ENVIRONMENTAL BIOTECHNOLOGY

Evaluation of solid polymeric organic materialsfor use in bioreactive sediment capping to stimulatethe degradation of chlorinated aliphatic hydrocarbons

Siavash Atashgahi & Farai Maphosa & Jo De Vrieze &

Pieter Jan Haest & Nico Boon & Hauke Smidt &Dirk Springael & Winnie Dejonghe

Received: 2 June 2013 /Revised: 15 July 2013 /Accepted: 17 July 2013 /Published online: 17 August 2013# Springer-Verlag Berlin Heidelberg 2013

Abstract In situ bioreactive capping is a promising technolo-gy for mitigation of surface water contamination by dischargingpolluted groundwater. Organohalide respiration (OHR)of chlorinated ethenes in bioreactive caps can be stim-ulated through incorporation of solid polymeric organicmaterials (SPOMs) that provide a sustainable electronsource for organohalide respiring bacteria. In this study, woodchips, hay, straw, tree bark and shrimp waste, were assessedfor their long term applicability as an electron donor for OHRof cis-dichloroethene (cDCE) and vinyl chloride (VC) insediment microcosms. The initial release of fermentationproducts, such as acetate, propionate and butyrate led to theonset of extensive methane production especially in micro-cosms amended with shrimp waste, straw and hay, while noconsiderable stimulation of VC dechlorination was obtainedin any of the SPOM amended microcosms. However, in thelonger term, short chain fatty acids accumulation decreased aswell as methanogenesis, whereas high dechlorination rates of

VC and cDCE were established with concomitant increase ofDehalococcoides mccartyi and vcrA and bvcA gene numbersboth in the sediment and on the SPOMs. A numeric simula-tion indicated that a capping layer of 40 cm with hay, straw,tree bark or shrimp waste is suffice to reduce the groundwaterVC concentration below the threshold level of 5 μg/l beforedischarging into the Zenne River, Belgium. Of all SPOMs, thepersistent colonization of tree bark by D. mccartyi combinedwith the lowest stimulation of methanogenesis singled out treebark as a long-term electron donor for OHR of cDCE/VC inbioreactive caps.

Keywords Bioreactive sediment capping . Solid polymericorganic materials . Reductive dechlorination .

Dehalococcoides mccartyi

Introduction

Contamination with chlorinated aliphatic hydrocarbons(CAHs) poses serious threats to groundwater quality in indus-trialized countries (Abelson 1990). In case of limited attenu-ation in the aquifer and riverbed sediment, discharges of CAHcontaminated groundwater into surface water systems presentan additional environmental concern (Hamonts et al. 2009). Insitu capping is a promising technology for reducing the expo-sure of biota present in the surface water column to contam-inants present in or leaking from the sediments. Passive in situcapping is achieved by isolating the contaminated sedimentthrough coverage with clean material media, such as sandgravel and clean sediment (Go et al. 2009). Such barriers arebased on physical separation and are especially useful forcontainment of sediments contaminated with strongly sorbedhydrophobic pollutants. They are, however, not effective formitigating contamination of the surface water by discharging

Electronic supplementary material The online version of this article(doi:10.1007/s00253-013-5138-9) contains supplementary material,which is available to authorized users.

S. Atashgahi : P. J. Haest :W. Dejonghe (*)Flemish Institute for Technological Research (VITO), Separation andConversion Technology, Boeretang 200, 2400 Mol, Belgiume-mail: winnie.dejonghe@vito.be

S. Atashgahi : F. Maphosa :H. SmidtLaboratory of Microbiology, Wageningen University, 6703HB Wageningen, The Netherlands

S. Atashgahi :D. SpringaelDivision Soil and Water Management, KU Leuven, KasteelparkArenberg 20, 3001 Heverlee, Belgium

J. De Vrieze :N. BoonLaboratory of Microbial Ecology and Technology (LabMET), GhentUniversity, Coupure Links 653, 9000 Ghent, Belgium

Appl Microbiol Biotechnol (2014) 98:2255–2266DOI 10.1007/s00253-013-5138-9

groundwater plumes containing mobile persistent contami-nants (Himmelheber et al. 2007). An innovative alternativein situ capping technique that addresses particularly sitesaffected by mobile contaminants is active capping, a technol-ogy that aims at sequestering or transforming contaminantsthat are easily released from the sediment (Himmelheber et al.2011; Knox et al. 2008; Reible et al. 2006). The focus in thedevelopment of reactive caps has been on the application ofphysicochemical processes for contaminant removal, primar-ily by retarding the contaminants through sorption. The draw-back of this approach is that the risk to human and environ-mental health returns once the capping material is saturated(Sun et al. 2010). Therefore, recent developments in activesediment capping include the in situ bioreactive capping ap-proach in which the capping layer hosts and stimulatesmicrobiota that can degrade the pollutants (Himmelheberet al. 2011).

A major biological process that contributes to biodegrada-tion of CAHs is organohalide respiration (OHR). In thisprocess, specific bacterial groups such as Dehalococcoidesmccartyi use in most cases hydrogen as electron donor forrespiration on CAH, which results in their dechlorination.Well known examples are the stepwise dechlorination of thegroundwater pollutants tetrachloroethene (PCE) andtrichloroethene (TCE) to ethene through cis-dichloroethene(cDCE) and vinyl chloride (VC) (Smidt and de Vos 2004).OHR is stimulated by the addition of a substrate that isconverted by fermentation into a suitable electron donor.Stimulation of OHR of CAHs in bioreactive caps can as suchbe achieved by incorporation of suitable solid polymericorganic materials (SPOMs), which allows a long-term slowrelease supply of electron donors, includingH2. However, otherH2 consuming anaerobic organisms like sulfate, and ferric iron-reducing bacteria, homoacetogens and methanogens will com-pete with organohalide respiring bacteria (OHRB) for the pro-duced H2. In particular, methanogens are considered unwantedand main competitors of dehalogenating microbial guilds, suchas D. mccartyi due to their exclusive dependence on reducingequivalents such as H2 and acetate (Smatlak et al. 1996; Weiand Finneran 2013).

Although the application of SPOMs as a slow releaseelectron source for biostimulation of reductive dechlorinationhas been reported (Brennan et al. 2006; Kassenga et al. 2003;Vera et al. 2001;Wu et al. 1998), this has not been related withthe microbial ecology of such systems. For instance, it is notknown how the presence of SPOMs affects the microbialcommunity and especially OHRB, such as D. mccartyi, asthe main mediators of CAH degradation, and methanogens asthe main competitors of OHRB for reducing equivalents. More-over, information on whether these species are colonizing thecapping material and in particular the SPOMs, is lacking. In thisstudy, the long-term potential of five different solid organicmaterials as electron donors to stimulate reductive dechlorination

of cDCE and VC was evaluated. The SPOM materials werechosen from waste products with the dual aim of stimulation ofreductive dechlorination and waste recycling. We hypothesizedthat different SPOMs will result in different degradation patternsand stimulation of D. mccartyi and methanogens. SPOMs wereintroduced into sediment microcosms prepared from the Zenneriver-bed sediment that were impacted by dischargingVC/cDCEpolluted groundwater. The sediments have previously beenshown to contain VC/cDCE OHRB, including D. mccartyi(Hamonts et al. 2012; Atashgahi et al. 2013). The intermediateshort chain fatty acids (SCFAs) formed during the hydrolysis/fermentation of SPOMs were identified, and their impact on theabundance of D. mccartyi and the genes involved in CAHrespiration, archaea and methanogens in the sediment and onthe SPOMs, were quantified by quantitative PCR (qPCR). Fi-nally, the efficiency of the different SPOMs in stimulating VCdechlorination when installed at the Zenne site as a 40-cmcapping layer, was evaluated in silico using a 1D numericalmodel.

Materials and methods

Materials

River-bed sediment samples were taken from the Zenne Riverin Machelen-Vilvoorde, Belgium in May 2010 at a site thathas been previously described (Hamonts et al. 2009). At thatsite, a VC/cDCE contaminated groundwater plume of approx-imately 1.4 km width flows in the northwestern direction intothe Zenne River (see Fig. S1). The sediment samples wereobtained from the top 80 cm river-bed sediment at the post 25location. Groundwater was obtained from the SB2 groundwa-ter extraction well near the right river bank (Fig. S1) asoutlined by Hamonts et al. (2012). The SPOMs used werewood chips, hay, straw, tree bark, and shrimpwaste containing48 %, 42 %, 46 %, 40 %, and 38 % carbon, respectively, and0.07 %, 0.44 %, 0.09 %, 0.28 %, and 8.94 % nitrogen,respectively.

Sediment microcosm set-up

Microcosms were prepared in 160-ml glass serum bottlescontaining 20 g wet and well-mixed sediment obtained fromthe top 80-cm depth of Zenne riverbed, 70 ml of groundwaterand a particular SPOM. All the preparations were performedin an anaerobic glove box (Don Whitley Scientific Ltd, WestYorkshire, UK) under N2 atmosphere. The amounts of SPOMadded to the microcosms were different for each SPOM, butwere chosen as to obtain an equal amount of initial dissolvedorganic carbon (DOC) in all set-ups. To determine this, in apreliminary experiment, 1 g of each SPOM, cut into pieces of4–5 mm, was incubated at room temperature on a shaker

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(50 rpm) in 100ml of groundwater in 500-ml glass vials. After1 and 8 h of incubation, samples were taken and analyzed forthe DOC concentration. Afterwards, based on the DOC re-lease from wood chips that was used as the reference, theamount of each SPOM was chosen. The SPOMs were placedinto non-sorptive and permeable membrane pockets madefrom polyamide membrane 49 PA 6/5 (Hendrickx et al.2005) and introduced into the corresponding sediment micro-cosms in duplicate. All bottles were sealed with Teflon-linedbutyl rubber stoppers followed by addition of 70 μl VC (equalto 5.5 μmol/bottle) and incubated at room temperature. Timecourse liquid samples were withdrawn and filtered over a0.45-μm filter for DOC/SCFA measurements. After degrada-tion of the first VC spike, microcosms were flushed with N2

followed by a second VC spike at identical initial concentra-tion. After three consecutive additions of VC, cDCE (3 μmol/bottle) was added to microcosms and its degradation wasfollowed for another three consecutive cycles, unless other-wise stated. To study longevity of SPOM effects, the polyam-ide bags were replaced with new bags containing freshSPOMs, and the microcosms were used for long-term cDCEdegradation. Each time when cDCE was degraded, the micro-cosms were flushed with N2 and cDCE was replenished.

Sampling for molecular analysis and DNA extraction

Before the initiation of the experiment, approximately 0.5 g ofeach SPOM was used for DNA extraction in duplicate. Afterdechlorination of the third VC spike, the sediment micro-cosms were decapped in the anaerobic glove box. The poly-amide bags containing the SPOMs were opened and fewpieces of SPOM were withdrawn from both replicates. TheSPOMs were washed with sterile water to remove sedimentparticles and used for DNA extraction. Afterwards, the bottleswere capped again, mixed thoroughly, opened and approxi-mately 2g slurry samples were withdrawn from all duplicatemicrocosms. The same procedure was followed at the end ofthe third cDCE spike and at the end of the long-term cDCEdechlorination cycle. DNAwas extracted from the samples asdescribed previously (Hendrickx et al. 2005).

Quantitative PCR

Real-time qPCR targeting bacteria,D. mccartyi, and reductivedehalogenase genes vcrA and bvcA was performed as de-scribed by Atashgahi et al. (2013). qPCR targeting archaeaand the methanogenic families Methanosarcinaceae andMethanosaetaceae was performed using primers previouslydescribed by Yu et al. (2005) on a StepOnePlus™ Real-TimePCR System (Applied Biosystems, Carlsbad, CA, USA). Allprimer sets used in the different qPCR assays are given inTable S1. The reaction mixture of 20 μl consisted of 10 μl ofGoTaq® qPCR Master Mix (Promega, Madison, WI, USA),

3.5 μl of nuclease-free water and 0.75 μl of each primer (finalconcentration of 375 nM) and 5 μl of template DNA. TheqPCR program consisted of a predenaturation step of 10 minat 94 °C, followed by 40 cycles of 10 s at 94 °C and simul-taneous annealing and extension for 1 min at 60 °C. All DNAsamples were analyzed in triplicate. The ratio of D. mccartyias compared to total bacteria was estimated by dividing thenumber of D. mccartyi 16S rRNA gene copies by the copynumber of bacterial 16S rRNA genes multiplied by 100. Theratios ofD. mccartyi to archaea, archaea to bacteria, vcrA toD.mccartyi, and bvcA to D. mccartyi were determined in thesame way.

Model development

The efficiency of the different capping materials wheninstalled at the Zenne site, was evaluated in silico by meansof a 1D numerical model using PHREEQC (Parkhurst andAppelo 1999). The model consists of 75 cells with a cell sizeof 2 cm to describe 10 cm of aquifer, 1 m of sediment and acapping layer of 40 cm. The site has been monitored inten-sively regarding CAH concentrations and temperature valuesin the sediments, and those results were used in the model(Hamonts et al. 2012; Ebrahim et al. 2013). The domain ischaracterized by a Darcy velocity of 2 cm/day and a smalldispersivity of 0.1 m. An average measured groundwatertemperature of 12 °Cwas used for the simulation. The influentcDCE and VC concentrations were set at 22 and 1,222 μg/l,respectively, which correspond to the largest concentrationmeasured at the bottom of the sediment during the entiremonitoring period. First-order degradation rates for cDCEand VC in the aquifer were obtained by calibration to ob-served concentrations. The degradation rates in sediment andin the capping materials were obtained by linear regression ofconcentration time series in batch experiments at 12 and20 °C, respectively. The observed degradation rates wererecalculated to 12 °C using the Arrhenius equation assumingthat the rate is halved with a temperature decrease of 10 °C(Table S2).

Analytical methods

Total mass of CAHs, methane, ethene and ethane were quan-tified by means of gas chromatography (GC) in a GC appara-tus equipped with a flame ionization detector (FID) as previ-ously described (Atashgahi et al. 2013). DOC was determinedfrom samples as the difference between total dissolved carbonand dissolved inorganic carbon, measured with a ShimadzuTOC-5000 analyzer with ASI-5000 auto-sampler. Total Ni-trogen was detected after all particulate carbon was removedby filtration (pore size 0.45 μm). Carbon and nitrogen contentof the SPOMs were determined using a Carlo Erba EA1110elemental analyzer, using acetanilide as a standard as reported

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by Nieuwenhuize et al. (1994). SCFAs (C2–C5) were ana-lyzed in ether extracts from aqueous solutions using a GCequipped with a FID as previously described (Calli et al.2008).

Results

AminimumDOC value of 60 mg/l was previously reported tobe required to support dechlorination of 23 μM PCE beyondDCE in microcosm studies (Lee et al. 1997). Therefore, theamount of DOC released from different SPOMs as a source ofcarbon and electrons after 1 h of incubation in ground waterwas assessed in this study (Table S3). Wood chips released50.4 mg/l of DOC, which was in the middle range of thereleased DOC from SPOMs and therefore was used as areference. To reach equal initial DOC values, 1, 0.173,0.725, 3.33, and 1.37 g of wood chips, hay, straw, tree bark,and shrimp waste, respectively, were added to microcosms.After the first VC spike, microcosms amended with tree barkand shrimp waste showed the shortest lag phase (around7 days) and high initial degradation rates (Fig. 1). Microcosmsamended with straw and wood chips showed the longest lagphase of 28 days after which high VC dechlorination rateswere established in both microcosms (Fig. 1a). Althoughdechlorination was also observed in microcosms that werenot amended with exogenous electron donor, addition ofSPOMs in general led to three to four times higher degrada-tion rates compared to incubations without SPOM (Table S4).After the second VC spike, dehalogenation started immedi-ately with pronounced rates of VC dechlorination, withhighest rates observed in microcosms amended with tree barkand shrimp waste (Table S4). Accordingly, complete VCremoval was obtained within 14 days. The produced ethenewas further reduced to ethane in microcosms containing hay,straw and tree bark (Fig. 1c). An identical trend was observedafter the third VC spike with the exception of microcosmsamended with wood chips, as those apparently lost theirstimulatory effect (Fig. 1a).

The stimulation of CAH dechlorination by SPOMsremained evident after VCwas replaced by cDCE as the targetCAH (Fig. 2a). The microcosms amended with hay, straw, andtree bark sustained high rates of cDCE dechlorination for threeconsecutive spikes (Table S4), with no accumulation of VC(Fig. 2b), and dechlorination proceeded to ethene (Fig. 2c) andeventually to ethane (Fig. 2d). In contrast to these micro-cosms, the microcosms amended with shrimp waste showedintermediate degradation rates in all cDCE spikes, whereaswoodchips amended microcosms did not show stimulated cDCEdegradation as compared with the unamended microcosms.

During hydrolysis and fermentation of organic matter re-leased from the different SPOMs, different profiles andamounts of SCFAs were produced (Fig. 3). Microcosms

amended with shrimp waste showed the highest accumulationof SCFAs, followed by straw and hay. Acetate showed thehighest concentration with a peak at day 14, followed bypropionate and butyrate. Propionate production was excep-tionally high with an increasing trend until day 60 in themicrocosms amended with straw reaching above 6 mM(Fig. 3b). Production of other SCFAs such as isobutyrate,valerate, isovalerate, caproate and isocaproate was negligible(data not shown). Overall, the SPOM amended microcosmsshowed maximum accumulation of SCFAs (and DOC;Fig. S2) during degradation of the first VC spike.

No substantial methane production was noticed in the first14 days of the experiment. However, between days 14 and 60,methanogenic activity increased, leading to 171, 152, and 112

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Fig. 1 Anaerobic dechlorination of VC (a) and accumulation of ethene(b) and ethane (c) in sediment microcosms containing different solidpolymeric organic materials (SPOMs) as a source of carbon and electrons.Data shown are average values obtained from duplicate microcosms. Thelines are disconnected at the end of each spike before the flushing of theheadspace and subsequent re-spiking. Error bars are not shown for clarity

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times higher methane production values in microcosmsamended with shrimp waste, straw and hay, respectively, as

compared with natural attenuation, while this ratio was only 28in microcosms amended with tree bark (Fig. 4). After the firstspike of VC was consumed at day 60, regardless of the SPOM,the concentration of SCFAs (Fig. 3) and DOC (Fig. S2) de-creased steadily until day 100 (end of the third VC spike),concomitant with a sharp decrease in methanogenesis (Fig. 4).Accordingly, after day 100, when VC was replaced with cDCEas target CAH, a dramatic decrease in methanogenesis wasnoticed in all microcosms (Fig. 4), while still high cDCEdechlorination rates were sustained in microcosms amended

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Fig. 2 Anaerobic dechlorination of cDCE (a) to VC (b), ethene (c) andethane (d) in sediment microcosms containing different solid polymericorganic materials (SPOMs) as a source of carbon and electrons. cDCEwas added at day 100 when degradation of VC was completed inmicrocosms. Data shown are average values obtained from duplicatemicrocosms. The lines are disconnected at the end of each spike beforethe flushing of the headspace and subsequent re-spiking. Error bars arenot shown for clarity

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Fig. 3 Production of acetate (a), propionate (b) and butyrate (c) insediment microcosms containing different solid polymeric organic mate-rials (SPOMs) during VC and cDCE dechlorination cycles. Data shownare average values obtained from duplicate microcosms. Note that differ-ent concentration scales are used for the Y-axes of panels a, b and c. Thelines are disconnected at the end of each spike before the flushing of theheadspace and subsequent re-spiking. Error bars are not shown for clarity

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with straw, hay and tree bark (Table S4). It is likely that theinitial fast release of easily degradable materials led to extensive

methanogenesis, while long-term incubation appears to confinemethane production without affecting dechlorination rate. Thishypothesis was further tested by a second round of SPOMaddition into the microcosms and by following cDCE dechlo-rination cycles for a period of 1 year. During that period, 23–72 % of the total methane produced in all of the SPOMamended microcosms was formed during the first cDCE spike(Table 1) and was concomitant with a high SCFA/DOC pro-duction (Fig. S3). While methane production decreased overtime, cDCE degradation rates were sustained in microcosmsamended with the lignocellulosic materials hay, straw, and treebark. This indicates that these SPOMs do not become morerecalcitrant over time and that their slow degradation maintainssufficiently reducing and nutrient-rich conditions to promotereductive dechlorination (Table 1). In contrast, extensive degra-dation of shrimp waste appeared to cause depletion of fermen-tation intermediates in a shorter time period leading to dimin-ished cDCE degradation compared with the lignocellulosic

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Table 1 First order cDCE degradation rates and proportion of methane production in microcosms during long-term cDCE degradation for 1 year

Spike number Natural attenuation Wood chips Hay Straw Tree bark Shrimp waste

ka (day−1) CH4b (%) k (day−1) CH4 (%) k (day−1) CH4 (%) k (day−1) CH4 (%) k (day−1) CH4 (%) k (day−1) CH4 (%)

1 0.08 33.8 0.20 63.4 0.24 68.1 0.41 26.3 0.31 23.32 0.24 72.9

2 0.09 27.3 0.15 19.1 0.76 8.94 0.75 8.01 0.76 5.82 0.74 13.2

3 0.08 22.3 0.04 10.1 1.62 5.34 1.63 7.44 1.16 6.27 0.36 5.65

4 0.08 16.5 0.02 7.31 1.13 5.31 2.03 5.06 1.14 10.94 0.34 3.51

5 –c – – – 0.58 3.04 1.64 7.90 0.59 5.28 0.58 1.34

6 – – – – 0.59 3.06 1.17 5.05 0.58 5.73 0.58 1.35

7 – – – – 0.58 1.37 1.16 3.98 0.58 3.42 0.15 0.87

8 – – – – 0.59 1.70 1.16 2.84 0.59 6.22 0.08 0.78

9 – – – – 0.58 1.26 1.17 2.83 0.57 5.46 0.11 0.04

10 – – – – 0.33 0.30 1.17 2.84 0.42 1.31 0.06 0.18

11 – – – – 0.34 0.46 1.16 2.93 0.58 0.91 0.03 0.06

12 – – – – 0.49 0.64 1.59 3.04 0.49 9.99 – –

13 – – – – 0.58 0.42 1.16 2.17 0.58 3.23 – –

14 – – – – 0.58 0.05 1.16 2.16 0.59 1.14 – –

15 – – – – 0.59 0.03 1.15 2.21 0.58 1.81 – –

16 – – – – 0.58 0.03 1.16 1.13 0.58 1.49 – –

17 – – – – 0.37 0.02 0.82 2.37 0.59 4.32 – –

18 – – – – 0.33 0.01 1.18 1.9 0.57 3.35 – –

19 – – – – – – 1.16 2.24 – – – –

20 – – – – – – 1.16 2.07 – – – –

21 – – – – – – 0.58 2.01 – – – –

22 – – – – – – 0.57 1.94 – – – –

23 – – – – – – 0.57 1.64 – – – –

New SPOMs were amended to the corresponding microcosms at the beginning of the experimenta Degradation rates obtained by linear regression of concentration time series in batch experimentsbMethane proportion for each condition obtained by dividing the amount of methane produced at the end of each spike to total amount of methaneproduced during long-term repeated batchesc No cDCE spike was performed

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materials. Straw-amended microcosms degraded the highestmass of cDCE with the highest degradation rates. However,the ratio between the amount of methane and ethene + ethaneproduced was 101.73 μmol/μmol in these microcosms, whilethis ratio was only 8.01 in microcosms amended with tree bark(Table S5).

Dynamics of biomarkers

qPCR was performed to quantify the number of 16S rRNAgene copies of total bacteria, archaea, Methanosarcinaceae,Methanosaetaceae and D. mccartyi, and functional genesassociated with reductive dechlorination of VC (vcrA andbvcA), to assess the dynamics in the population sizes ofmicrobial guilds involved in OHR and methanogenesis inthe sediment and associated with the SPOM materials.

At the end of the VC dechlorination cycles, the 16S rRNAgene copy numbers of bacteria (Fig. S4) and archaea (Fig. 5)increased in all microcosms, and in particular on SPOM mate-rials, where their abundance increased generally more than twoorders of magnitude. Although growth of D. mccartyi was also

noticed in the non-amended microcosms, addition of lignocel-lulosic material led to a higher relative abundance of D.mccartyi 16S rRNA gene copies (D. mccartyi/bacteria), espe-cially in the microcosms amended with tree bark where thisratio was 21-fold higher compared with non-amended micro-cosms (Fig. 6a, Table 2, end of the third VC spike). An identicaltrend was noted for the relative abundances of the VC reductivedehalogenase-encoding genes vcrA and bvcA (Figs. S5 and S6)and their ratio over D. mccartyi (Table 2, end of the third VCspike), which was the highest in both the sediment slurry andon SPOM samples withdrawn from the microcosms amendedwith tree bark.

At the end of the third VC spike, the ratio of archaea/bacteria was lowest in both sediment slurry and on SPOMsamples in microcosms amended with tree bark (Table 2, endof the third VC spike), which was in line with the lowestmethane production in these microcosms (Fig. 4). In contrast,the ratio of archaea/bacteria was the highest in the sedimentslurry samples from microcosms amended with shrimp waste(Table 2, end of the third VC spike), which is in agreementwith the highest methane production in these microcosms(Fig. 4). The apparent extensive growth of archaea in the

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Fig. 5 16S rRNA gene copy numbers of total archaea on day 0 and endof VC, cDCE, and long-term cDCE degradation in sediment slurry (a)and solid polymeric organic materials (SPOMs) (b) as determined by real-time PCR. Each value represents the average value obtained from tripli-cate qPCRs performed on one sample of each of the duplicate micro-cosms (n=6). NA natural attenuation, WC wood chips, TB tree bark, SWshrimp waste

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a

b

Fig. 6 16S rRNA gene copy numbers of Dehalococcoides mccartyi onday 0 and at the end of VC, cDCE, and long-term cDCE degradation insediment slurry (a) and solid polymeric organic materials (SPOMs) (b) asdetermined by real-time PCR. NoD.mccartyiwas detected in T0 samplesof SPOMs. Each value represents the average value obtained fromtriplicate qPCRs performed on one sample of each of the dupli-cate microcosms (n=6). NA natural attenuation,WCwood chips, TB treebark, SW shrimp waste

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Table 2 The ratio of biomarkersin the sediment slurry and SPOMssamples

Values in boldface are minimumvalues; values in italics are mid-point values; values in romandenote maximum valuesa The ratio was estimated by di-viding the number of D. mccartyi16S rRNA gene copies by thecopy number of bacterial and ar-chaeal 16S rRNA genes multi-plied by 100b Time zeroc End of third VC spiked End of third cDCE spikee End of long-term cDCE spikes

Sample type Target biomarkers Conditions Ratio (%)a

T0b VCc cDCEd Lt-cDCEe

Slurry D. mccartyi/bacteria Natural attenuation 0.01 0.07 0.09 0.13

Wood chips 0.01 0.26 1.20 0.59

Hay 0.01 0.29 1.23 1.13

Straw 0.01 0.20 0.53 2.24

Tree bark 0.01 1.53 0.97 0.09

Shrimp waste 0.01 0.01 1.77 0.02

D. mccartyi/archaea Natural attenuation 0.21 19.86 147.20 28.08

Wood chips 0.21 1.53 10.06 0.59

Hay 0.21 67.39 80.10 5.55

Straw 0.21 15.71 2.07 38.32

Tree bark 0.21 2,748.92 34.79 103.99

Shrimp waste 0.21 0.03 0.64 0.05

Aechaea/bacteria Natural attenuation 3.45 0.36 0.06 0.46

Wood chips 3.45 16.71 11.97 100.60

Hay 3.45 0.44 1.54 20.41

Straw 3.45 1.25 25.72 5.85

Tree bark 3.45 0.06 2.79 0.09

Shrimp waste 3.45 21.64 276.33 28.25

vcrA/D. mccartyi Natural attenuation 0.00 81.97 41.92 176.02

Wood chips 0.00 19.54 7.77 6.63

Hay 0.00 2.85 3.98 24.70

Straw 0.00 1.47 164.62 3.07

Tree bark 0.00 11.21 47.86 4.61

Shrimp waste 0.00 19.79 41.65 47.93

bvcA/D. mccartyi Natural attenuation 0.00 213.66 190.18 190.74

Wood chips 0.00 66.61 62.17 9.12

Hay 0.00 15.18 19.19 28.59

Straw 0.00 3.54 5,447.27 7.75

Tree bark 0.00 204.47 198.83 15.06

Shrimp waste 0.00 53.02 51.56 69.78

SPOMs D. mccartyi/bacteria Wood chips 0.00 0.16 0.17 0.04

Hay 0.00 2.45 1.10 0.42

Straw 0.00 2.22 4.07 4.01

Tree bark 0.00 0.27 2.19 1.78

D. mccartyi/archaea Wood chips 0.00 0.01 0.00 0.00

Hay 0.00 0.14 0.17 0.06

Straw 0.00 0.38 0.02 0.24

Tree bark 0.00 0.35 0.23 0.35

Aechaea/bacteria Wood chips 335.64 16.70 178.43 150.52

Hay 17.45 17.04 6.55 7.24

Straw 14.49 5.85 170.98 16.60

Tree bark 9.44 0.77 9.66 5.00

vcrA/D. mccartyi Wood chips 0.00 34.56 23.42 77.43

Hay 0.00 7.99 8.70 166.03

Straw 0.00 2.43 14.93 60.38

Tree bark 0.00 41.72 101.33 1,389.73

bvcA/D. mccartyi Wood chips 0.00 19.25 47.25 37.11

Hay 0.00 33.89 27.13 204.90

Straw 0.00 6.20 118.33 29.93

Tree bark 0.00 141.22 228.73 1,081.19

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microcosms amended with shrimp waste (Fig. 5a) resultedalso into the lowest ratio of D. mccartyi/archaea (Table 2, endof the third VC spike). However, due to the degradation ofshrimp waste, no SPOM could be sampled to study the local-ization of the target biomarkers on shrimp waste. Among thelignocellulosic materials, microcosms amended with woodchips had the lowest ratio of D. mccartyi/archaea (Table 2,end of the third VC spike), which is in line with thelowest degradation rates observed in these microcosms(Table S4).

At the end of the cDCE degradation cycles at day 200,among the amended microcosms, the D. mccartyi/archaearatio was the highest in sediment slurry and on SPOM samplesin the microcosms amended with hay and tree bark, althoughit should be noted that this ratio was the highest in sedimentslurries of unamended microcosms (Table 2, end of the thirdcDCE spike). The abundance of archaea (Fig. 5a), Methano-sarcinaceae (Fig. S7a) and Methanosaetaceae 16S rRNA genecopies (Fig. S8a) decreased more than one order of magnitudeonly in the sediment slurry of the microcosms amended withshrimp waste. Nevertheless, due to the concomitant decreaseof bacterial 16S rRNA gene copies (more than 2 orders ofmagnitude), the ratio of archaea/bacteria was still the highestin the sediment slurry of the microcosms amended withshrimp waste (Table 2, end of the third cDCE spike). Theratio of archaea/bacteria increased more than 20-fold in bothsediment slurry and on SPOM samples withdrawn from mi-crocosms amended with straw (Table 2, end of the third cDCE

spike), which was due to further proliferation of archaea inthese microcosms (Fig. 5), which relates to the continuousproduction of methane by the end of the third cDCE spike(Fig. 4).

At the end of the long-term cDCE degradation experimen-tal period, the microcosms amended with tree bark had thehighest ratio of D. mccartyi/archaea and the lowest ratio ofarchaea/bacteria in both sediment slurry and on SPOM sam-ples (Table 2, end of the long-term cDCE spikes). Microcosmsamended with wood chips on the other hand showed thelowest ratio of D. mccartyi/archaea and the highest ratio ofarchaea/bacteria among the lignocellulosic materials.

Prediction of reactive transport of CAH in in situ caps

The degradation rates of cDCE and VC that were obtained inthis study compare well to the range of values reported in theliterature with the exception of cDCE degradation using chitinas reported by Brennan et al. (2006) (Table S2). A simulationof the reactive transport in the hyporheic zone with thepresented constants showed that VC concentrations abovethe threshold of 5 μg/l would discharge in the surface waterif no additional measures are being taken (Fig. 7). In the caseof a pore water velocity of 5.4 cm/day, a capping layer of40 cm using hay, straw, tree bark or shrimp waste as SPOMswould suffice to bring the VC concentration below the thresh-old level.

Fig. 7 Numeric modellinganalysis of VC reactive transportin different capping materialsapplied at the Zenne site,Belgium. The graphs show theconcentrations of VC (red) andethene (green). The top graphshows the modelled 1D domainof 10 cm aquifer (brown), 1 m ofsediment (gray), 40 cm ofcapping material (white) and withthe river in blue. The bottomgraphs show the concentration inthe capping material at a moredetailed scale, with the legalthreshold for VC in groundwaterindicated by the dashed line.Under natural attenuation,application of a sand cap wasconsidered. A distance of 0 m(x-axis) was defined as a depth of1.1 m from the sediment–surfacewater interface without cap

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Discussion

We examined whether SPOMs can be incorporated into bio-reactive sediment caps to stimulate bioremediation of CAH-contaminated sediments by means of OHR, and whether thechanneling of reducing equivalents to methanogenesis can beof technical and environmental concern. This was achieved bystudying the effect of the presence of SPOMs on reductivedechlorination of CAHs and the dynamics of OHRB andmethanogenic archaea in sediments that have been previouslyexposed to CAH contamination.

During the first 2 weeks, extensive accumulation of SCFAs(Fig. 3) and DOC (Fig. S2) was noticed in SPOM containingmicrocosms, which could be due to direct release and fermen-tation of sugars, amino acids, and proteins (Robinson-Loraand Brennan 2009). The observed accumulation of SCFAswas likely due to the initial low numbers of SCFA consumingmicroorganisms, such as Methanosarcina (Fig. S7) as themain acetoclastic methanogens (De Vrieze et al. 2012). Thisis in line with the fact that no accumulation of methane as theproduct of methanogenesis was noticed in this phase. Anotherreason for such a delay in onset of methanogenesis might bethe dominance of iron and sulfate reduction which typicallytakes place during the early stages of anoxic incubations andhas been reported in anoxic paddy soil amended with ricestraw (Glissmann and Conrad 2000). Although no accumula-tion of ethene as the product of reductive dechlorination wasnoticed during the first 2 weeks (Fig. 1b), a slight decrease inVC concentration was evident in SPOM amended microcosmswhich might be due to anaerobic VC oxidation coupled to ironreduction (Bradley and Chapelle 1996). The onset of extensivemethane production at day 14 in microcosms amended withshrimp waste, straw and hay (Fig. 4) was concomitant with adecrease in acetate and butyrate concentrations (Fig. 3). Incontrast, no considerable stimulation of VC dechlorinationwas obtained in SPOM amended microcosms during this peri-od of high methane production, suggesting dominance ofmethanogenesis in the early stage of the experiment. The clearstimulation of dechlorination in SPOM amended microcosmsafter the second and third VC spikes was coupled to a substan-tial enrichment ofD. mccartyi at day 100. Although the ratio ofD. mccartyi/archaea in the slurry samples of the unamendedmicrocosmswas higher thanmicrocosms amendedwith shrimpwaste, wood chips and even straw (Table 2), the obtaineddegradation rates in the unamended microcosms were notsufficient to reduce the VC concentration below the thresholdlevel (Fig. 7). The fact thatD. mccartyiwas not detected on theSPOM materials before incubation, and the extensive coloni-zation of SPOM samples by D. mccartyi in the course of theexperiment, suggest that the SPOMs can host sediment-bornD.mccartyi populations capable of CAH biotransformation.

The sustained high degradation rates during subsequentcDCE addition (Fig. 2) with drastic decrease ofmethanogenesis

(Fig. 4) after day 100 showed that long-term incubation mightlead to the dominance of OHRB instead of methanogens.Accordingly, during the long-term cDCE dechlorination exper-imental period with new addition of SPOMs, confinedmethanogenesis and sustained dechlorination rates over timeespecially in tree bark amended microcosms were recorded(Table 1), verifying that long-term incubation favors dechlo-rinating communities over methanogens. It can be speculatedthat after the depletion of immediately available easily degrad-able molecules, such as organic acids and alcohols, with con-comitant predominance of methanogens as evidenced by thefact that most methane is produced in the initial phase afterSPOM addition (Table 1), anaerobic degradation of more com-plex materials, such as cellulose by cellulolytic and fermenta-tive microbes, provide a long-term slow release source ofacetate and H2, which is sufficient to sustain reductive dechlo-rination. The success of reductive dechlorination in this phasecould be due to the competitive advantage of D. mccartyi overmethanogens in conditions of low H2 fluxes, as based on H2

half-saturation rate constants, D. mccartyi strains are excellentscavengers of H2 under H2-limiting conditions (Fennell andGossett 1998). Hydrogen concentrations were not measured inour experiments, but Brennan (2003) showed that the majorityof the total electron donating capacity released during chitinfermentation was acetate while hydrogen represented less than0.01 %. Accordingly, other reports showed that the majorfraction of electron equivalents during fermentation is routedto organic products and not to H2 (Aulenta et al. 2007; Fennellet al. 1997), even though the measured concentrations of hy-drogen do not necessarily reflect hydrogen consumption rates(Löffler et al. 1999).

Although straw-amended microcosms showed the highestcDCE degradation rates (Table 1), the ratio between theamount of methane and ethene + ethane produced was 12.7-fold higher in these microcosms than in microcosms amendedwith tree bark (Table S5), while the D. mccartyi populationsize (Fig. 6) and rdh gene numbers (Fig. S5 and S6) wereidentical in both microcosms, and the ratio of D. mccartyi/archaea was higher in microcosms amended with tree barkboth in the sediment slurry and on the SPOM (Table 2). Thisshows that the major fraction of reducing equivalents from thestraw was routed to the production of organic intermediatesexceeding the requirement for reductive dechlorination, whichcan lead to stimulation of methanogenesis and an increase ofbiochemical oxygen demand in the water. Recently, Wei andFinneran (2013) reported that adding ten times more acetatedid not increase the rate or extent of TCE reduction, butonly increased methane production. Taking into account thatother main SCFAs, such as propionate and butyrate areexpected to be converted to acetate, CO2, formate and hydro-gen by syntrophic microbial communities (Stams and Plugge2009), the excess of electron donors will be consumed formethane production, which is unwanted for several reasons.

2264 Appl Microbiol Biotechnol (2014) 98:2255–2266

The impact of methanogenesis is not merely restricted to theconsumption of reducing equivalents and production of thegreenhouse gas methane. Extensive methanogenesis in situcan also strip volatile CAHs by the formed methane bubbles(Vroblesky et al. 1989; Viana et al. 2007), compromise thephysical stability of the bioreactive cap and as such provideadditional pathways for contaminant release (Reible et al.2006). Therefore, control of methanogenesis is important toguarantee effectiveness of SPOM-based bioreactive caps.

Taking into account that lignocellulosic materials are agri-cultural by-products/wastes with little or no economic value,their use into bioreactive caps is interesting for both pollutantdegradation and waste load reduction. However, it should benoted that due to differences between SPOMs, site character-istics, and contaminants of concern, the selection of an appro-priate SPOM can have critical impact on the success ofbioreactive sediment capping technology. As shown here,addition of wood chips did not translate to a better dechlori-nation, and most of the reducing equivalents produced fromaddition of straw and hay were channeled to methanogenesis.Our results showed that tree bark is the best SPOM to stimu-late CAH degradation in the sediments of the Zenne site.Rapid and persistent colonization of the tree bark by a D.mccartyi community shows that a SPOM-based layer canpotentially form a sustainable biotransformation layer in abioreactive cap in which the release of electron donor andpresence of CAHs is coupled with the dehalogenating micro-organisms in the same place. The fact that D. mccartyi nativeto the sediment colonized the SPOM suggests thatD.mccartyieluted from deeper sediment by the upwelling ground watercan colonize the bioreactive layer in an in situ application.Previously, localization of microorganisms indigenous toaquatic sediments on the overlaying sand cap material wasshown in up-flow column experiments (Himmelheber et al.2009). In conclusion, the results of this study showed thatalthough lignocellulosic materials are promising candidates tobe incorporated into bioreactive caps, before embarking intofull-scale field bioreactive capping efforts, bench-scale screen-ing of SPOMs is necessary to stimulate the target microbialgroups without generating new environmental problems.

Acknowledgements This study was supported by the EU projectAQUAREHAB (grant 226565) and a VITO/KU Leuven PhD grant toS. Atashgahi. Additional support to HS and FMwas obtained through theEcoLinc project of the Netherlands Genomics Initiative. We thank B.Horemans and S. Bouillon for assistance in analysis of carbon and nitrogencontent.

References

Abelson PH (1990) Volatile contaminants of drinking water. Science247:141

Atashgahi S, Maphosa F, Doğan E, Smidt H, Springael D, Dejonghe W(2013) Small-scale oxygen distribution determines the vinyl chlo-ride biodegradation pathway in surficial sediments of river-bedhyporheic zones. FEMS Mic Ecol 84:133–142

Aulenta F, Pera A, Rossetti S, Papini M, Majone M (2007) Relevance ofside reactions in anaerobic reductive dechlorination microcosmsamended with different electron donors. Water Res 41:27–38

Bradley PM, Chapelle FH (1996) Anarobic mineralization of vinyl chlo-ride in Fe(III) reducing aquifer sediments. Environ Sci Technol30:2084–2086

Brennan RA (2003) PhD dissertation. University of Illinois at Urbana–Champaign, US

Brennan RA, Sanford RA, Werth CJ (2006) Chitin and corncobs aselectron donor sources for the reductive dechlorination oftetrachloroethene. Water Res 40:2125–2134

Calli B, Schoenmaekers K, Vanbroekhoven K, Diels L (2008) Darkfermentative H2 production from xylose and lactose—effects ofonline pH control. Int J Hydrogen Energy 33:522–530

DeVrieze J, Hennebel T, BoonN, VerstraeteW (2012)Methanosarcina: therediscovered methanogen for heavy duty biomethanation. BioresourTechnol 112:1–9

Ebrahim GY, Hamonts K, van Griensven A, Jonoski A, Dejonghe W,Mynett A (2013) Effect of temporal resolution of water level andtemperature inputs on numerical simulation of groundwater–surfacewater flux exchange in a heavily modified urban river. HydrolProcess 27:1634–1645

Fennell DE, Gossett JM (1998) Modeling the production of and compe-tition for hydrogen in a dechlorinating culture. Environ Sci Technol32:2450–2460

Fennell DE, Gossett JM, Zinder SH (1997) Comparison of butyric acid,ethanol, lactic acid, and propionic acid as hydrogen donors for thereductive dechlorination of tetrachloroethene. Environ Sci Technol31:918–926

Glissmann K, Conrad R (2000) Fermentation pattern of methanogenicdegradation of rice straw in anoxic paddy soil. FEMS MicrobiolEcol 31:117–126

Go JL, Lampert DJ, Stegemann JA, Reible DD (2009) Predicting con-taminant fate and transport in sediment caps: mathematical model-ing approaches. Appl Geochem 24:1347–1353

Hamonts K, Hamonts K, Maesen M, Bronders J, Lookman R, Kalka H,Diels L,Meckenstock RU, Springael D, DejongheW (2009) Factorsdetermining the attenuation of chlorinated aliphatic hydrocarbons ineutrophic river sediment impacted by discharging polluted ground-water. Environ Sci Technol 43:5270–5275

Hamonts K, Kuhn T, Vos J, Maesen M, Kalka H, Smidt H, Springael D,Meckenstock RU, Dejonghe W (2012) Temporal variations in nat-ural attenuation of chlorinated aliphatic hydrocarbons in eutrophicriver sediments impacted by a contaminated groundwater plume.Water Res 46:1873–1888

Hendrickx B, Dejonghe W, Boenne W, Brennerova M, Cernik M,Lederer T, Bucheli- Witschel M, Bastiaens L, Verstraete W, TopEM, Diels L, Springael D (2005) Dynamics of an oligotrophicbacterial aquifer community during contact with a groundwater plumecontaminated with benzene toluene, ethylbenzene, and xylenes: an insitu mesocosm study. Appl Environ Microbiol 71:3815–3825

Himmelheber DW, Pennell Kurt D, Hughes JB (2007) Natural attenuationprocesses during in situ capping. Environ Sci Technol 41:5306–5313

Himmelheber DW, Thomas SH, Löffler FE, Taillefert MT, Hughes JB(2009) Microbial colonization of an in situ sediment cap and corre-lation to stratified redox zones. Environ Sci Technol 43:66–74

Himmelheber DW, Pennell KD, Hughes JB (2011) Evaluation of alaboratory-scale bioreactive in situ sediment cap for the treatmentof organic contaminants. Water Res 45:5365–5374

Kassenga GR, Pardue JH, Blair S, Ferraro T (2003) Treatment of chlo-rinated volatile organic compounds in upflow wetland mesocosms.Ecol Eng 19:305–323

Appl Microbiol Biotechnol (2014) 98:2255–2266 2265

Knox AS, Paller MH, Reible DD, Ma XM, Petrisor IG (2008) Seques-tering agents for active caps— remediation of metals and organics.Soil Sediment Contam 17:516–532

Lee MD, Quinton GE, Beeman RE, Biehle AA, Liddle RL, Ellis DE,Buchanan RJ Jr (1997) Scaleup issues for in situ anaerobic tetra-chloroethene bioremediation. J Ind Microbiol Biotechnol 18:106–115

Löffler FE, Tiedje JM, Sanford RA (1999) Fraction of electrons con-sumed in electron acceptor reduction and hydrogen thresholds asindicators of halorespiratory physiology. Appl Environ Microbiol65:4049–4056

Nieuwenhuize J, Maas YEM, Middleburg JJ (1994) Rapid analysis oforganic carbon and nitrogen in particulate materials. Mar Chem44:217–224

Parkhurst DL, Appelo CAJ (1999) User's guide to PHREEQC (Version 2)– a computer program for speciation, batch-reaction, one-dimensionaltransport and inverse geochemical calculations. Water-Resources In-vestigations, Denver, Report 99-4259, 312 pp

Reible DD, Lambert D, Constant DW, Mutch RD, Zhu Y (2006) Activecapping demonstration in the Anacostia River, Washington, DC.Remediation 17:39–53

Robinson-Lora MA, Brennan RA (2009) The use of crab-shell chitin forbiological denitrification: batch and column tests. Bioresour Technol100:534–541

Smatlak CR, Gossett JM, Zinder SH (1996) Comparative kinetics ofhydrogen utilization for reductive dechlorination of tetrachloroetheneand methanogenesis in an anaerobic enrichment culture. Environ SciTechnol 30:2850–2858

Smidt H, de Vos WM (2004) Anaerobic microbial dehalogenation. AnnuRev Microbiol 58:43–73

Stams AJM, Plugge CM (2009) Electron transfer in syntrophic commu-nities of anaerobic bacteria and archaea. Nat Rev Microbiol 7:568–577

Sun M, Yan F, Zhang R, Reible DD, Lowry GV, Gregory KB (2010)Redox control and hydrogen production in sediment caps usingcarbon cloth electrodes. Environ Sci Technol 44:8209–8215

Vera SM,Werth CJ, Sanford RA (2001) Evaluation of different polymericorganic materials for stimulating reductive dechlorination. Bioreme-diation J 5:169–181

Viana PZ, Yin K, Zhao X, Rockne KJ (2007)Active sediment capping forpollutant mixtures: control of biogenic gas production under highlyintermittent flows. Land Contam Reclam 15:413–425

Vroblesky DA, Lorah MW, Oliveros JP (1989) Ground-water, surface-water, and bottom-sediment contamination in the O-field area, Ab-erdeen Proving Ground, Maryland, and the possible effects ofselected remedial actions. Open-File report 1989-399. U.S. Geolog-ical Survey, Washington, DC

Wei N, Finneran KT (2013) Low and high acetate amendments areequally as effective at promoting complete dechlorination of trichlo-roethylene (TCE). Biodegradation 24:413–425

WuW, Nye J,Mahendra KJ, Kickey RF (1998) Anaerobic dechlorinationof trichloroethylene (TCE) to ethylene using complex organic ma-terials. Water Res 32:1445–1454

Yu Y, Lee C, Kim J, Hwang S (2005) Group-specific primer and probesets to detect methanogenic communities using quantitative real-time polymerase chain reaction. Biotechnol Bioeng 89:670–679

2266 Appl Microbiol Biotechnol (2014) 98:2255–2266