al., 2013; Pozo et al., 2016) and perchlorates (Butler et al., 2010) (Table 2 and Fig. 2). Nitrate (Mencio et al., 2011; Spra-gue et al., 2011), ammonium (Mastrocicco et al., 2013; Scheiber et al., 2016) and perchlorate (Bohlke et al., 2009; Izbicki et al., 2015) are mainly found in groundwater due to anthropogenic activities. In contrast, sulfates can also be accumulated because of aquifer's geology (Burg et al., 2017) and seawater intrusion (Bottrell et al., 2008), but it poses a lower risk for human health (Liamleam and Annachhatre, 2007).

Nitrates are one of the most widespread contaminants threatening groundwater's usage as drinking water. It can be found in several regions around the world as the bad face of intensive agriculture and livestock production (Mencio et al., 2011; Sprague et al., 2011). Separation-based technologies, such as reverse osmosis, reverse electrodialysis and ion exchange have been used to deal with nitrates in groundwater. These technologies are effective on removing nitrate, but they are energy-intensive and they produce waste brine concentrated with nitrates of difficult disposal (Twomey et al., 2010). For this reason, technologies based on converting nitrates (to dinitrogen gas preferably) are being investigated. They can be divided into two main groups: abiotic and biologic. The abiotic alternatives are mainly based on electrocatalysis or the usage of a chemical catalyser, such as zero valent iron (ZVI) (Duca and Koper, 2012; Fu et al., 2014). Besides they could become effective strategies for removing nitrate, their main challenge is the low reduction specificity to dinitrogen gas (N2) as end-product. Nitrate is converted into ammonium in most of the cases, which requires a post-treatment (Duca and Koper, 2012; Fu et al., 2014). On the contrary, biologic treatments rely on bacteria, which are considered to be low-cost and self-renewable catalysers. Bacteria are able to convert nitrate into dinitrogen gas through the

Table 2. Summary of inorganic pollutants treated in groundwater using microbial electro-remediation.

Operational WE potential Dominant associated

Pollutant Reaction Placement mode (mV vs. SHE) microbiome


U(VI) U(VI) ? U(IV) Ex-situ 3-MEC -303 Geobacter sufurreducens

In-situ 3-MEC -303 Desulfotomaculum,


As(III) As(lll) ? As(V) Ex-situ 3-MEC +497 8, c-proteobacteria

+500 Achromobacter sp., Ensifer

sp., Sinorhizobium sp.


Se(IV) Se(IV) ? Se(0) Ex-situ MFC - -

3-MEC -300 Cronobacter

Cr(VI) Cr(VI) ? Cr(III) Ex-situ MFC - -

Shewanella sp.

Shewanella oneidensis


Shewanella oneidensis

2-MEC -303 Proteobacteria

3-MEC Shewanella oneidensis

Cu(II) Cu(II) ? Cu(0) Ex-situ 2-MEC - Proteobacteria

MFC - Stenotrophomonas maltiphilia,

Citrobacter sp., Pseudomonas

aeruginosa, Stenotrophomonas

Cd(II) Cd(II) ? Cd(0) Ex-situ 2-MEC - Proteobacteria


NO- NO3- ? N2 Ex-situ 2-MEC - -

a, ß, c-proteobacteria,



Nitratireductor sp., Shinella sp.,

Aeromonas sp., Pseudomonas

sp., Curtobacterium sp.,

Dyella sp.

3-MEC -303 Geobacter sp.

Geobacter metallireducens

-123 -

-700 Shinella sp., Alicycliphilus sp.

MFC - -

In-situ 2-MEC - -

3-MEC -700 Thiobacillus sp., Paracoccus sp.

ClO4- ClO4- ? Cl- Ex-situ 3-MEC -303 Dechloromonas, Azospira

MFC _ ß-proteobacteria, Bacteroidetes

Bacteroidetes, Firmicutes,


2-MEC - Aureibacter sp., Fulvivirga sp.,



Ex-situ Ex-situ


-260 -900

-1100 -800

Thermotalea sp., Thauera sp.

Methanobacterium, Desulfovibrio Methanobacteriales Desulfovibrio sp., Sulfuricurvum sp. 8-proteobacteria Alcaligenes sp., Paracoccus sp.

(Gregory and Lovley, 2005)

(Pous et al, 2015a) (Nguyen et al, 2016d)

(Nguyen et al, 2017) (Catal et al., 2009) (Nguyen et al, 2016c) (Huang et al., 2010) (Hsu et al., 2012) (Xafenias et al., 2013) (Wu et al., 2015) (Xafenias et al., 2015) (Song et al., 2016) (Huang et al., 2015) (Xafenias et al., 2013) (Huang et al., 2015) (Shen et al., 2017)

(Huang et al., 2015)

(Sakakibara and Kuroda, 1993) (Feleke et al., 1998) (Park et al., 2005) (Park et al., 2006)

(Tong et al., 2013) (Kondaveeti and Min, 2013) (Kondaveeti et al, 2014) (Huang et al., 2013) (Nguyen et al, 2015)

(Gregory et al., 2004)

(Pous et al., 2015a,b,c) (Nguyen et al, 2016a) (Pous et al., 2013) (Tong and He, 2013) (Nguyen et al, 2016b) (Thrash et al., 2007) (Shea et al., 2008) (Butler et al., 2010) (Mieseler et al, 2013)

(Wang et al., 2014)

(Coma et al., 2013) (Pozo et al., 2015)

(Pozo et al., 2016) (Blazquez et al., 2016)

(Rakoczy et al., 2013) (Rabaey et al., 2006)

WE accounts for Working Electrode; MFC indicates Microbial Fuel Cell; 2-MEC indicates a Microbial Electrolysis Cell with a 2-electrodes configuration and 3-MEC accounts for a Microbial Electrolysis Cell with a 3-electrodes configuration.

SO24- ? S2

denitrification process. Biologic nitrate removal in METs has been widely studied because of its possible application to wastewater treatment (Clauwaert et al., 2007; Vir-dis etal., 2010; Puig et al., 2011; Pous et al., 2015b; Vilajeliu-Pons etal., 2015). Although bioelectrochemical dissimilatory nitrate reduction (i.e. nitrate conversion to ammonium) has been described (Sander etal., 2015), nitrate removal in METs naturally follows the conventional denitrifying pathway in most of the cases (Clauwaert et al., 2007; Virdis et al., 2008). Nitrates are reduced to dinitrogen gas in the cathode compartment. However, literature regarding the treatment of nitrate-polluted groundwater using MET is not abundant. The difference between treating nitrate in wastewater or groundwater using MET is relevant, as it has been demonstrated that the low conductivity of groundwater (< 1 mS cm-1) limits the MET performance (Puig et al., 2012). Thus, groundwater treatment is expected to have higher restrictions compared to wastewaters with higher conductivities and buffer capacities.

In the first studies regarding microbial electro-remediation of nitrate, the mechanism was based on electrochemical water splitting to provide hydrogen to hydrogenotrophic denitrifiers (Sakakibara and Kuroda, 1993; Prosnansky et al., 2002). This process was considered an alternative to conventional hydrogenotrophic denitrification (Karanasios etal., 2010), in which hydrogen gas is directly supplied to a biological reactor. But this process is mass transfer limited due to the low solubility of hydrogen [1.6 mg l-1 at 20 °C (Soares, 2000)]. Sakakibara and Kuroda (1993) demonstrated that the complete reduction of nitrate to dinitrogen gas could be accomplished by applying different currents from 0 to 40 mA, which lead to increase the denitrification rate up to 0.15 mmol h-1. Although the authors stated that denitrification was mediated by H2 (produced in situ by electrochemical water splitting), it cannot be excluded that denitrification using the electrode as electron donor was taking place simultaneously. Besides the fact that the in situ electrochemical production of hydrogen for nitrate reduction was effective [nitrate removal rates up to 394 mgN l-1 day-1 (Prosnansky et al., 2002)], it implied a certain lack of process control. The hydrogen generated in the cathode may or may not be used for nitrate reduction. Hence, lower columbic efficiency can be expected for this type of configuration.

In 2004, Gregory and co-workers observed that auto-trophic denitrifiers were able to use a poised cathode electrode (-500 mV versus Ag/AgCl, -303 mV versus SHE) as electron donor, getting an electrode predominantly covered by Geobacter sp. (Gregory et al., 2004). Electron uptake from an electrode to perform denitrifica-tion was also demonstrated in groundwater (Park et al., 2005). In this case, by applying 200 mA, a nitrate

removal rate of 435 mgN Г1 h-1 (10440 mgN Г1 day-1) was achieved in batch mode (Park et al., 2005). In groundwater, the electrode was predominantly covered by a-, p-, y-proteobacteria and Flavobacteriia, which indicated that not only Geobacter sp. (Gregory et al., 2004) were capable to perform bioelectrochemical denitrifica-tion. From there on, the investigation of nitrate removal in groundwater has been focused on determining the best operational strategies to increase nitrate removal rates.

If a MFC strategy is chosen to treat nitrate-polluted groundwater, organic matter needs to be dosed into the anode compartment. Despite organic matter is not directly added to groundwater (it is added in a different compartment), it implies an extra cost. Hence, to convince future stakeholders that a BES operated as a MFC is suitable for groundwater bioremediation (Pous et al., 2013; Zhang and Angelidaki, 2013), the denitrification rates should be objectively higher than those obtained in conventional heterotrophic denitrification systems. By now, the highest denitrification rate reported in a denitrifying MFC has been around 500 mgN l-1 day-1 treating either groundwater (Zhang and Angelidaki, 2013) or synthetic wastewater (Clauwaert et al., 2009). A conventional heterotrophic treatment of nitrate-polluted groundwater as membrane bioreactors (MBR) can reach values up to 1700 mgN l-1 day-1 (Wasik et al., 2001).

MET can be a market alternative for treating nitrate-contaminated groundwater if it moves towards the idea of developing a fully autotrophic treatment. In this sense, a MEC operation is preferred, where external energy can be used to directly empower the denitrifying activity (Sakakibara and Kuroda, 1993). The fully autotrophic nitrate removal in groundwater has been evaluated in both MEC 2-electrode (Sakakibara and Kuroda, 1993; Feleke etal., 1998; Park etal., 2005, 2006; Huang etal., 2013; Kondaveeti and Min, 2013; Kondaveeti et al., 2014; Nguyen et al., 2015) or 3-electrode arrangement (Pous etal., 2015c; Nguyen etal., 2016a,b). Except for the case of Park et al. (2005), who reported 435 mgN l-1 h-1 in a 2-MEC, and Pous et al. (2017), who reported 849 mgN l-1 day-1 in a 3-MEC, the other authors obtained nitrate removal rates below 200 mgN l-1 day-1. A lower capital cost is required for a MEC 2-electrodes, as it only needs a conventional power supply (e.g. power supply 0-30 V, 0-3 A has a cost of around 150 €). But MEC 2-electrodes have a risk of side reactions (i.e. hydrogen evolution). On the contrary, the capital cost is higher for a MEC 3-electrodes because a potentiostat is needed (e.g. potentiostat 0-20 V, 0-1 A has a cost of around 5000 €). However, in MEC 3-electrodes, the cathode potential is controlled, which gives a better control over the electrode reactions. Thus, with both presenting advantages and

disadvantages, the decision of choosing one or another will depend on each real application case.

In order to deliver drinking water, the plethora of configurations to deal with nitrate in groundwater is usually thought as ex situ applications (intensive treatment). However, experiences on in situ microbial electro-remediation have also been explored, giving promising results (Tong and He, 2013; Zhang and Angelidaki, 2013; Nguyen et al., 2016b).

Another less common, but sometimes present, nitrogen compound is ammonium. It is a contaminant that can be found in subsurface waters that have received industrial or petrochemical pollution (Voyevoda et al., 2012). In those spills where oxygen is at low concentrations, ammonium is not oxidized into nitrate at the surface neither during the percolation (Buss et al., 2004). The main strategy to treat ammonium using METs is based on oxidizing ammonium aerobically into nitrate, which is then reduced into dinitrogen gas in a denitrifying biocathode (Virdis et al., 2008, 2010; Vilajeliu-Pons et al., 2015, 2017). This strategy has been used to treat ammonium from real contaminated groundwater with satisfactory results in terms of ammonium oxidation, but low efficiencies of nitrate removal (Wei et al., 2015a,b). Wei et al., 2015a observed a 100% ammonium oxidation (20 mgN T1) but an insufficient nitrate removal in a 0.16-l reactor. While Wei et al., 2015b reached an stable ammonium removal of 100% during an operation time of 200 days in a MET presenting a 26 l volume and operated at 15 days HRT, but again an insufficient nitrate removal was observed. Another strategy that is being developed for treating ammonium is the ammonium oxidation using the anode as the final electron acceptor (Zhan et al., 2012, 2014; Zhu et al., 2016), but still low ammonium oxidation rates have been obtained [around 60 mgN T1 day-1 (Zhan et al., 2014)].

Perchlorate is an emerging pollutant in groundwater, which consumption can cause a depression of thyroid hormone formation (Greer et al., 2002). The biologic treatment of perchlorate is performed by perchlorate-reducing bacteria, which are able to convert ClO4- into CP. Besides no literature is available on perchlorate treatment in real groundwater, electro-remediation of perchlorate in organic matter-free media has been already proved (Butler et al., 2010). Like other biocathode-based processes, the investigation of ClO4- reduction has been evaluated in MFC and MEC modes. Butler et al. (2010) were able to obtain electrical current by perchlorate cathodic reduction at a maximum rate of 24 mg l-1 day-1 (Butler et al., 2010). Under MEC mode, the perchlorate reduction was also possible at poised cathode potential of -500 mV versus Ag/AgCl (-303 mV versus SHE) (3-electrodes) (Thrash et al., 2007) or by supplying a fixed current (2-MEC) (Wang et al., 2014). However, the way to enrich this

kind of reactors is one of the critical steps for MET application. For this reason, different inoculation strategies have been tested, such as the enrichment perchlorate-reducing bacteria fed with acetate (Mieseler et al., 2013) or the adaptation of a denitrifying MET to perform perchlorate reduction (Shea et al., 2008). Both of them showed promising results, which should encourage further research on perchlorate bioremediation using METs.

Sulfates occurrence in groundwater also presents interest for microbial electro-remediation, despite its low risk for human health. Some subsurface waters can present sulfate concentrations above the guideline value, and it also represents a risk for the utility infrastructure because of its possible conversion into hydrogen sulfide, even at low concentrations. Because of its low reduction potential [E0 (SO4°/HS°) = 0.252 V versus SHE, E0 (SO4O/S0) = 0.357 V versus SHE (Rabaey et al, 2009)] compared to organic matter oxidation [E0 (CH3COOO/ HCO3) = 0.187 V versus SHE (Logan et al., 2006)], the reduction of SO43 in the cathode of a MFC is not feasible (Coma et al., 2013). Hence, it is necessary to apply external power to reach relevant removal rates. For example, Coma et al. (2013) observed a sulfate removal rate of 2 gSO4° mo3 day31 when operating as MFC (0 V applied), but a removal of around 65 gSO4° mo3 day31 when operating as MEC and applying 0.7 V. Not only the achievement of sulfate removal rates is important, but it is also important to determine which reduction product has been produced. In order to remove the sulfates from water using MET, two strategies have been evaluated: (i) sulfate conversion to sulfide, which could be extracted by promoting its precipitation as metal sulfide (Su et al., 2012; Coma et al., 2013; Pozo et al., 2016); (ii) sulfate conversion into elemental sulfur, which would allow S recovery for further usage if a cheap strategy for extraction is developed (Blazquez et al., 2016; Chatterjee et al., 2017). Nevertheless, the highest importance of studying sulfates bioelectrocatalysis for groundwater application is its coexistences together with other contaminants that posses higher risks for human health [e.g. together with chlorinated hydrocarbons (Lai et al., 2015) or with nitrates (Nguyen et al., 2016a)]. Therefore, the importance of the understanding of microbial electro-remediation of inorganic non-metallic pollutants in groundwater relies not only on the capacity of MET to treat these contaminants, but also on the possible interferences that these common contaminants can provoke to the electro-remediation of others.

Hurdles and challenges for groundwater microbial electro-remediation

The scarcity of nutrients is one of the main hurdles that microbial electro-remediation of groundwater has to face.

From a chemical-specific sight, N'Guessan etal. (2010) investigated the effect of phosphate limitation in Geobacter sp. The authors demonstrated that G. sulfurreducens is able to reduce U(VI) at phosphate-limiting conditions (0.217 mM phosphate) (N'Guessan etal., 2010). Thus, the electroactive microorganism G. sulfurreducens was not limited by low nutrient availability, which gives good perspectives for their survival when treating groundwa-ter.

From a general perspective, a clear indication of the low availability of chemical species itself is the low conductivity of groundwater (< 1.6 mS cm-1). The low conductivity can have a negative impact on MET, it implies higher ohmic and transport losses (Logan et al., 2006). For example, in the case of MET-based nitrate removal, the decrease in conductivity from 4.3 to 1.3 mS cm-1 implied a decrease of 44% on nitrate removed (from 13.5 to 7.5 mgN l-1) (Puig etal., 2012). Moreover, the low conductivity can also lead to pH gradients by promoting to acidic pHs in the anode and basification in cathode. pH shifts can directly harm the electroactive bacteria and their removal performance (Clauwaert et al., 2008; Fornero etal., 2010), and it can lead to additional problems for the specific application of groundwater treatment. Depending on the aquifer's geochemistry, groundwater can present a high concentration of calcium, magnesium and bicarbonate (i.e. hardness) (Briggs and Ficke, 1977). The reductive nature of cathodes, together with the low buffering capacity of ground-water, can promote basified zones on the electrode surface. This induces scaling with the consequent blockage of the cathode electrodes, which can end up in MET deactivation (Santini etal., 2016). Besides it could be seen as a new application for MET (water softening) (Gabrielli etal., 2006; Zeppenfeld, 2011), strategies for solving this issue must be explored.

Another challenge for MET treatment of groundwater is the presence of mixtures of different contaminants (Squillace et al., 2002). The study of electro-remediation of co-contaminants in MET is limited, and few examples, such as perchlorate/nitrate (Xie et al., 2014) or cis-DCE/ nitrate/sulfate (Lai et al., 2015), can be found.

The cocktail perchlorate/nitrate is of a high interest, as they both can occur simultaneously (Dasgupta et al., 2005). On the one hand, anthropogenic perchlorate contamination has been linked to ammonium perchlorate (a missile propellant) (Hogue, 2003) and to nitrate-based fertilizers, which also contain perchlorate (Susarla et al., 1999; Urbansky etal., 2000). It is relevant the case of the Chilean nitrate, since its perchlorate content is about 0.05-0.2 wt % ClO- (Urbansky etal., 2001). On the other hand, perchlorate can be naturally produced by sea salt aerosol photolysis in the atmosphere. This process can also involve nitrogen oxides, which can end up

with nitrate deposition (Dasgupta et al., 2005). Xie et al. (2014) evaluated the occurrence of both nitrate and per-chlorate in a MET. The experiments were performed in a perchlorate-reducing biocathode grown at a poised cathode potential of -252 mV versus SHE (-500 mV versus SCE). After testing the perchlorate removal (initial concentration of 0.70 mM ClO-) together with different nitrate concentrations (0-2.10 m MNO-), the authors observed lower perchlorate reductions when higher nitrate concentrations were present. In batch experiments, a perchlorate concentration of 0.70 mM was totally consumed in 4 days when spiked alone. Twelve days were needed for its removal when 0.07 mM of nitrate was added, and perchlorate reduction was totally suppressed when nitrate was added at 2.10 mM (Xie etal., 2014). This inhibition of perchlorate reduction in the presence of nitrate is not specific of bioelectrochemi-cal perchlorate reduction, and it has also been observed when using organic carbon or hydrogen as electron donors (Zhao etal., 2011; Ricardo etal., 2012). The reduction potentials of nitrate and perchlorate are similar (E0 NO3/N2 = 1.25 V; E0 ClO 4/Cl- = 1.28 V), which make them electron competitors (Bardiya and Bae, 2011). In fact, most of the perchlorate-reducing bacteria identified so far are also able to denitrify (Nozawa-Inoue et al., 2011). However, nitrate consumption allows higher cell growth. In consequence, the perchlorate reduction starts only after nitrate is depressed in most of the cases described (Bardiya and Bae, 2011). Hence, the decrease in perchlorate reduction in the presence of nitrate is linked to a substrate preference over nitrate. Thus, the tendency of bacteria over denitrification should be taken into account when dealing with a perchlorate/nitrate cocktail, and strategies for allowing perchlorate reduction should be implemented.

On the removal of cis-DCE, the presence of nitrate and sulfate can also be possible, as they are one of the most widespread contaminants. For this reason, Lai etal. (2015) investigated whether nitrate and sulfate presence could affect bioelectrochemical reductive dechlorination of cis-DCE (Lai etal., 2015). They observed that the cathode potential had a key role on selecting the target pollutant. In the cathode potential range evaluated (-550/-750 mV versus SHE), nitrate reduction always took place. As cathode potential was lowered, sulfate reduction and methanogenesis increased their activity. Besides reductive dechlorination was not inhibited, the electricity consumption incremented due to crossed reactions at lower cathode potentials. In this case, reductive dechlorination contribution was < 1% of the electrons consumed. The effect of sulfate was also evaluated on bioelectrochemical nitrate reduction (Nguyen etal., 2016a). Nguyen and co-workers compared the denitrifying activity with or without

sulfate (50 mgS-SO23 lo1), and they observed that the presence of sulfate suppressed, somehow, the overall denitrifying activity. Not only the nitrate removal rate decreased but also nitrite was accumulated as undesired denitrification intermediate. Therefore, it would be welcomed a further understanding on chemical species that coexist with the target pollutant in groundwater.

Outlook for the future of microbial electro-remediation of groundwater

Microbial electro-remediation represents a unique opportunity to develop a robust, resilient and sustainable technology in a circular economy context to deal with different contaminants that are already present in our groundwater bodies. A considerable development has been done in the last 20 years in this field. Contaminants of different chemical nature (e.g. polycyclic heteroaromatic hydrocarbons, nutrients or metals) have been successfully treated using microbial electrochemical technologies. The technology proved its flexibility, as it has been adapted for ex situ or in situ treatment applications depending on the target pollutant. Moreover, MET-based knowledge can also be applied to develop biosensors for contaminant or microbial monitoring in groundwater. However, in order to keep paving the way to its future implementation, specific development might be required for each specific pollutant, as their characteristics require different operational strategies. Strategies to overcome the restricting characteristics of groundwa-ter and to face problems like carbonate scaling or those related to cocktails of contaminants need to be investigated and implemented. Moreover, testing at pilot plant level is still scarce, which demands an increase in scaling-up orientated research to avoid technological stagnation.


This research was financially supported by the Spanish Government (CTQ2014-53718-R and CTM2015-71982-REDT) and the University of Girona (MPCUdG2016/ 137). LEQUIA has been recognized as a consolidated research group by the Catalan Government with code 2014-SGR-1168.

Conflict of interest

None declared.


Acar, Y.B., Gale, R.J., Alshawabkeh, A.N., Marks, R.E., Pup-pala, S., Bricka, M., and Parker, R. (1995) Electrokinetic

remediation: basics and technology status. J Hazard Mater 40: 117-137.

Altalyan, H.N., Jones, B., Bradd, J., Nghiem, L.D., and Alya-zichi, Y.M. (2016) Removal of volatile organic compounds (VOCs) from groundwater by reverse osmosis and nanofiltration. J. Water Process Eng 9: 9-21.

Anderson, R.T., Vrionis, H.A., Ortiz-Bernad, I., Resch, C.T., Long, P.E., Dayvault, R., et al. (2003) Stimulating the in situ activity of Geobacter species to remove uranium from the groundwater of a uranium-contaminated aquifer. Appl Environ Microbiol 69: 5884-5891.

Aulenta, F., Catervi, A., Majone, M., Panero, S., Reale, P., and Rossetti, S. (2007) Electron transfer from a solidstate electrode assisted by methyl viologen sustains efficient microbial reductive dechlorination of TCE. Environ Sci Technol 41: 2554-2559.

Aulenta, F., Reale, P., Catervi, A., Panero, S., and Majone, M. (2008) Kinetics of trichloroethene dechlorination and methane formation by a mixed anaerobic culture in a bio-electrochemical system. Electrochim Acta 53: 5300-5305.

Aulenta, F., Reale, P., Canosa, A., Rossetti, S., Panero, S., and Majone, M. (2010) Characterization of an electro-active biocathode capable of dechlorinating trichloroethene and cis-dichloroethene to ethene. Biosens Bioelectron 25: 1796-1802.

Aulenta, F., Tocca, L., Verdini, R., Reale, P., and Majone, M. (2011) Dechlorination of trichloroethene in a continuous-flow bioelectrochemical reactor: effect of cathode potential on rate, selectivity, and electron transfer mechanisms. Environ Sci Technol 45: 8444-8451.

Aulenta, F., Verdini, R., Zeppilli, M., Zanaroli, G., Fava, F., Rossetti, S., and Majone, M. (2013) Electrochemical stimulation of microbial cis-dichloroethene (cis-DCE) oxidation by an ethene-assimilating culture. N Biotechnol 30: 749755.

Bamforth, S.M., and Singleton, I. (2005) Bioremediation of polycyclic aromatic hydrocarbons: current knowledge and future directions. J Chem Technol Biotechnol 80: 723736.

Bardiya, N., and Bae, J.-H. (2011) Dissimilatory perchlorate reduction: a review. Microbiol Res 166: 237-254.

Blazquez, E., Gabriel, D., Baeza, J.A., and Guisasola, A. (2016) Treatment of high-strength sulfate wastewater using an autotrophic biocathode in view of elemental sulfur recovery. Water Res 105: 395^05.

Bohlke, J.K., Hatzinger, P.B., Sturchio, N.C., Gu, B., Abbene, I., and Mroczkowski, S.J. (2009) Atacama perchlorate as an agricultural contaminant in groundwater: isotopic and chronologic evidence from Long Island, New York. Environ Sci Technol 43: 5619-5625.

Bottrell, S., Tellam, J., Bartlett, R., and Hughes, A. (2008) Isotopic composition of sulfate as a tracer of natural and anthropogenic influences on groundwater geochemistry in an urban sandstone aquifer, Birmingham, UK. Appl Geo-chemist 23: 2382 2394.

Briggs, J.C. and Ficke, J.F. (1977) Quality of rivers of the United States, 1975 water year; based on the National Stream Quality Accounting Network (NASQAN) Reston, VA.

Burg, A., Gavrieli, I., and Guttman, J. (2017) Concurrent salinization and development of anoxic conditions in a

confined aquifer, Southern Israel. Groundwater 55: 183198.

Buss, S.R., Herbert, A.W., Morgan, P., Thornton, S.F., and Smith, J.W.N. (2004) A review of ammonium attenuation in soil and groundwater. Q J Eng Geol Hydrogeol 37(3): 47-359.

Butler, C.S., Clauwaert, P., Green, S.J., Verstraete, W., and Nerenberg, R. (2010) Bioelectrochemical perchlorate reduction in a microbial fuel cell. Environ Sci Technol 44: 4685-4691.

Catal, T., Bermek, H., and Liu, H. (2009) Removal of selen-ite from wastewater using microbial fuel cells. Biotechnol Lett 31: 1211-1216.

Chambon, J.C., Bjerg, P.L., Scheutz, C., Bffilum, J., Jakob-sen, R., and Binning, P.J. (2013) Review of reactive kinetic models describing reductive dechlorination of chlorinated ethenes in soil and groundwater. Biotechnol Bioeng 110: 1-23.

Chang, S.-H., Wu, C.-H., Wang, R.-C., and Lin, C.-W.

(2016) Electricity production and benzene removal from groundwater using low-cost mini tubular microbial fuel cells in a monitoring well. J Environ Manage 193: 551557.

Chatterjee, P., Ghangrekar, M.M., Rao, S., and Kumar, S.

(2017) Biotic conversion of sulphate to sulphide and abiotic conversion of sulphide to sulphur in a microbial fuel cell using cobalt oxide octahedrons as cathode catalyst. Bioprocess Biosyst Eng 40: 759-768.

Clauwaert, P., Rabaey, K., Aelterman, P., De Schamphe-laire, L., Pham, T.H., Boeckx, P., etal. (2007) Biological denitrification in microbial fuel cells. Environ Sci Technol 41: 3354-3360.

Clauwaert, P., Aelterman, P., Pham, T.H., De Schamphe-laire, L., Carballa, M., Rabaey, K., and Verstraete, W. (2008) Minimizing losses in bio-electrochemical systems: the road to applications. Appl Microbiol Biotechnol 79: 901-913.

Clauwaert, P., Desloover, J., Shea, C., Nerenberg, R., Boon, N., and Verstraete, W. (2009) Enhanced nitrogen removal in bio-electrochemical systems by pH control. Biotechnol Lett 31: 1537-1543.

Coma, M., Puig, S., Pous, N., Balaguer, M.D., and Colprim, J. (2013) Biocatalysed sulphate removal in a BES cathode. Bioresour Technol 130: 218-223.

Daghio, M., Aulenta, F., Vaiopoulou, E., Franzetti, A., Arends, J.B.A., Sherry, A., et al. (2017) Electrobioremedi-ation of oil spills. Water Res 114: 351-370.

Dasgupta, P.K., Martinelango, P.K., Jackson, W.A., Anderson, T.A., Tian, K., Tock, R.W., and Rajagopalan, S. (2005) The origin of naturally occurring perchlorate: the role of atmospheric processes. Environ Sci Technol 39: 1569-1575.

Dominguez-Garay, A., Boltes, K., and Esteve-Nunez, A. (2016) Cleaning-up atrazine-polluted soil by using microbial electroremediating cells. Chemosphere 161: 365371.

Dominguez-Garay, A., Rodrigo Quejigo, J., Dörfler, U., Schroll, R. and Esteve-Nunez, A. (2017) Bioelectrovent-ing: an electrochemical-assisted bioremediation strategy for cleaning-up atrazine-polluted soils. Microb Biotechnol (In Press) DOI: 10.1111/1751-7915.12687.

Duca, M., and Koper, M.T.M. (2012) Powering denitrification: the perspectives of electrocatalytic nitrate reduction. Energy Environ Sci 5: 9726-9742.

Environmental Protection Agency Office of water, office of Science, and Technology: Washington, D., U.S. (2004) National Recommended Water Quality Criteria; 4304T.

Feleke, Z., Araki, K., Sakakibara, Y., Watanabe, T., and Kuroda, M. (1998) Selective reduction of nitrate to nitrogen gas in a biofilm-electrode reactor. Water Res 32: 2728-2734.

Feng, C., Yue, X., Li, F., and Wei, C. (2013) Bio-current as an indicator for biogenic Fe(II) generation driven by dis-similatory iron reducing bacteria. Biosens Bioelectron 39: 51-56.

Fornero, J.J., Rosenbaum, M., Cotta, M.A., and Angenent, L.T. (2010) Carbon dioxide addition to microbial fuel cell cathodes maintains sustainable catholyte pH and improves anolyte pH, alkalinity, and conductivity. Environ Sci Technol 44: 2728-2734.

Friman, H., Schechter, A., Nitzan, Y., and Cahan, R. (2013) Phenol degradation in bio-electrochemical cells. Int Biode-terior Biodegradation 84: 155-160.

Fu, F., Dionysiou, D.D., and Liu, H. (2014) The use of zero-valent iron for groundwater remediation and wastewater treatment: a review. J Hazard Mater 267: 194-205.

Gabrielli, C., Maurin, G., Francy-Chausson, H., Thery, P., Tran, T.T.M., and Tlili, M. (2006) Electrochemical water softening: principle and application. Desalination 201: 150-163.

Gavrilescu, M., Pavel, L.V., and Cretescu, I. (2009) Characterization and remediation of soils contaminated with uranium. J Hazard Mater 163: 475-510.

Greer, M.A., Goodman, G., Pleus, R.C., and Greer, S.E. (2002) Health effects perchlorate contamination: the dose response for inhibition of thyroidal radioiodine uptake in humans. Environ Health Perspect 110: 927-937.

Gregory, K.B., and Lovley, D.R. (2005) Remediation and recovery of uranium from contaminated subsurface environments with electrodes. Environ Sci Technol 39: 89438947.

Gregory, K.B., Bond, D.R., and Lovley, D.R. (2004) Graphite electrodes as electron donors for anaerobic respiration. Environ Microbiol 6: 596-604.

Hedbavna, P., Rolfe, S.A., Huang, W.E., and Thornton, S.F. (2016) Biodegradation of phenolic compounds and their metabolites in contaminated groundwater using microbial fuel cells. Bioresource Technol 200: 426-434.

Hogue, C. (2003) Rocket-fueled river. Chem Eng News 81: 37-46.

Holliger, C., and Schraa, G. (1994) Physiological meaning and potential for application of reductive dechlorination by anaerobic bacteria. FEMS Microbiol Rev 15: 297-305.

Holliger, C., Wohlfarth, G., and Diekert, G. (1998) Reductive dechlorination in the energy metabolism of anaerobic bacteria. FEMS Microbiol Rev 22: 383-398.

Holmes, D.E., Giloteaux, L., Chaurasia, A.K., Williams, K.H., Luef, B., Wilkins, M.J., etal. (2015) Evidence of Geobac-ter-associated phage in a uranium-contaminated aquifer. ISME J 9: 333-346.

Hsu, L., Masuda, S.A., Nealson, K.H., and Pirbazari, M. (2012) Evaluation of microbial fuel cell Shewanella

biocathodes for treatment of chromate contamination. RSCAdv 2: 5844-5855.

Huang, L., Chen, J., Quan, X., and Yang, F. (2010) Enhancement of hexavalent chromium reduction and electricity production from a biocathode microbial fuel cell. Bioprocess Biosyst Eng 33: 937-945.

Huang, B., Feng, H., Ding, Y., Zheng, X., Wang, M., Li, N., et al. (2013) Microbial metabolism and activity in terms of nitrate removal in bioelectrochemical systems. Elec-trochim Acta 113: 29-36.

Huang, L., Wang, Q., Jiang, L., Zhou, P., Quan, X., and Logan, B.E. (2015) Adaptively evolving bacterial communities for complete and selective reduction of Cr(VI), Cu (II), and Cd(II) in biocathode Bioelectrochemical Systems. Environ Sci Technol 49: 9914-9924.

Izbicki, J.A., Teague, N.F., Hatzinger, P.B., Böhlke, J.K., and Sturchio, N.C. (2015) Groundwater movement, recharge, and perchlorate occurrence in a faulted alluvial aquifer in California (USA). Hydrogeol J 23: 467-491.

Karanasios, K.A., Vasiliadou, I.A., Pavlou, S., and Vayenas, D.V. (2010) Hydrogenotrophic denitrification of potable water: a review. J Hazard Mater 180: 20-37.

Katsoyiannis, I.A., Hug, S.J., Ammann, A., Zikoudi, A., and Hatziliontos, C. (2007) Arsenic speciation and uranium concentrations in drinking water supply wells in Northern Greece: correlations with redox indicative parameters and implications for groundwater treatment. Sci Total Environ 383: 128-140.

Kim, D.-H., Park, S., Kim, M.-G., and Hur, H.-G. (2014) Accumulation of amorphous Cr(III)-Te(IV) nanoparticles on the surface of Shewanella oneidensis MR-1 through reduction of Cr(VI). Environ Sci Technol 48: 1459914606.

Kondaveeti, S., and Min, B. (2013) Nitrate reduction with biotic and abiotic cathodes at various cell voltages in bio-electrochemical denitrification system. Bioprocess Biosyst Eng 36: 231-238.

Kondaveeti, S., Lee, S.-H., Park, H.-D., and Min, B. (2014) Bacterial communities in a bioelectrochemical denitrifica-tion system: the effects of supplemental electron acceptors. Water Res 51: 25-36.

Lai, A., Verdini, R., Aulenta, F., and Majone, M. (2015) Influence of nitrate and sulfate reduction in the bioelectrochemi-cally assisted dechlorination of cis-DCE. Chemosphere 125: 147-154.

Lai, A., Aulenta, F., Mingazzini, M., Palumbo, M.T., Papini, M.P., Verdini, R., and Majone, M. (2017) Bioelectrochemical approach for reductive and oxidative dechlorination of chlorinated aliphatic hydrocarbons (CAHs). Chemosphere 169: 351-360.

Leitäo, P., Rossetti, S., Nouws, H.P.A., Danko, A.S., Majone, M., and Aulenta, F. (2015) Bioelectrochemically-assisted reductive dechlorination of 1,2-dichloroethane by a Dehalococcoides-enriched microbial culture. Biore-source Technol 195: 78-82.

Leitäo, P., Rossetti, S., Danko, A.S., Nouws, H., and Aulenta, F. (2016) Enrichment of Dehalococcoides mccartyi spp. from a municipal activated sludge during AQDS-mediated bioelectrochemical dechlorination of 1,2-dichloroethane to ethene. Bioresource Technol 214: 426-431.

Leitäo, P., Nouws, H., Danko, A.S., and Aulenta, F. (2017) Bioelectrochemical dechlorination of 1,2-DCA with an AQDS-functionalized cathode serving as electron donor. Fuel Cells (In Press) DOI: 10.1002/fuce.201700045.

Li, D.-B., Cheng, Y.-Y., Wu, C., Li, W.-W., Li, N., Yang, Z.-C., et al. (2014) Selenite reduction by Shewanella oneidensis MR-1 is mediated by fumarate reductase in periplasm. Sci Rep 4: 3735.

Liamleam, W., and Annachhatre, A.P. (2007) Electron donors for biological sulfate reduction. Biotechnol Adv 25: 452-463.

Logan, B.E., Hamelers, B., Rozendal, R., Schröder, U., Keller, J., Freguia, S., et al. (2006) Microbial fuel cells: methodology and technology. Environ Sci Technol 40: 5181-5192.

Lovley, D.R., Ueki, T., Zhang, T., Malvankar, N.S., Shrestha, P.M., Flanagan, K.A., et al. (2011) Geobacter the microbe electric's physiology, ecology, and practical applications. Adv Microb Physiol 59: 1-100.

Mastrocicco, M., Giambastiani, B.M.S., and Colombani, N. (2013) Ammonium occurrence in a salinized lowland coastal aquifer (ferrara, italy). Hydrol Process 27: 34953501.

McAdam, E.J., and Judd, S.J. (2006) A review of membrane bioreactor potential for nitrate removal from drinking water. Desalination 196: 135-148.

Mencio, A., Boy, M., and Mas-Pla, J. (2011) Analysis of vulnerability factors that control nitrate occurrence in natural springs (Osona Region, NE Spain). Sci Total Environ 409: 3049-3058.

Mieseler, M., Atiyeh, M.N., Hernandez, H.H., and Ahmad, F. (2013) Direct enrichment of perchlorate-reducing microbial community for efficient electroactive perchlorate reduction in biocathodes. J Ind Microbiol Biotechnol 40: 1321-1327.

Moran, M.J., Zogorski, J.S., and Squillace, P.J. (2007) Chlorinated solvents in groundwater of the United States. Environ Sci Technol 41: 74-81.

Mu, Y., Rozendal, R.A., Rabaey, K., and Keller, J. (2009) Nitrobenzene removal in bioelectrochemical systems. Environ Sci Technol 43: 8690-8695.

Mulligan, C.N., Yong, R.N., and Gibbs, B.F. (2001) Remediation technologies for metal-contaminated soils and groundwater: an evaluation. Eng Geol 60: 193-207.

N'Guessan, A.L., Elifantz, H., Nevin, K.P., Mouser, P.J., Methe, B., Woodard, T.L., et al. (2010) Molecular analysis of phosphate limitation in Geobacteraceae during the bioremediation of a uranium-contaminated aquifer. ISME J 4: 253 266.

Nguyen, V.K., Hong, S., Park, Y., Jo, K., and Lee, T. (2015) Autotrophic denitrification performance and bacterial community at biocathodes of bioelectrochemical systems with either abiotic or biotic anodes. J Biosci Bioeng 119: 180187.

Nguyen, V.K., Park, Y., Yang, H., Yu, J., and Lee, T. (2016a) Effect of the cathode potential and sulfate ions on nitrate reduction in a microbial electrochemical denitrifi-cation system. J Ind Microbiol Biotechnol 43: 783-793.

Nguyen, V.K., Park, Y., Yu, J., and Lee, T. (2016b) Bioelec-trochemical denitrification on biocathode buried in simulated aquifer saturated with nitrate-contaminated groundwater. Environ Sci Pollut Res 23: 15443-15451.

Nguyen, V.K., Park, Y., Yu, J., and Lee, T. (2016c) Microbial selenite reduction with organic carbon and electrode as sole electron donor by a bacterium isolated from domestic wastewater. Bioresource Technol 212: 182-189.

Nguyen, V.K., Park, Y., Yu, J., and Lee, T. (2016d) Simultaneous arsenite oxidation and nitrate reduction at the electrodes of bioelectrochemical systems. Environ Sci Pollut Res 23: 19978-19988.

Nguyen, V.K., Tran, H.T., Park, Y., Yu, J. and Lee, T. (2017) Microbial arsenite oxidation with oxygen, nitrate, or an electrode as the sole electron acceptor. J Ind Microbiol Biotechnol 44: 857-868.

Nozawa-Inoue, M., Jien, M., Yang, K., Rolston, D.E., Hris-tova, K.R., and Scow, K.M. (2011) Effect of nitrate, acetate and hydrogen on native perchlorate-reducing microbial communities and their activity in vadose soil. FEMS Microbiol Ecol 76: 278-288.

Obiri-Nyarko, F., Grajales-Mesa, S.J., and Malina, G. (2014) An overview of permeable reactive barriers for in situ sustainable groundwater remediation. Chemosphere 111: 243-259.

Orellana, R., Leavitt, J.J., Comolli, L.R., Csencsits, R., Janot, N., Flanagan, K.A., etal. (2013) U(VI) reduction by diverse outer surface c-type cytochromes of Geobac-ter sulfurreducens. Appl Environ Microbiol 79: 63696374.

Oremland, R.S., and Stolz, J.F. (2005) Arsenic, microbes and contaminated aquifers. Trends Microbiol 13: 45-49.

Ortiz-Bernad, I., Anderson, R.T., Vrionis, H.A., and Lovley, D.R. (2004) Vanadium respiration by Geobacter metallire-ducens: Novel strategy for in situ removal of vanadium from groundwater. Appl Environ Microbiol 70: 3091-3095.

Park, H.Il., Kim, D.K., Choi, Y.-J., and Pak, D. (2005) Nitrate reduction using an electrode as direct electron donor in a biofilm-electrode reactor. Process Biochem 40: 33833388.

Park, H.Il., Kim, J.S., Kim, D.K., Choi, Y.-J., and Pak, D. (2006) Nitrate-reducing bacterial community in a biofilm-electrode reactor. Enzyme Microb Technol 39: 453-458.

Pous, N., Puig, S., Coma, M., Balaguer, M.D., and Colprim, J. (2013) Bioremediation of nitrate-polluted groundwater in a microbial fuel cell. J Chem Technol Biotechnol 88: 1690-1696.

Pous, N., Casentini, B., Rossetti, S., Fazi, S., Puig, S., and Aulenta, F. (2015a) Anaerobic arsenite oxidation with an electrode serving as the sole electron acceptor: a novel approach to the bioremediation of arsenic-polluted groundwater. J Hazard Mater 283: 617-622.

Pous, N., Koch, C., Vila-Rovira, A., Balaguer, M.D., Colprim, J., Mühlenberg, J., etal. (2015b) Monitoring and engineering reactor microbiomes of denitrifying bioelectrochemical systems. RSC Adv 5: 68326-68333.

Pous, N., Puig, S., Balaguer, M.D., and Colprim, J. (2015c) Cathode potential and anode electron donor evaluation for a suitable treatment of nitrate-contaminated groundwater in bioelectrochemical systems. Chem Eng J 263: 151159.

Pous, N., Puig, S., Balaguer, M.D., and Colprim, J. (2017) Effect of hydraulic retention time and substrate availability in denitrifying bioelectrochemical systems. Environ Sci: Water Res Technol 3: 922-929.

Pozo, G., Jourdin, L., Lu, Y., Ledezma, P., Keller, J., and Freguia, S. (2015) Methanobacterium enables high rate electricity-driven autotrophic sulfate reduction. RSC Adv 5: 89368-89374.

Pozo, G., Jourdin, L., Lu, Y., Keller, J., Ledezma, P., and Freguia, S. (2016) Cathodic biofilm activates electrode surface and achieves efficient autotrophic sulfate reduction. Electrochim Acta 213: 66-74.

Prosnansky, M., Sakakibara, Y., and Kuroda, M. (2002) High-rate denitrification and SS rejection by biofilm-elec-trode reactor (BER) combined with microfiltration. Water Res 36: 4801-4810.

Puig, S., Serra, M., Vilar-Sanz, A., Cabró, M., Bañeras, L., Colprim, J., and Balaguer, M.D. (2011) Autotrophic nitrite removal in the cathode of microbial fuel cells. Bioresource Technol 102: 4462-4467.

Puig, S., Coma, M., Desloover, J., Boon, N., Colprim, J., and Balaguer, M.D. (2012) Autotrophic denitrification in microbial fuel cells treating low ionic strength waters. Environ Sci Technol 46: 2309-2315.

Rabaey, K., Van de Sompel, K., Maignien, L., Boon, N., Ael-terman, P., Clauwaert, P., et al. (2006) Microbial fuel cells for sulfide removal. Environ Sci Technol 40: 5218-5224.

Rabaey, K., Angenent, L., derSchroö, U. and Keller, J. (2009) Bioelectrochemical systems: from extracellular electron transfer to biotechnological application. London : International water association publishing.

Rakoczy, J., Feisthauer, S., Wasmund, K., Bombach, P., Neu, T.R., Vogt, C., and Richnow, H.H. (2013) Benzene and sulfide removal from groundwater treated in a microbial fuel cell. Biotechnol Bioeng 110: 3104-3113.

Ricardo, A.R., Carvalho, G., Velizarov, S., Crespo, J.G., and Reis, M.A.M. (2012) Kinetics of nitrate and perchlo-rate removal and biofilm stratification in an ion exchange membrane bioreactor. Water Res 46: 4556-4568.

Rodrigo, J., Boltes, K., and Esteve-Nuñez, A. (2014) Micro-bial-electrochemical bioremediation and detoxification of dibenzothiophene-polluted soil. Chemosphere 101: 6165.

Rooney-Varga, J.N., Anderson, R.T., Fraga, J.L., Ringelberg, D., and Lovley, D.R. (1999) Microbial communities associated with anaerobic benzene degradation in a petroleum-contaminated aquifer. Appl Environ Microbiol 65: 3056-3063.

Rosenbaum, M., Aulenta, F., Villano, M., and Angenent, L.T. (2011) Cathodes as electron donors for microbial metabolism: which extracellular electron transfer mechanisms are involved? Bioresource Technol 102: 324-333.

Sakakibara, Y., and Kuroda, M. (1993) Electric prompting and control of denitrification. Biotechnol Bioeng 42: 535537.

Sander, E.M., Virdis, B., and Freguia, S. (2015) Dissimila-tory nitrate reduction to ammonium as an electron sink during cathodic denitrification. RSC Adv 5: 86572-86577.

Santini, M., Marzorati, S., Fest-Santini, S., Trasatti, S., and Cristiani, P. (2016) Carbonate scale deactivating the biocathode in a microbial fuel cell. J Power Sources 356: 400-407.

Scheiber, L., Ayora, C., Vázquez-Sunó., E., Cendón, D.I., Soler, A. and Baquero, J.C. (2016) Origin of high ammonium, arsenic and boron concentrations in the proximity of

a mine: natural vs. anthropogenic processes. Sci Total Environ 541: 655-666.

Schröder, U., Harnisch, F., and Angenent, L.T. (2015) Microbial electrochemistry and technology: terminology and classification. Energy Environ Sci 8: 513-519.

Shea, C., Clauwaert, P., Verstraete, W., and Nerenberg, R. (2008) Adapting a denitrifying biocathode for perchlorate reduction. Water Sci Technol 58: 1941-1946.

Shelobolina, E.S., Vrionis, H.A., Findlay, R.H., and Lovley, D.R. (2008) Geobacter uraniireducens sp. nov., isolated from subsurface sediment undergoing uranium bioremedi-ation. Int J Syst Evol Microbiol 58: 1075-1078.

Shen, J., Huang, L., Zhou, P., Quan, X., and Puma, G.L. (2017) Correlation between circuital current, Cu(II) reduction and cellular electron transfer in EAB isolated from Cu (II)-reduced biocathodes of microbial fuel cells. Bioelectro-chemistry 114: 1-7.

Shukla, A.K., Upadhyay, S.N., and Dubey, S.K. (2014) Current trends in trichloroethylene biodegradation: a review. Crit Rev Biotechnol 34: 101-114.

Soares, M.I.M. (2000) Biological denitrification of groundwater. Water Air Soil Pollut 123: 183-193.

Song, T.-S., Jin, Y., Bao, J., Kang, D., and Xie, J. (2016) Graphene/biofilm composites for enhancement of hexava-lent chromium reduction and electricity production in a biocathode microbial fuel cell. J Hazard Mater 317: 7380.

Sprague, L.A., Hirsch, R.M., and Aulenbach, B.T. (2011) Nitrate in the Mississippi River and its tributaries, 1980 to 2008: are we making progress? Environ Sci Technol 45: 7209-7216.

Squillace, P.J., Scott, J.C., Moran, M.J., Nolan, B.T., and Kolpin, D.W. (2002) VOCs, pesticides, nitrate, and their mixtures in groundwater used for drinking water in the United States. Environ Sci Technol 36: 1923-1930.

Stolz, J.F., Basu, P., Santini, J.M., and Oremland, R.S. (2006) Arsenic and selenium in microbial metabolism. Annu Rev Microbiol 60: 107-130.

Strycharz, S.M., Woodard, T.L., Johnson, J.P., Nevin, K.P., Sanford, R.A., Löffler, F.E., and Lovley, D.R. (2008) Graphite electrode as a sole electron donor for reductive dechlorination of tetrachloroethene by Geobacter lovleyi. Appl Environ Microbiol 74: 5943-5947.

Strycharz, S.M., Gannon, S.M., Boles, A.R., Franks, A.E., Nevin, K.P., and Lovley, D.R. (2010) Reductive dechlorination of 2-chlorophenol by Anaeromyxobacter dehaloge-nans with an electrode serving as the electron donor. Environ Microbiol Rep 2: 289 294.

Su, W., Zhang, L., Tao, Y., Zhan, G., and Li, D.D. (2012) Sulfate reduction with electrons directly derived from electrodes in bioelectrochemical systems. Electrochem Commun 22: 37^0.

Susarla, S., Collette, T.W., Garrison, A.W., Wolfe, N.L., and Mccutcheon, S.C. (1999) Perchlorate identification in fertilizers. Environ Sci Technol 33: 3469-3472.

Ter Heijne, A., Liu, F., Weijden, R.Van.Der., Weijma, J., Buisman, C.J.N. and Hamelers, H.V.M. (2010) Copper recovery combined with electricity production in a microbial fuel cell. Environ Sci Technol 44: 4376^381.

Teuten, E.L., Saquing, J.M., Knappe, D.R.U., Barlaz, M.A., Jonsson, S., Björn, A., et al. (2009) Transport and release

of chemicals from plastics to the environment and to wildlife. Philos Trans R Soe B Biol Sei 364: 2027-2045.

Thrash, J.C., Van Trump, J.I., Weber, K.A., Miller, E., Achenbach, L.A., and Coates, J.D. (2007) Electrochemical stimulation of microbial perchlorate reduction. Environ Sei Teehnol 41: 1740-1746.

Tong, Y., and He, Z. (2013) Nitrate removal from groundwater driven by electricity generation and heterotrophic denitrification in a bioelectrochemical system. J Hazard Mater 262: 614-619.

Tong, S., Zhang, B., Feng, C., Zhao, Y., Chen, N., Hao, C., et al. (2013) Characteristics of heterotrophic/biofilm-elec-trode autotrophic denitrification for nitrate removal from groundwater. Bioresouree Teehnol 148: 121-127.

Turney, G.L., and Goerlitz, D.F. (1990) Organic contamination of ground water at gas works park, Seattle, Washington. Groundw Monit Remediat 10: 187-198.

Twomey, K.M., Stillwell, A.S., and Webber, M.E. (2010) The unintended energy impacts of increased nitrate contamination from biofuels production. J Environ Monit 12: 218224.

Urbansky, E.T., Magnuson, M.L., Kelty, C.A., Gu, B., Brown, G.M., Susarla, S., et al. (2000) Comment on "perchlorate identification in fertilizers" and the subsequent addition/ correction (multiple letters). Environ Sei Teehnol 34: 4452 4454.

Urbansky, E.T., Brown, S.K., Magnuson, M.L., and Kelty, C.A. (2001) Perchlorate levels in samples of sodium nitrate fertilizer derived from Chilean caliche. Environ Pollut 112: 299-302.

Van Halem, D., Bakker, S.A., Amy, G.L., and van Dijk, J.C. (2009) Arsenic in drinking water: a worldwide water quality concern for water supply companies. Drink Water Eng Sei 2: 29-34.

Velasquez-Orta, S.B., Werner, D., Varia, J.C., and Mgana, S. (2017) Microbial fuel cells for inexpensive continuous in-situ monitoring of groundwater quality. Water Res 117: 9-17.

Verdini, R., Aulenta, F., De Tora, F., Lai, A., and Majone, M. (2015) Relative contribution of set cathode potential and external mass transport on TCE dechlorination in a continuous-flow bioelectrochemical reactor. Chemosphere 136: 72-78.

Vilajeliu-Pons, A., Puig, S., Pous, N., Salcedo-Davila, I., Bañeras, L., Balaguer, M.D., and Colprim, J. (2015) Microbiome characterization of MFCs used for the treatment of swine manure. J Hazard Mater 288: 60-68.

Vilajeliu-Pons, A., Puig, S., Salcedo-Dávila, I., Balaguer, M.D. and Colprim, J. (2017) Long-term assessment of six-stacked scaled-up MFCs treating swine manure with different electrode materials. Environ Sei: Water Res Technol 3: 947 959.

Virdis, B., Rabaey, K., Yuan, Z., and Keller, J. (2008) Microbial fuel cells for simultaneous carbon and nitrogen removal. Water Res 42: 3013-3024.

Virdis, B., Rabaey, K., Rozendal, R.A., Yuan, Z., and Keller, J. (2010) Simultaneous nitrification, denitrification and carbon removal in microbial fuel cells. Water Res 44: 29702980.

Voyevoda, M., Geyer, W., Mosig, P., Seeger, E.M., and Mothes, S. (2012) Evaluation of the effectiveness of

different methods for the remediation of contaminated groundwater by determining the petroleum hydrocarbon content. Clean - Soil, Air, Water 40: 817-822.

Wang, A.-J., Cheng, H.-Y., Liang, B., Ren, N.-Q., Cui, D., Lin, N., et al. (2011) Efficient reduction of nitrobenzene to aniline with a biocatalyzed cathode. Environ Sci Technol 45: 10186-10193.

Wang, Z., Gao, M., Zhang, Y., She, Z., Ren, Y., Wang, Z., and Zhao, C. (2014) Perchlorate reduction by hydrogen autotrophic bacteria in a bioelectrochemical reactor. J Environ Manage 142: 10-16.

Wardman, C., Nevin, K.P., and Lovley, D.R. (2014) Realtime monitoring of subsurface microbial metabolism with graphite electrodes. Front Microbiol 5: 621.

Wasik, E., Bohdziewicz, J., and Blaszcyk, M. (2001) Removal of nitrates from ground water by a hybrid process of biological denitrification and microfiltration membrane. Process Biochem 37: 57-64.

Webster, D.P., TerAvest, M.A., Doud, D.F.R., Chakravorty, A., Holmes, E.C., Radens, C.M., et al. (2014) An arsenic-specific biosensor with genetically engineered Shewanella oneidensis in a bioelectrochemical system. Biosens Bio-electron 62: 320-324.

Wei, M., Harnisch, F., Vogt, C., Ahlheim, J., Neue, T.R., and Richnowa, H.H. (2015a) Harvesting electricity from benzene and ammonium-contaminated groundwater using a microbial fuel cell with an aerated cathode. RSC Adv 5: 5321-5330.

Wei, M., Rakoczy, J., Vogt, C., Harnisch, F., Schumann, R., and Richnow, H.H. (2015b) Enhancement and monitoring of pollutant removal in a constructed wetland by microbial electrochemical technology. Bioresource Technol 196: 490-499.

Wen, Q., Yang, T., Wang, S., Chen, Y., Cong, L., and Qu, Y. (2013) Dechlorination of 4-chlorophenol to phenol in bioelectrochemical systems. J Hazard Mater 244-245: 743-749.

Williams, K.H., Nevin, K.P., Franks, A., Englert, A., Long, P.E., and Lovley, D.R. (2010) Electrode-based approach for monitoring in situ microbial activity during subsurface bioremediation. Environ Sci Technol 44: 47-54.

Williams, K.H., Bargar, J.R., Lloyd, J.R., and Lovley, D.R.

(2013) Bioremediation of uranium-contaminated ground-water: a systems approach to subsurface biogeochem-istry. Curr Opin Biotechnol 24: 489-497.

Wu, X., Zhu, X., Song, T., Zhang, L., Jia, H., and Wei, P. (2015) Effect of acclimatization on hexavalent chromium reduction in a biocathode microbial fuel cell. Bioresour Technol 180: 185-191.

Xafenias, N., Zhang, Y., and Banks, C.J. (2013) Enhanced performance of hexavalent chromium reducing cathodes in the presence of Shewanella oneidensis MR-1 and lac-tate. Environ Sci Technol 47: 4512-4520.

Xafenias, N., Zhang, Y., and Banks, C.J. (2015) Evaluating hexavalent chromium reduction and electricity production in microbial fuel cells with alkaline cathodes. Int J Environ Sci Technol 12: 2435-2446.

Xie, D., Yu, H., Li, C., Ren, Y., Wei, C., and Feng, C.

(2014) Competitive microbial reduction of perchlorate and

nitrate with a cathode directly serving as the electron donor. Electrochim Acta 133: 217-223.

Xue, A., Shen, Z.-Z., Zhao, B., and Zhao, H.-Z. (2013) Arsenite removal from aqueous solution by a microbial fuel cell-zerovalent iron hybrid process. J Hazard Mater 261C: 621-627.

Yan, F., and Reible, D. (2015) Electro-bioremediation of contaminated sediment by electrode enhanced capping. J Environ Manage 155: 154-161.

Yu, J., Park, Y., Nguyen, V.K., and Lee, T. (2016) PCE dechlorination by non-Dehalococcoides in a microbial electrochemical system. J Ind Microbiol Biotechnol 43: 1095-1103.

Yun, H., Liang, B., Kong, D.-Y., Cheng, H.-Y., Li, Z.-L., Gu, Y.-B., et al. (2017) Polarity inversion of bioanode for bio-cathodic reduction of aromatic pollutants. J Hazard Mater 331: 280-288.

Zeppenfeld, K. (2011) Electrochemical removal of calcium and magnesium ions from aqueous solutions. Desalination 277: 99-105.

Zhan, G., Zhang, L., Li, D., Su, W., Tao, Y., and Qian, J. (2012) Autotrophic nitrogen removal from ammonium at low applied voltage in a single-compartment microbial electrolysis cell. Bioresource Technol 116: 271-277.

Zhan, G., Zhang, L., Tao, Y., Wang, Y., Zhu, X., and Li, D.

(2014) Anodic ammonia oxidation to nitrogen gas catalyzed by mixed biofilms in bioelectrochemical systems. Electrochim Acta 135: 345-350.

Zhang, Y., and Angelidaki, I. (2013) A new method for in situ nitrate removal from groundwater using submerged microbial desalination-denitrification cell (SMDDC). Water Res 47: 1827-1836.

Zhang, T., Gannon, S.M., Nevin, K.P., Franks, A.E., and Lovley, D.R. (2010) Stimulating the anaerobic degradation of aromatic hydrocarbons in contaminated sediments by providing an electrode as the electron acceptor. Environ Microbiol 12: 1011-1020.

Zhang, T., Bain, T.S., Nevin, K.P., Barlett, M.A., and Lovley, D.R. (2012) Anaerobic benzene oxidation by Geobacter species. Appl Environ Microbiol 78: 8304-8310.

Zhang, T., Tremblay, P.-L., Chaurasia, A.K., Smith, J.A., Bain, T.S., and Lovley, D.R. (2014) Identification of genes specifically required for the anaerobic metabolism of benzene in Geobacter metallireducens. Front Microbiol 5: 245.

Zhang, B., Tian, C., Liu, Y., Hao, L., Liu, Y., Feng, C., etal.

(2015) Simultaneous microbial and electrochemical reductions of vanadium (V) with bioelectricity generation in microbial fuel cells. Bioresource Technol 179C: 91-97.

Zhao, H.-P., Van Ginkel, S., Tang, Y., Kang, D.-W., Rittmann, B., and Krajmalnik-Brown, R. (2011) Interactions between perchlorate and nitrate reductions in the biofilm of a hydrogen-based membrane biofilm reactor. Environ Sci Technol 45: 10155-10162.

Zhu, T., Zhang, Y., Bu, G., Quan, X., and Liu, Y. (2016) Producing nitrite from anodic ammonia oxidation to accelerate anammox in a bioelectrochemical system with a given anode potential. Chem Eng J 291: 184-191.