kinetics during the redox biotransformation of pollutants mediated by immobilized and soluble humic...

9
ENVIRONMENTAL BIOTECHNOLOGY Kinetics during the redox biotransformation of pollutants mediated by immobilized and soluble humic acids Francisco J. Cervantes & Claudia M. Martínez & Jorge Gonzalez-Estrella & Arturo Márquez & Sonia Arriaga Received: 13 March 2012 / Accepted: 2 April 2012 / Published online: 9 May 2012 # Springer-Verlag 2012 Abstract The aim of this study was to elucidate the kinetic constraints during the redox biotransformation of the azo dye, Reactive Red 2 (RR2), and carbon tetrachloride (CT) mediat- ed by soluble humic acids (HA s ) and immobilized humic acids (HA i ), as well as by the quinoid model compounds, anthraquinone-2,6-disulfonate (AQDS) and 1,2-naphthoqui- none-4-sulfonate (NQS). The microbial reduction of both HA s and HA i by anaerobic granular sludge (AGS) was the rate- limiting step during decolorization of RR2 since the reduction of RR2 by reduced HA i proceeded at more than three orders of magnitute faster than the electron-transferring rate observed during the microbial reduction of HA i by AGS. Similarly, the reduction of RR2 by reduced AQDS proceeded 1.6- and 1.9- fold faster than the microbial reduction of AQDS by AGS when this redox mediator (RM) was supplied in soluble and immobilized form, respectively. In contrast, the reduction of NQS by AGS occurred 1.6- and 19.2-fold faster than the chemical reduction of RR2 by reduced NQS when this RM was supplied in soluble and immobilized form, respectively. The microbial reduction of HA s and HA i by a humus-reducing consortium proceeded 1,400- and 790-fold faster than the transfer of electrons from reduced HA s and HA i , respectively, to achieve the reductive dechlorination of CT to chloroform. Overall, the present study provides elucidation on the rate- limiting steps involved in the redox biotransformation of priority pollutants mediated by both HA s and HA i and offers technical suggestions to overcome the kinetic restrictions identified in the redox reactions evaluated. Keywords Humus . Immobilization . Recalcitrant pollutants . Redox mediator . Wastewater Introduction During the last two decades, evidence has been reported indicating that humus, the most abundant organic matter accumulating in terrestrial and aquatic environments, has active roles accelerating the redox biotransformation of azo dyes (Rau et al. 2002; Cervantes et al. 2011a; Liu et al. 2011), nitroaromatics (Bhushan et al. 2006; Kwon and Finneran 2006), polychlorinated compounds (Collins and Picardal 1999; Cervantes et al. 2004, 2011a; Alvarez et al. 2011), among other electron-accepting contaminants by serving as redox mediators (RMs) (Van der Zee and Cervantes 2009). The redox-mediating properties of hu- mus have mainly been attributed to quinone moieties (Scott et al. 1998), which are very abundant in humus; thus, quinoid model compounds have extensively been used as humus analogs in several studies (Van der Zee and Cervantes 2009). The information available in the literature provides a broad perspective of the different experimental conditions (pH, temperature, type and concentration of RM, type and concentration of electron donor, and redox potential), which affect the redox-mediating capacity of humus and quinones Electronic supplementary material The online version of this article (doi:10.1007/s00253-012-4081-5) contains supplementary material, which is available to authorized users. F. J. Cervantes : C. M. Martínez : J. Gonzalez-Estrella : A. Márquez : S. Arriaga División de Ciencias Ambientales, Instituto Potosino de Investigación Científica y Tecnológica (IPICyT), Camino a la Presa San José 2055, Col. Lomas 4ª Sección, San Luis Potosí 78216, Mexico F. J. Cervantes (*) Department of Biotechnology, Norwegian University of Science and Technology (NTNU), 7491 Trondheim, Norway e-mail: [email protected] Appl Microbiol Biotechnol (2013) 97:26712679 DOI 10.1007/s00253-012-4081-5

Upload: claudia-m-martinez

Post on 08-Dec-2016

216 views

Category:

Documents


2 download

TRANSCRIPT

Page 1: Kinetics during the redox biotransformation of pollutants mediated by immobilized and soluble humic acids

ENVIRONMENTAL BIOTECHNOLOGY

Kinetics during the redox biotransformation of pollutantsmediated by immobilized and soluble humic acids

Francisco J. Cervantes & Claudia M. Martínez &

Jorge Gonzalez-Estrella & Arturo Márquez &

Sonia Arriaga

Received: 13 March 2012 /Accepted: 2 April 2012 /Published online: 9 May 2012# Springer-Verlag 2012

Abstract The aim of this study was to elucidate the kineticconstraints during the redox biotransformation of the azo dye,Reactive Red 2 (RR2), and carbon tetrachloride (CT) mediat-ed by soluble humic acids (HAs) and immobilized humic acids(HAi), as well as by the quinoid model compounds,anthraquinone-2,6-disulfonate (AQDS) and 1,2-naphthoqui-none-4-sulfonate (NQS). The microbial reduction of both HAs

and HAi by anaerobic granular sludge (AGS) was the rate-limiting step during decolorization of RR2 since the reductionof RR2 by reduced HAi proceeded at more than three orders ofmagnitute faster than the electron-transferring rate observedduring the microbial reduction of HAi by AGS. Similarly, thereduction of RR2 by reduced AQDS proceeded 1.6- and 1.9-fold faster than the microbial reduction of AQDS by AGSwhen this redox mediator (RM) was supplied in soluble andimmobilized form, respectively. In contrast, the reduction ofNQS by AGS occurred 1.6- and 19.2-fold faster than thechemical reduction of RR2 by reduced NQS when this RMwas supplied in soluble and immobilized form, respectively.The microbial reduction of HAs and HAi by a humus-reducingconsortium proceeded 1,400- and 790-fold faster than the

transfer of electrons from reduced HAs and HAi, respectively,to achieve the reductive dechlorination of CT to chloroform.Overall, the present study provides elucidation on the rate-limiting steps involved in the redox biotransformation ofpriority pollutants mediated by both HAs and HAi and offerstechnical suggestions to overcome the kinetic restrictionsidentified in the redox reactions evaluated.

Keywords Humus . Immobilization . Recalcitrantpollutants . Redox mediator . Wastewater

Introduction

During the last two decades, evidence has been reportedindicating that humus, the most abundant organic matteraccumulating in terrestrial and aquatic environments, hasactive roles accelerating the redox biotransformation ofazo dyes (Rau et al. 2002; Cervantes et al. 2011a; Liu etal. 2011), nitroaromatics (Bhushan et al. 2006; Kwon andFinneran 2006), polychlorinated compounds (Collins andPicardal 1999; Cervantes et al. 2004, 2011a; Alvarez et al.2011), among other electron-accepting contaminants byserving as redox mediators (RMs) (Van der Zee andCervantes 2009). The redox-mediating properties of hu-mus have mainly been attributed to quinone moieties(Scott et al. 1998), which are very abundant in humus;thus, quinoid model compounds have extensively beenused as humus analogs in several studies (Van der Zee andCervantes 2009).

The information available in the literature provides abroad perspective of the different experimental conditions(pH, temperature, type and concentration of RM, type andconcentration of electron donor, and redox potential), whichaffect the redox-mediating capacity of humus and quinones

Electronic supplementary material The online version of this article(doi:10.1007/s00253-012-4081-5) contains supplementary material,which is available to authorized users.

F. J. Cervantes :C. M. Martínez : J. Gonzalez-Estrella :A. Márquez : S. ArriagaDivisión de Ciencias Ambientales, Instituto Potosino deInvestigación Científica y Tecnológica (IPICyT),Camino a la Presa San José 2055, Col. Lomas 4ª Sección,San Luis Potosí 78216, Mexico

F. J. Cervantes (*)Department of Biotechnology,Norwegian University of Science and Technology (NTNU),7491 Trondheim, Norwaye-mail: [email protected]

Appl Microbiol Biotechnol (2013) 97:2671–2679DOI 10.1007/s00253-012-4081-5

Page 2: Kinetics during the redox biotransformation of pollutants mediated by immobilized and soluble humic acids

(Van der Zee and Cervantes 2009). However, scarce in-formation has been reported to elucidate the mechanismsby which RMs are involved in different redox reactionsand the bottlenecks to consider during redox catalysis. Insome cases, the reduction of RMs by different microor-ganisms has been pointed out as the rate-limiting stepduring redox reactions (Rau et al. 2002). In contrast, otherstudies evidenced that the reduction of electron-acceptingcontaminants by reduced RMs was the rate-limiting step(Rau et al. 2002; Jiang and Kappler 2008; Martínez andCervantes 2012). Therefore, further research is demandedin order to explain the mechanisms limiting the catalyticeffects of RMs during the anaerobic biotransformation ofpriority pollutants.

One of the main limitations for applying humus andquinoid analogs as RM in wastewater treatment systems isthat continuous addition of RM should be supplied in orderto achieve increased conversion rates, which is economical-ly and environmentally unviable. An approach to eliminatethe prerequisite of continuous supply of RM is by immobi-lizing them in supporting matrixes within anaerobic bio-reactors. However, few attempts to apply immobilizedquinoid model compounds for the anaerobic conversion ofpriority pollutants have been reported (Guo et al. 2007; Li etal. 2008; Wang et al. 2009; Cervantes et al. 2010). Morerecently, immobilized humic acids (HA), which representthe fraction of humus containing the highest concentrationof quinones (Stevenson 1994), were demonstrated as effec-tive solid-phase RMs accelerating the redox biotransforma-tion of carbon tetrachloride (CT) and the azo modelcompound, Reactive Red 2 (RR2) (Cervantes et al. 2011a).The aim of the present study was to elucidate the kineticconstraints during the redox biotransformation of RR2 andCT mediated by immobilized and soluble HA, as well as bythe quinoid model compounds, anthraquinone-2,6-disulfo-nate (AQDS) and 1,2-naphthoquinone-4-sulfonate (NQS).An anion exchange resin (AER) previously used to immo-bilize HA and the quinoid model compounds (Cervantes etal. 2010, 2011a) was used as an immobilizing matrix. CTwas selected as a model contaminant because it is animportant intermediate for the production of other chem-icals, such as chlorinated paraffin wax and chlorofluoro-carbons, nowadays. Furthermore, contamination ofgroundwater and soils by CT still prevails in many pol-luted sites (Penny et al. 2010). RR2 was selected becauseit is a very recalcitrant azo compound commonly used torepresent reductive decolorization processes for textilewastewater treatment (Dos Santos et al. 2005; Pavlostathisand Beydilli 2005; Cervantes et al. 2010). RR2 and CThave widely been used to represent environmentally rele-vant redox reactions, such as decolorization of azo dyesand dehalogenation of polychlorinated contaminants (Vander Zee and Cervantes 2009).

Materials and methods

Chemicals and inocula

The AER used during the immobilization of HA was pro-vided by Rohm and Haas (AMBERJET 4600 CL, Philadel-phia, PA). This AER has previously been characterized(Cervantes et al. 2010). The source of HA used during thepresent study was leonardite obtained from the InternationalHumic Substances Society (catalog no. 1BS104L). RR2,AQDS, and NQS were obtained from Sigma-Aldrich andused as received from the supplier without further purifica-tion. All other chemicals were reagent grade and purchasedfrom J.T. Baker. Two different humus-reducing inocula wereused for measuring the microbial reduction of HA. A soilsample, referred to as humus-reducing soil (HRS), capableof reducing HA, which has a content of 8.43 % of volatilesolids (VS), was obtained from Poza Rica, Mexico. HRShas previously been phylogenetically characterized and hasbeen demonstrated to degrade benzene (Cervantes et al.2011b) and CT (Cervantes et al. 2011a) under humus-reducing conditions. An anaerobic granular sludge (AGS),which has previously been studied during the reductivedecolorization of RR2 (Cervantes et al. 2010, 2011a), wasalso evaluated.

Microbial reduction of HA and quinones

To determine the reduction rate of HA and quinones byhumus-reducing consortia (first step in Fig. 1), incuba-tions were performed in 120-mL glass serum bottles.Portions of 50 mL of basal medium, which compositionhas previously been described (Cervantes et al. 2010),were firstly dispensed in vials. Afterwards, inoculationtook place by adding 30 g VS/L or 0.1 g VS/L forincubations conducted with AGR and HRS, respectively.Vials were sealed with rubber stoppers and aluminumcaps. Glucose (2 g/L) was provided as an external elec-tron donor. The pH was controlled at 7.2 ± 0.1 by adding5,000 mg/L of NaHCO3 and saturating the headspacewith N2/CO2 (80/20 %). HA were supplied in soluble(HAs) or immobilized (HAi) form with an equivalentamount to have 1.2 g total organic carbon (TOC)/L inall cases. Both HAs and HAi were sulfonated as previ-ously described (Yudov et al. 2005) to increase solubilityand achieve immobilization, respectively. Furthermore,HAi were immobilized in an AER as previously described(Cervantes et al. 2011a). Incubations were also performedwith the quinoid model compounds, AQDS and NQS, asterminal electron acceptors both in soluble and immobilizedform with an initial concentration of 4.8 mM in all cases.Experiments were set up in triplicate and incubated at 25 °Cin the dark under mild shaking (~200 rpm). All experimental

2672 Appl Microbiol Biotechnol (2013) 97:2671–2679

Page 3: Kinetics during the redox biotransformation of pollutants mediated by immobilized and soluble humic acids

conditions were established in order to elucidate the rate-limiting step in previously reported experiments (Cervanteset al. 2010, 2011a). Reduction of HAs, HAi, AQDS, and NQSwas followed over time as described below.

Reduction rates were determined on the maximum slopeobserved on linear regressions considering at least threeexperimental sampling points. The coefficient of determina-tion (R2) was higher than 0.9 in all rates calculated.

Chemical reduction of CT and RR2 by reduced HAand hydroquinones

The reduction rate of RR2 and CT by previously reducedHA and hydroquinones (second step in Fig. 1) was alsomeasured. HAs and HAi (1.2 g TOC/L), as well as AQDSand NQS (4.8 mM), were chemically reduced in basalmedium by bubbling H2/CO2 (80/20 %) in the presence of

H2N

N

N

N

Cl

Cl

NH

-O3S SO3-

NH2

N

N

SO3--O3S

N

N

N

Cl

Cl

NHB

acte

ria

Reduced electron donor

Oxidized electron donor

C

Cl

Cl

Cl

Cl

Carbon Tetrachloride

CH2

Cl

Cl

Dichloro Methane

CH

Cl

Cl

Cl

Chloroform

Reactive Red 2

HOOC OH

OH

OH

O

HO

O

NCH

HO

OH

R

HOOC

O

HN

HO

CH

R

C

O

NH

R

-O3S

SO3-

-O3S

N+

H3C

CH3

COOHHO

O

O

O

OH

O

N CH

O

O

R

COOH

O

NH

OH

CH

R

C

O

HN

R

SO3-

-O3S

SO3-

N+

CH3

CH3

Aromatic Amines

H3C

H3C

+

+

Fig. 1 Proposed mechanisms involved in the reductive biotransforma-tion of Reactive Red 2 and Carbon Tetrachloride by humus-reducingbacteria mediated by immobilized sulfonated humic acids (HA) on ananion exchange resin (AER). Immobilized HA serve as terminal elec-tron acceptors for humus-reducing bacteria supporting the anaerobic

oxidation of an external electron donor during the first step. ReducedHA then transfer the reducing equivalents to the electron-acceptingpollutants in the second step. Dashed squares indicate the interactionbetween sulfonate groups inserted in HA and the quaternary aminepresent at the AER

Appl Microbiol Biotechnol (2013) 97:2671–2679 2673

Page 4: Kinetics during the redox biotransformation of pollutants mediated by immobilized and soluble humic acids

palladium as a catalyst until complete reduction. After re-moving the catalyst and saturating the headspace of vialswith N2/CO2 (80/20 %), the initial concentrations of CT(100 μM) and RR2 (0.3 mM) were established by addingthe corresponding amount from anaerobic stock solutions.Control incubations were also performed in vials containingnonreduced HA or quinones in order to verify adsorptionof CT and RR2. CT dechlorination was determined bymeasuring the concentration of this contaminant and thedehalogenated products expected (chloroform (CF) anddichloromethane (DCM)) from previous studies (Cervanteset al. 2011a). The reductive decolorization of RR2 was deter-mined by quantifying the concentration of the azo dye asdescribed below.

Electron-transferring rates were determined on the max-imum slope observed on linear regressions considering atleast three sampling points. The coefficient of determination(R2) was higher than 0.9 in all rates calculated.

Analytical techniques

The concentrations of CT, CF, and DCM were measured bygas chromatography (Agilent Technologies 7890 series)coupled to a microcell electron capture detector as previous-ly described (Alvarez et al. 2011). Identification of volatilechlorinated compounds (CT, CF, and DCM) was confirmedin headspace samples using a gas chromatograph (AgilentTechnologies 7890) coupled to a mass selective detector(Agilent Technologies 5975-C), with the same column andchromatographic conditions.

Decolorization of RR2 was measured spectrophotometri-cally at 539 nm as previously described (Cervantes et al.2010). TOC concentrations were determined in previouslyfiltered samples (0.22 μm) in a TOC meter (Shimadzu Co.Model TOC-VCSN). VS concentrations were determinedaccording to standard methods.

The reduction of quinones was determined spectrophoto-metrically at 450 and 350 nm for AQDS and NQS, respec-tively, in an anaerobic hood with an atmosphere composedof N2/H2 (96/4 %). The reduction of immobilized quinones,HAs, and HAi was determined indirectly by the ferrozinetechnique under the same atmosphere described above.Vials were sacrificed at different incubation periods by add-ing Fe(III) citrate (10 mM). After 30 min reaction, subsam-ples were filtered (0.2-μm pore diameter filter) and taken todetermine the amount of Fe(II) produced. Preliminaryexperiments indicated that 30 min was enough time tocompletely transfer the electrons from HA and quinones toFe(III) citrate. All measurements were corrected for intrinsicFe(II) present in samples by using subsamples in which Fe(III) citrate was not added.

The presence of bulk functional groups in HAwas deter-mined directly in samples using a Nicolet-6700 Fourier

transform infrared (FTIR) spectrometer (Thermo Scientific)equipped with a Smart iTR accessory, which is a versatileaccessory of attenuated total reflectance with ZnSe crystal.A deuterated triglycine sulfate detector was used to collectthe IR spectra in the 4,000–650 cm−1 spectral range.

Results

Kinetics during the reductive decolorization of RR2

Previously, HA and quinones were demonstrated as effectiveRMs (both in soluble and immobilized form) accelerating thereductive decolorization of RR2 by AGS (Cervantes et al.2010, 2011a). In order to elucidate the kinetic limitationsduring the redox reactions involved in the mediated decolor-ization of RR2, the electron-transferring rates involved in bothsteps illustrated in Fig. 1 were determined. Kinetic studiesperformed with the quinoid model compound, NQS, revealedthat the transfer of electrons from the reduced form of NQS(NH2QS) to RR2 is the rate-limiting step when decolorizationis mediated by this RM regardless if it is soluble orimmobilized (supplementary data (SD), Fig. S1). Certainly,the reduction of NQS by AGS occurred 1.6- and 19.2-foldfaster than the chemical reduction of RR2 by NH2QS whenthis RM was supplied in soluble and immobilized form,respectively (Tables 1 and 2). In contrast, the reduction ofAQDS by AGS appeared as the rate-limiting step when the

Table 1 Reduction rates of quinoid model compounds and humicacids by anaerobic consortia

Consortium Electron acceptor Reduction rate(μEq/L-h)a

Specific reductionrate (μEq/g VS h)a

HRSb HAs 24.7±0.7 247±7

HRSb HAi 9.5±0.5 95.7±5

AGSb HAs 11.6±0.2 0.39±0.03

AGSb HAi 1.4±0.01 0.047±0.01

AGSc AQDSs 743±21 24.8±1.3

AGSc AQDSi 163±11 5.43±0.3

AGSc NQSs 331±22 11.03±0.7

AGSc NQSi 442±31 14.73±1.1

AGS anaerobic granular sludge, HRS humus-reducing soil, HAS solublehumic acids, HAi immobilized humic acids, AQDS anthraquinone-2,6-disulfonate, NQS 1,2-naphthoquinone-4-sulfonate, AQDSs solubleAQDS, AQDSi immobilized AQDS, NQSs soluble NQS, NQSi immo-bilized NQS, μEq microelectron equivalent, VS volatile solidsa Data represent average from triplicate determinations ± standarddeviationb Conditions applied as described in Cervantes et al. 2011a duringcarbon tetrachloride dechlorination by this consortiumc Conditions applied as described in Cervantes et al. 2010 during thereductive decolorization of Reactive Red 2

2674 Appl Microbiol Biotechnol (2013) 97:2671–2679

Page 5: Kinetics during the redox biotransformation of pollutants mediated by immobilized and soluble humic acids

decolorization of RR2 was mediated by this RM. Indeed,the reductive decolorization of RR2 by AH2QDS proceeded1.6- and 1.9-fold faster than the microbial reduction ofAQDS by AGS when this RM was supplied in solubleand immobilized form, respectively (Tables 1 and 2).

Further experiments were conducted with HAs and HAi

in order to elucidate the kinetic constraints in incubationsmediated by HA (SD, Fig. S2). In this case, the microbialreduction of HA by AGS (first step in Fig. 1) was the rate-limiting step, namely, the reduction of RR2 by HAs and HAi

(second step in Fig. 1) proceeded 5.5- and 1,380-fold fasterthan the microbial reduction of HAs and HAi by AGS,respectively (Tables 1 and 2).

Adsorption of RR2 on HAs, HAi, and on immobilizedquinones did not occur as evidenced in controls performedwith nonreduced HA and quinones in which no removal ofRR2 was observed (data not shown). Thus, RR2 removal wasdue to its reduction to aromatic amines as previously docu-mented under the same experimental conditions (Cervantes etal. 2010, 2011a).

Kinetics during the reductive dechlorination of CT

Further experiments were conducted in order to understandthe kinetic limitations involved during the reductive dechlori-nation of CTmediated by HAs and HAi, which has previouslybeen documented (Cervantes et al. 2011a). As observed withAGS, HRS reduced HAs faster than HAi (SD, Fig. S3).

Nevertheless, in both cases, the rate-limiting step was thedechlorination of CT to CF by reduced HA (Tables 1 and 2).In fact, the microbial reduction of HAs and HAi by HRSproceeded 1,400- and 790-fold faster than the transfer ofelectron from reduced HAs and HAi, respectively, to achievethe reductive dechlorination of CT to CF.

CF was the only product detected from CT dechlorina-tion accounting for 90 and 97 % of the CT reduced inincubations amended with HAs and HAi, respectively (SD,Fig. S4). No CT removal was observed in controls incubatedwith nonreduced HAs; however, 40 % of CT initially spikeddisappeared from nonreduced HAi-amended vials, whichwas due to adsorption of CT on HAi as previously observed(Cervantes et al. 2011a). A clear distinction between adsorp-tion and reduction of CT could be established as no CFproduction occurred in vials incubated with nonreducedHAi. Furthermore, adsorption of CT to HAi only occurredduring the first 24 h of incubation.

Two clear phases were observed during the reductivedechlorination of CT by HAs and HAi (SD, Fig. S4). InHAs-amended vials, a first-order dechlorination constant(KCT) of 8×10

−4 h−1 was observed during the first 72 h ofincubation, which increased ~10-fold (8.6×10−3 h−1) afterthis incubation period. Meanwhile, in HAi-amended vials, aKCT value of 5.8×10−3 h−1 was achieved in the incubationperiod between 24 and 120 h in which no further adsorptionof CT was detected in nonreduced HAi-amended vials (datanot shown).

Discussion

The aim of the present study was to elucidate the kineticlimitations governing the reductive biotransformation ofelectron-accepting pollutants mediated by HAs and HAi.The available literature provides plentiful information doc-umenting the enhancement of redox reactions involved dur-ing the biotransformation of several organic and inorganicpollutants by the application of RMs (Van der Zee andCervantes 2009). Nevertheless, scarce information is avail-able in order to elucidate the kinetic mechanisms limitingthe catalytic effects of RMs during the anaerobic biotrans-formation of priority pollutants.

Regarding the experiments to elucidate the kinetic con-straints during the mediated decolorization of RR2, resultsindicated that the microbial reduction of both HAs and HAi

by AGS was the rate-limiting step during the decolorizationprocess (first step in Fig. 1). The kinetic disparity between thismicrobially mediated step and the chemical reaction of RR2with reduced HA (second step in Fig. 1) was much strongerwhen HAi were used as RM. Certainly, the reductive decol-orization of RR2 by reduced HAi proceeded at more than threeorders of magnitute faster than the electron-transferring rate

Table 2 Reduction of carbon tetrachloride and Reactive Red 2 byreduced humic acids and hydroquinone model compounds

Reduction process Electron donor Reduction rate (μEq/L h)a

Decolorization of RR2b AH2QDSs 1.19×103±98

Decolorization of RR2b AH2QDSi 302±18

Decolorization of RR2b NH2QSs 213±11

Decolorization of RR2b NH2QSi 23±5.1

Decolorization of RR2c Reduced HAS 64.1±6.2

Decolorization of RR2c Reduced HAi 1.93×103±97

Dechlorination of CTc Reduced HAS 1.7×10−2±1×10−3

Dechlorination of CTc Reduced HAi 1.2×10−2±0.9×10−3

AH2QDSs soluble anthrahydroquinone-2,6-disulfonate, AH2QDSi·im-mobilized AH2QDS, NH2QSs soluble 1,2-naphthohydroquinone-4-sul-fonate, NH2QSi immobilized NH2QS, HAS soluble humic acids, HAi

immobilized humic acids, μEq microelectron equivalent, TOC totalorganic carbon, CT carbon tetrachloride, RR2, Reactive Red 2a Data represent average from triplicate determinations ± standarddeviation. Hydroquinones and humic acids were previously reducedin a H2/Pd reaction system before incubation and were supplied at4.8 mM and 1.2 g TOC/L, respectivelyb Conditions applied as described in Cervantes et al. 2010 during thereductive decolorization of RR2c Conditions applied as described in Cervantes et al. 2011a during thereductive decolorization of RR2 and dechlorination of CT

Appl Microbiol Biotechnol (2013) 97:2671–2679 2675

Page 6: Kinetics during the redox biotransformation of pollutants mediated by immobilized and soluble humic acids

observed during the microbial reduction of HAi by AGS(Tables 1 and 2). Since HAi have been proposed as solid-phase RMs to be applied in anaerobic wastewater treatmentsystems to accelerate the reductive biotransformation of recal-citrant pollutants (Cervantes et al. 2011a), strategies to over-come the kinetic limitations elucidated in the present studyshould be considered. Firstly, to accelerate the microbial re-duction of HAi in continuous bioreactors, a high concentrationof humus-reducing microorganisms (HRM) should be consid-ered, such as that usually applied in high-rate anaerobic treat-ment systems (Lettinga et al. 1980). It has previously beendemonstrated that enrichment and immobilization of quinone-reducing bacteria can be achieved in this kind of anaerobicbioreactors (Cervantes et al. 2003). Therefore, considering thelarge variation in humus-reducing activities reported in theliterature ranging from 0.048 to 333.3 microelectron equiva-lent (μEq)/g VS h with consortia derived from a wide diver-sity of environments (Bradley et al. 1998; Cervantes et al.2000, 2001, 2008, 2011b), it would be very practical to enrichand immobilize a high concentration of HRM in high-rate

bioreactors in order to enhance the first part of the redoxreactions illustrated in Fig. 1.

The microbial reduction of HAi by AGS proceeded eight-fold slower as compared to HAs (Table 1), suggesting thatmass transfer is also limiting the mediated decolorization ofRR2; thus, direct contact between HAi and HRM should bepromoted in anaerobic treatment systems by coimmobiliz-ing HA and HRM in AER, for instance, by forming abiofilm over HAi-coated AER. Direct contact between HAi

and HRM could in this way decrease the mass transferlimitations during the first step illustrated in Fig. 1. Recent-ly, it has been demonstrated that attachment of γ-Al2O3

nanoparticles coated with HA on the surface of HRM pro-moted a faster reduction rate of HAi as compared to thereduction of HAs (Alvarez and Cervantes 2011). On theother hand, if a biofilm of HRM is formed over HAi-coatedAER, then the reduction of electron-accepting contaminants,such as RR2 and CT, by microbially reduced HAi woulddepend on the diffusion of these contaminants through thebiofilm (Fig. 2).

CO2 + H2O

Glucose

R

O

O

Oxidized HA

Reduced HA

RR2 or CT

Biotransformationproducts

Biofilm

OH

OH

R

Biofilm

Electron flow

AER

HA

Fig. 2 Redox reactions involved in the reductive biotransformation ofReactive Red 2 (RR2) and carbon tetrachloride (CT) mediated byimmobilized humic acids (HA) on an anion exchange resin (AER).Humus-reducing bacteria present in the biofilm firstly reduce HA

coupled to the anaerobic oxidation of glucose. Reduced HA thentransfer the electrons to RR2 and CT, which diffuse through the biofilmfrom the liquid bulk

2676 Appl Microbiol Biotechnol (2013) 97:2671–2679

Page 7: Kinetics during the redox biotransformation of pollutants mediated by immobilized and soluble humic acids

On the other hand, mass transfer is not the only aspectdetermining the kinetics involved in the reductive decoloriza-tion of RR2 mediated by HA because it is expected that HAs

would have promoted a higher decolorization of RR2 com-pared to HAi, but the abiotic reduction of RR2 by HAs

occurred at a 30-fold lower rate compared to the reductivedecolorization achieved with HAi (Table 2). Previously, HAs

decreased the rate of decolorization of RR2 byAGS (Cervanteset al. 2011a), which was proposed to be the result of electro-static repulsion between the sulfonate groups presentboth in HAs and in RR2 (Fig. 1). Since decolorization experi-ments were performed at pH 7.2, it is expected that mostsulfonate groups (pKa101.76 and pKa207.2 (Tartar andGarretson 1941) prevailed negatively charged, thus repellingthe interaction between HAs and RR2. Results derived from

sludge incubations supplied with HAi support this conclusionfirstly because HAi were demonstrated as an effective solid-phase RMs during RR2 decolorization by AGS (Cervantes etal. 2011a); secondly, the spectral signal of sulfonate groups (at~1,150 cm−1 (Silverstein and Webster 1992) were no longerdetected in HAi once adsorbed on the surface of the AER(Fig. 3), presumably because they were interacting with qua-ternary amines, which serve as ion-exchange functionalgroups in the AER (Fig. 1). Furthermore, nonsulfonated HAs

(NSHAs) were previously demonstrated as effective RM dur-ing the reductive decolorization of RR2 by AGS (Cervantes etal. 2011a). NSHAs did not have incorporated sulfonate groupsin its structure; hence, the electrostatic repulsion betweenNSHAs and RR2 would have been weaker than that hypoth-esized for the HAs–RR2 interaction, thus allowing the role of

NSHA on the AER

SHA on the AER

NSHA

SHA

A

B

Fig. 3 FTIR spectra of SHAand NSHA immobilized on theanion exchange resinAMBERJET 4600 CL (a) andFTIR spectra of SHA and NSHAsamples before immobilization(b). Arrows indicate the spectralsignals for quinone (1,700–1,630 cm−1) and sulfonate(1,200–1,100 cm−1) groups.SHA sulfonated humic acids,NSHA nonsulfonated humicacids, FTIR Fourier transforminfrared

Appl Microbiol Biotechnol (2013) 97:2671–2679 2677

Page 8: Kinetics during the redox biotransformation of pollutants mediated by immobilized and soluble humic acids

NSHAs as RMs during the reductive decolorization of RR2.On the other hand, NSHAi did not show a significant effect onthe rate of decolorization of RR2 by AGS (Cervantes et al.2011a). Since this HA sample was not sulfonated, it isexpected that it was adsorbed on the AER through groups,which might have a role in the redox process (e.g., quinones),whereas these redox functional groups might have been moreavailable in sludge incubations supplied with sulfonated HAi

because inserted sulfonate groups participated in the adsorp-tion process. FTIR spectra confirmed this theory since NSHAi

did not show the typical spectral signal associated with qui-none moieties (at ~1,650 cm−1), whereas it was present insulfonated HAi on the AER (Fig. 3).

Experiments conducted with the quinoid model com-pounds, AQDS and NQS, suggested that the standard redoxpotential (E°′) of the RM might partly determine the rate-limiting step during the reductive decolorization of RR2.Particularly, the reduction of RR2 by AH2QDS and NH2QSdepended on the E°′ of the quinoid model compound, name-ly, the reduction of RR2 by AH2QDS (E°′0−184 mV(Straub et al. 2001)) occurred at a 5.6- and 13.1-fold higherrate than that observed with NH2QS (E°′0+75 mV (Fieserand Fieser 1935) when the quinoid model compounds weresupplied in soluble and immobilized form, respectively(Table 2). These results are in agreement with Rau et al.(2002), who demonstrated that the lower the E°′ of qui-nones used to reductively decolorize Acid Red 27, thehigher decolorization rate was observed. Further experi-ments revealed that the reductive dechlorination of CT byAH2QDS and NH2QS followed the same dependency onthe E°′. Certainly, the reductive dechlorination of CT toCF by AH2QDS and NH2QS occurred at 11.9 and 3.5μEq/L h, respectively (data not shown).

Regarding the experiments to elucidate the kinetic con-straints during the mediated dechlorination of CT, resultsindicated that the reduction of CT to CF by both reducedHAs and HAi (second step in Fig. 1) was the rate-limitingstep during the redox process. Indeed, the microbial reduc-tion of HAs and HAi by HRS proceeded 1,400- and 790-foldfaster than the transfer of electron from reduced HAs andHAi, respectively, to achieve the reductive dechlorination ofCT to CF (Tables 1 and 2). Our results agree with thoseobtained by Curtis and Reinhard (1994), who observed anormalized KCT of 0.25×10−3 day−1 (mg TOC/L)−1 duringthe reduction of CT to CF by AH2QDS (the normalized KCT

values observed in the present study with reduced HAs andHAi were 0.17×10-3 and 0.12×10−3 day−-1 (mg TOC/L)−1,respectively). The dechlorination rate of CT observed withreduced HAs and HAi was even one order of magnitudelower than that observed during the reductive dechlorina-tion of hexachloroethane by electrochemically reduced HA(Kappler and Haderlein 2003). Therefore, strategies toovercome the kinetic limitations elucidated here should

be considered when applying HAi as RMs in continuousbioreactors for accelerating the redox biotransformation ofrecalcitrant pollutants, such as polyhalogenated solvents.Engineered HA with enhanced redox properties have beenproposed as potentially effective RMs for bioremediationpurposes (Perminova et al. 2005). It has also beenreported that mercaptoquinones are more effective RMsthan either semiquinones or hydroquinones for transferringelectrons to CT (Doong and Chiang 2005). Thus, there isa great potential for synthesizing HA with enhanced redoxproperties suitable for accelerating the redox biotransfor-mation of priority pollutants.

On the other hand, the increased dechlorination rate ofCT observed after 72 h of incubation, as compared to thatobserved during the initial incubation period (SD, Fig. S4),could be explained by the production of redox species with agreater reactivity with CT. The production of radicals duringthe reduction of both CT and HA has previously beendocumented (Scott et al. 1998; Doong and Chiang 2005).Thus, it is conceivable to assume that redox species with agreater reactivity with CT, which might have been producedduring the initial incubation period, would have promoted a10-fold higher KCT value. This observation agrees withKappler and Haderlein (2003), who observed two well-defined steps during the dechlorination of CT by electro-chemically reduced HA and suggested that this was due tothe presence of redox-active functional groups with differentreactivity to transfer electrons towards the reduction of CTto CF.

Overall, the present study provides elucidation on therate-limiting steps involved in the redox biotransformationof priority pollutants mediated by both HAs and HAi.Coimmobilization of HA and HRM in high-rate anaero-bic treatment systems is proposed as a mechanism toovercome the kinetic constraints encountered. Further-more, the development of engineered HA with enhancedredox properties is also proposed for accelerating theredox conversion of recalcitrant pollutants commonly foundin industrial wastewaters.

Acknowledgments We thank the technical assistance of Luis H.Alvarez, Dulce Partida, Guillermo Vidriales, and Ma. Carmen Rocha.CT and CF analyses were conducted at Laboratorio Nacional de Bio-tecnología Agrícola, Médica y Ambiental (LANBAMA, IPICYT). Thepresent study was financially supported by the Council of Science andTechnology of Mexico (SEP-CONACYT Grants 40808 and 155656).

References

Alvarez LH, Cervantes FJ (2011) Assessing the impact of aluminananoparticles in an anaerobic consortium: methanogenic and hu-mus reducing activity. Appl Microbiol Biotechnol. doi:10.1007/s00253-011-3759-4

2678 Appl Microbiol Biotechnol (2013) 97:2671–2679

Page 9: Kinetics during the redox biotransformation of pollutants mediated by immobilized and soluble humic acids

Alvarez LH, Jimenez-Bermudez L, Hernández-Montoya V, CervantesFJ (2011) Enhanced dechlorination of carbon tetrachloride byimmobilized fulvic acids on alumina particles. Water Air SoilPollut. doi:10.1007/s11270-011-0994-3

Bhushan B, Halasz A, Hawari J (2006) Effect of iron(III), humic acidsand anthraquinone-2,6-disulfonate on biodegradation of cyclicnitramines by Clostridium sp. EDB2. J Appl Microbiol 100(3):555–563

Bradley PM, Chapelle FH, Lovley DR (1998) Humic acids as electronacceptors for anaerobic microbial oxidation of vinyl chloride anddichloroethene. Appl Environ Microbiol 64:3102–3105

Cervantes FJ, Dijksma W, Duong-Dac T, Ivanova A, Lettinga G, FieldJA (2001) Anaerobic mineralization of toluene by enriched sedi-ments with quinones and humus as terminal electron acceptors.Appl Environ Microbiol 67:4471–4478

Cervantes FJ, Duong-Dac T, Roest K, Akkermans ADL, Lettinga G, FieldJA (2003) Enrichment and immobilization of quinone-respiring bac-teria in anaerobic granular sludge. Water Sci Technol 48(6):9–16

Cervantes FJ, García-Espinosa A, Moreno-Reynosa MA, Rangel-MéndezR (2010) Immobilized redox mediators on anion exchange resins andtheir role on the reductive decolorization of azo dyes. Environ SciTechnol 44(5):1747–1753

Cervantes FJ, Gonzalez-Estrella J, Marquez A, Alvarez LH, Arriaga S(2011a) Immobilized humic substances on an anion exchangeresin and their role on the redox biotransformation of contami-nants. Bioresour Technol 102(2):2097–2100

Cervantes FJ, Gutiérrez CH, López KY, Estrada-Alvarado MI, Meza-Escalante ER, Texier AC, Cuervo F, Gómez J (2008) Contributionof quinone-reducing microorganisms on the anaerobic biodegra-dation of organic compounds under different redox conditions.Biodegradation 19:235–246

Cervantes FJ, Mancilla AR, Ríos-del Toro EE, Alpuche-Solis AG,Montoya-Lorenzana L (2011b) Anaerobic benzene oxidation byenriched inocula with humic acids as terminal electron acceptors.J Hazard Mat 195(1):201–207

Cervantes FJ, van der Velde S, Lettinga G, Field JA (2000) Competitionbetween methanogenesis and quinone respiration for ecologicallyimportant substrates in anaerobic consortia. FEMS Microbiol Ecol34:161–171

Cervantes FJ, Vu-Thi-Thu L, Lettinga G, Field JA (2004) Quinone-respiration improves dechlorination of carbon tetrachloride byanaerobic sludge. Appl Microbiol Biotechnol 64(5):702–711

Collins R, Picardal F (1999) Enhanced anaerobic transformations ofcarbon tetrachloride by soil organic matter. Environ ToxicolChem 18(12):2703–2710

Curtis GP, Reinhard M (1994) Reductive dehalogenation of hexachloro-ethane, carbon tetrachloride, and bromoform by anthraquinonedisulfonate and humic acid. Environ Sci Technol 28(13):2393–2401

Doong R-A, Chiang H-C (2005) Transformation of carbon tetrachlo-ride by thiol reductants in the presence of quinone compounds.Environ Sci Technol 39(19):7460–7468

Dos Santos AB, Traverse J, Cervantes FJ, Van Lier JB (2005) Enhanc-ing the electron transfer capacity and subsequent color removal inbioreactors by applying thermophilic anaerobic treatment andredox mediators. Biotechnol Bioeng 89(1):42–52

Fieser LF, Fieser M (1935) The reduction potentials of various naph-thoquinones. J Am Chem Soc 57(3):491–494

Guo J, Zhou J, Wang D, Tian C, Wang P, Salah Uddin M, Yu H (2007)Biocalalyst effects of immobilized anthraquinone on the anaero-bic reduction of azo dyes by the salt-tolerant bacteria. Water Res41(2):426–432

Jiang J, Kappler A (2008) Kinetics of microbial and chemical reduc-tion of humic substances: implications for electron shuttling.Environ Sci Technol 42(10):3563–3569

Kappler A, Haderlein SB (2003) Natural organic matter as reductantfor chlorinated aliphatic pollutants. Environ Sci Technol 37(12):2714–2719

Kwon MJ, Finneran KT (2006) Microbially mediated biodegradationof hexahydro-1,3,5-trinitro-1,3,5-triazine by extracellular electronshuttling compounds. Appl Environ Microbiol 72(9):5933–5941

Lettinga G, Van Velsen AFM, Hobma SW, De Zeeuw W, Klapwijk A(1980) Use of upflow sludge blanket (USB) reactor concept forbiological wastewater treatment, especially for anaerobic treat-ment. Biotechnol Bioeng 22(4):699–734

Li L, Wang J, Zhou J, Yang F, Jin C, Qu Y, Li A, Zhang L (2008)Enhancement of nitroaromatic compounds anaerobic biotransfor-mation using a novel immobilized redox mediator prepared byelectropolymerization. Bioresour Technol 99(15):6908–6916

Liu G, Zhou J, Wang J, Wang X, Jin R, Lv H (2011) Decolorization ofazo dyes by Shewanella oneidensis MR-1 in the presence ofhumic acids. Appl Microbiol Biotechnol 91(2):417–424

Martínez CM, Cervantes FJ (2012) Simultaneous biodegradation ofphenol and carbon tetrachloride mediated by humic acids. Bio-degradation. doi:10.1007/s10532-012-9539-8

Pavlostathis SG, Beydilli MI (2005) Decolorization kinetics of the azodye reactive red 2 under methanogenic conditions: effect of long-term culture acclimation. Biodegradation 16(2):135–146

Penny C, Vuilleumier S, Bringel F (2010) Microbial degradation oftetrachloromethane: mechanisms and perspectives for bioremedi-ation. FEMS Microbiol Ecol 74(2):257–275

Perminova IV, Kovalenko AN, Schmitt-Kopplin P, Hatfield K, HertkornN, Belyaeva EY, Petrosyan VS (2005) Design of quinonoid-enriched humic materials with enhanced redox properties. EnvironSci Technol 39(21):8518–8524

Rau J, Knackmuss HJ, Stolz A (2002) Effects of different quinoidredox mediators on the anaerobic reduction of azo dyes by bac-teria. Environ Sci Technol 36(7):1497–1504

Scott DT, McKnight DM, Blunt Harris EL, Kolesar SE, Lovley DR(1998) Quinone moieties act as electron acceptors in the reductionof humic substances by humics-reducing microorganisms. Envi-ron Sci Technol 32(19):2984–2989

Silverstein MR, Webster FX (1992) Spectrometric identification oforganic compounds. Wiley, New York

Stevenson FJ (1994) Humus chemistry: genesis, composition, reac-tions. Wiley, New York

Straub KL, Benz M, Schink B (2001) Iron metabolism in anoxicenvironments at near neutral pH. FEMS Microbiol Ecol 34(3):181–186

Tartar HV, Garretson HH (1941) The thermodynamic ionization con-stants of sulfurous acid at 25°. J Am Chem Soc 63(3):808–816

Van der Zee FP, Cervantes FJ (2009) Impact and application of electronshuttles on the redox (bio)transformation of contaminants: a re-view. Biotechnol Adv 27(3):256–277

Wang J, Li L, Zhou J, Lu H, Liu G, Jin R, Yang F (2009) Enhancedbiodecolorization of azo dyes by electropolymerization-immobilizedredox mediator. J Hazard Mater 168(2–3):1098–1104

Yudov MV, Zhilin DM, Pankova AP, Rusanov AG, Perminova IV,Petrosyan VS, Matorin DN (2005) Synthesis, metal-binding prop-erties and detoxifying ability of sulphonated humic acids. In:Perminova IV, Hatfield K, Hertkorn N (eds) Use of humic sub-stances to remediate polluted environments: from theory to prac-tice. Springer, Dordrecht, pp 485–498

Appl Microbiol Biotechnol (2013) 97:2671–2679 2679