Chemosphere 97 (2014) 54–63

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Enhanced dechlorination of carbon tetrachloride by Geobacter sulfurreducens in the presence of naturally occurring quinones and ferrihydrite Ruey-an Doong ⇑, Chun-chi Lee, Chia-min Lien Department of Biomedical Engineering and Environmental Sciences, National Tsing Hua University, 101, Sec. 2, Kuang Fu Road, Hsinchu 30013, Taiwan

h i g h l i g h t s

g r a p h i c a l a b s t r a c t

 The bioreduction efficiency and rate

of Fe(OH)3 was enhanced by LQ and NQ.  The conversion efficiency of Fe(OH)3 was up to maximum of 73 ± 4% at 10 lM LQ.  Rate constant for CT dechlorination and production of Fe(II) was highly correlated.  NQ was reduced to semiquinone while only hydroquinone was produced in LQ system.

a r t i c l e

i n f o

Article history: Received 12 April 2013 Received in revised form 5 November 2013 Accepted 6 November 2013 Available online 28 November 2013 Keywords: Carbon tetrachloride Naturally occurring quinones Ferrihydrite Geobacter sulfurreducens Biogenic Fe(II)

a b s t r a c t The effect of naturally occurring quinones including lawsone (LQ), ubiquinone (UQ), juglone (JQ), and 1,4naphthoquinone (NQ) on the biotransformation of carbon tetrachloride (CT) in the presence of Geobacter sulfurreducens and ferrihydrite was investigated. AQDS was used as the model compound for comparison. The reductive dissolution of ferrihydrite by G. sulfurreducens was enhanced by AQDS, NQ, and LQ. However, addition of UQ and JQ had little enhancement effect on Fe(II) production. The bioreduction efficiency and rate of ferrihydrite was highly dependent on the natural property and concentration of quinone compounds and the addition of low concentrations of LQ and NQ significantly accelerated the biotransformation rate of CT. The pseudo-first-order rate constants for CT dechlorination (kobsCT) in AQDS-, LQ- and NQ-amended batches were 5.4–5.8, 4.6–7.4 and 2.4–5.8 times, respectively, higher than those in the absence of quinone. A good relationship between kobsCT for CT dechlorination and bioreduction ratio of ferrihydrite was observed, indicating the important role of biogenic Fe(II) in dechlorination of CT under iron-reducing conditions. Spectroscopic analysis showed that AQDS and NQ could be reduced to semiquinones and hydroquinones, while only hydroquinones were generated in LQ-amended batches. Ó 2013 Elsevier Ltd. All rights reserved.

1. Introduction Chlorinated hydrocarbons are one of the frequently encountered solvents widely used for different industrial and ⇑ Corresponding author. Tel.: +886 3 5726785; fax: +886 3 5718649. E-mail address: [email protected] (R.-a. Doong). 0045-6535/$ - see front matter Ó 2013 Elsevier Ltd. All rights reserved. http://dx.doi.org/10.1016/j.chemosphere.2013.11.004

domestic purposes (Doong et al., 2003; Penny et al., 2010; Muchitsch et al., 2011). For example, carbon tetrachloride (CT) is highly persistent in the environment compared with other chlorinated hydrocarbons and is still produced as an intermediate in the production of other chlorinated compounds (Penny et al., 2010). These compounds have become one of the most often found recalcitrant compounds in the environment, and are promulgated as the

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suspected carcinogens which could cause toxicity to human beings and ecosystem. Therefore, the understanding of fate and transport of these compounds in the natural environments and the development of effective technology for treatment of chlorinated hydrocarbons in contaminated sites are urgently necessary. Iron-reducing environment is one of the most important naturally occurring environments (Straub et al., 2001; Maithreepala and Doong, 2005; Minyard and Burgos, 2007; Lovley et al., 2011). It is believed that the chlorinated hydrocarbons can undergo both biotic and abiotic reductions under iron-reducing conditions (Maithreepala and Doong, 2008; Borch et al., 2010). Several studies have reported the importance of surface-bound iron species to the abiotically reductive transformation of chlorinated hydrocarbons (Maithreepala and Doong, 2005, 2009; Tao et al., 2010). Addition of ferrous ions to the various iron oxide minerals such as green rust, mackinawite, and goethite has been found to effectively reduce CT, tetrachloroethylene, and nitroaromatic compounds (Elsner et al., 2004; Tobler et al., 2007; Jeong and Hayes, 2007; Maithreepala and Doong, 2009; Bae and Lee, 2012; Amir and Lee, 2012). The surface associated with ferrous ions is a bio-geochemically important reaction and the decomposition of chlorinated hydrocarbons is controlled by the iron oxides species and surface density of Fe(II) (Elsner et al., 2004; Tobler et al., 2007; Buchholz et al., 2011; Larese-Casanova et al., 2012). Dissimilatory iron-reducing bacteria (DIRB) such as Geobacter and Shewanella are ubiquitous in soil and subsurface environments (Straub et al., 2001; Lovley et al., 2011). The microbial reduction of iron oxides by DIRB has recently been recognized as an important process for the generation of Fe(II) as well as the degradation of organic contaminants under anaerobic conditions (Doong and Schink, 2002; Maithreepala and Doong, 2008; Bose et al., 2009; Garcia-Balboa et al., 2010). When poorly and crystalline iron oxides are reductively dissolved to ferrous ions, the ferrous ions would adsorb onto the surface of ferric oxides to form reactive surface-bound iron species. Compared with the abiotic reactions, biological transformation is relatively slow but the biogenic ferrous ion concentration could increase with time, resulting in the increase in degradation efficiency and rate of chlorinated hydrocarbons. Several studies have demonstrated that humic acid and 9,10-anthraquinone-2,6-disulfonic acid (AQDS) can serve as the electron mediators to significantly enhance the dissolution efficiency and rate of iron oxides (Lovley et al., 1996; Maithreepala and Doong, 2009; Wu et al., 2013). Previous studies have mainly focused on the enhancement effect of AQDS on the dissolution of iron oxides and the dechlorination efficiency and rate of chlorinated hydrocarbons (Kwon and Finneran, 2006; Jiang et al., 2009; Aulenta et al., 2010; Cao et al., 2012). It is noteworthy that quinone compounds are ubiquitously present in trace amounts in natural organic matter and 50–60% fluorophores in the humic acid are quinone moieties (Cory and McKnight, 2005), while AQDS is an artificial analogue mainly produced from chemical processes and is hard to find in the natural environment. This gives a great impetus to understand the effect of various naturally occurring quinone compounds on the enhancement of Fe(II) production as well as the dechlorination of chlorinated hydrocarbons. In this study, the effect of four naturally occurring quinones including lawsone (LQ), 1,4-naphthoquinone (NQ), ubiquinone (UQ), and juglone (JQ) on the reductive dissolution of ferric oxides and the dechlorination of CT by Geobacter sulfurreducens were systematically evaluated under iron-reducing conditions. AQDS was employed as the electron mediator for comparison. The kinetics for biogenic Fe(II) formation and CT transformation in the presence of electron mediators were examined and compared. In addition, the relationship between the dechlorination rate of CT and produced ferrous concentrations was elucidated.

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2. Materials and methods 2.1. Chemicals Carbon tetrachloride (CT, CCl4, >99.8%, GC grade), chloroform (CHCl3, >99.8%, GC grade), and dichloromethane (CH2Cl2, >98%, GC grade) were purchased from Merck Co. (Darmstadt, Germany). Fumarate disodium salt (C4H2O4Na2, 99%), ferrozine monosodium salt (C20H13N4O6S2Na), lawsone (2-hydroxy-1,4-naphthoquinone, LQ), juglone (5-hydroxy-1,4-naphthoquinone, JQ), 9,10-anthraquinone-2,6-disulfonaic acid disodium salt (AQDS, >98%), ubiquinone (2,3-dimethoxy-5-methyl-p-benzoquinone, UQ), and 1,4-naphthoquinone (NQ, 97%) were purchased form Sigma–Aldrich Co. (Milwaukee, WI). All other chemicals were of analytical grade and were used as received without further purification. Solutions were prepared with deoxygenated deionized water using a vacuum and high-purity N2 purging system (Maithreepala and Doong, 2004). Ferrihydrite was synthesized according to the method previously described (Maithreepala and Doong, 2008, 2009), and was characterized using X-ray powder diffractometry (XRD) and surface area analyzer. Briefly, 500 mL of 1 M FeCl3 was titrated with 1 M NaOH solution with vigorous stirring until the brown color suspensions were formed at pH 7.0–7.2. The precipitates were harvested by centrifugation at 10 000 rpm for 5 min and washed three times with anoxic bidistilled water to remove dissolved ions. The resulting ferric oxide suspensions were then degassed and gassed with N2 gas several times and maintained under N2 atmosphere in tightly sealed serum bottles. The ferrihydrite suspensions were autoclaved afterwards and stored at room temperature under anoxic conditions. The surface area of ferrihydrite, determined by a BET N2 adsorption surface area analyzer (Micromeritics, ASAP 2020), was 313 m2 g 1. 2.2. Microorganism and cultivation G. sulfurreducens was cultivated in bicarbonate-buffered mineral medium at pH 7.2 ± 0.1 as previously reported (Doong and Schink, 2002). The compositions of mineral media used for growth contained the following mineral salts (g L 1): NH4Cl, 0.25; NaCl, 1; MgCl2
6H2O, 0.4; KCl, 0.5; CaCl2
2H2O, 0.15; KH2PO4, 0.2. After autoclaving and cooling under an atmosphere of N2/CO2 (80/20, v/v), 30 mM sodium bicarbonate buffer solution, 1 mL of trace mineral, vitamin, and selenite–tungstate solutions were added per liter. In addition, 20 mM acetate and 40 mM fumarate solutions were added as electron donor and acceptor, respectively for the growth of G. sulfurreducens. Trace elements and selenite–tungstate solution contained the following chemicals (g L 1): Fe(NH4)2(SO4)
6H2O, 0.8, CoCl2
6H2O, 0.2, ZnSO4
7H2O, 0.2, CuCl2
2H2O, 0.02, NiCl2
6H2O, 0.02, Na2SeO4, 0.02, Na2MoO4
2H2O, 0.02, and Na2WO4, 0.02. All cultures were incubated at 25 ± 1 °C in the dark, and the purity was checked by optical microscopy at regular intervals. In addition, the optical density at 660 nm (OD660) was used to monitor the growth of G. sulfurreducens during the incubation period. In addition, the growth rate of G. sulfurreducens was determined by the slope of growth curve in the exponential phase. 2.3. Reductive dissolution of ferrihydrite Batch experiments for microbial reduction of ferrihydrite were conducted under anoxic conditions using 60-mL serum bottles. High-purity N2 was introduced into the bottles to maintain the anaerobic conditions during the experimental periods. Anoxic 30 mM bicarbonate buffer solutions were used to maintain the solution pH at 7.2 ± 0.1. The 10 mM ferrihydrite and 20 mM acetate

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were introduced into serum bottles using N2-purged sterilized syringes. In addition, 10–200 lM quinone compounds including AQDS, LQ, NQ, UQ, and JQ and 1 mL of G. sulfurreducens in the exponential phase (OD660 0.5–0.6) were then injected into the serum bottles to get the total volume of 50 mL. The serum bottles were then incubated horizontally in an orbital shaker at 120 rpm at 25 ± 1 °C in dark. All the samples were performed triplicate or quadruplicate. 2.4. Dechlorination of CT The dechlorination experiments were conducted in batch-fed modes using serum bottles sealed with Teflon-lined rubber septa and aluminum crimp caps. An appropriate amount of CT stock solution dissolved in deoxygenated methanol was delivered into the serum bottles through a sterilized gas-tight glass syringe to obtain a final concentration of 5 lM. Fig. SM-1 (see Supplementary data) shows the growth of G. sulfurreducens in the presence and absence of methanol. It is clear that addition of 10–100 lM methanol has little effect on the growth of G. sulfurreducens under anaerobic conditions. In addition, 10–20 lM quinone compounds, 10 mM ferrihydrite, 20 mM acetate, and 1 mL of G. sulfurreducens in the exponential phase were added to the solution containing 30 mM bicarbonate buffer. The total volume of liquid phase in the serum bottle was maintained at 50 mL, resulting in a 10-mL volume for headspace analysis. In sterilized experiments, media were introduced into the serum bottles through 0.2-lm sterilized PTFE membrane filters to understand the effect of solution ingredients on abiotic reduction of iron oxides. In addition, parallel control experiments in the absence of Fe(III) oxides, electron shuttle and/ or bacteria (sterilized control) were also performed to evaluate the activity of G. sulfurreducens and electron shuttles toward CT dechlorination. All the serum bottles were incubated in the dark on an orbital shaker at 120 rpm and 25 ± 1 °C. 2.5. Analytical methods The headspace analytical technique was used for the determination of chlorinated hydrocarbons. The concentrations of CT and the byproducts in the headspace of the test bottles were monitored by withdrawing 80 lL of gas in the headspace using a 100-lL gastight syringe. The headspace sample was immediately injected into a gas chromatograph (GC) equipped with an electron capture detector (ECD) and a flame ionization detector (FID) (Perkin–Elmer, Autosystem, Norwark, CT). A 60-m VOCOL fused-silica megabore capillary column (0.545 mm  3.0 lm, Supelco Co.) was used to separate the organic compounds. The column was connected to the FID and ECD simultaneously by a Y-splitter; and an optimal sensitivity for chlorinated hydrocarbons was achieved with 40% of the flow (1.85 mL min 1) to ECD. The column temperature was maintained isothermally at 120 °C using ultra-high purity nitrogen (>99.9995%) as the carrier gas at a flow rate of 5.6 mL min 1 (linear velocity of 42.3 cm s 1). The temperatures of ECD and FID were maintained at 325 °C and 250 °C, respectively. The relative standard deviation (RSD) for analysis was controlled within 10%. The reduction of Fe(III) was examined by periodical analysis of Fe(II) by ferrozine method. Concentrations of total HCl-extractable Fe(II) in serum bottles were monitored by withdrawing 0.5 mL of suspension using N2-purged syringes and were immediately acidified with 0.5 mL of 1 M HCl. After mixing vigorously, the acidified samples were centrifuged at 10 000 rpm for 10 min to remove particles and the Fe(II) concentrations in solutions were determined by UV/Vis spectrophotometer (HITACH U-3010, Tokyo, Japan) at 562 nm. The organic radicals were examined using the electron paramagnetic resonance (EPR) spectrometer (Bruker, EMX-10,

Germany) working at X-band frequency of 9.78 GHz with microwave power of 8.02 mW. For UV–Vis spectroscopic absorbance analysis, aqueous samples were transferred from the serum bottles through a N2-purged syringe to 1-cm quartz cuvettes capped with rubber septa. A Hitachi U-3010 UV–Vis spectrophotometer running UV solutions software was used to identify the changes in the redox states of quinone compounds over a wavelength range from 200 to 600 nm at a scan rate of 120 nm min 1. In addition, the reduced quinone compounds were chemically generated by adding quinone compounds into the oxygen-free solution containing 50 mg palladium powder (Pd0) under hydrogen atmosphere.

3. Results and discussion 3.1. Transformation of ferrihydrite in the presence of quinone moieties Quinone moiety may have a negative impact on the growth of DIRB. Therefore, the growth of G. sulfurreducens in the presence and absence of 100 lM quinone compounds by using 20 mM acetate and 40 mM fumarate as the electron donor and acceptor, respectively, was first investigated to evaluate the toxicity of quinine compounds to G. sulfurreducens. Fig. SM-2 (see Supplementary data) shows the growth of G. sulfurreducens in the presence of various quinones under anaerobic conditions. The growth of G. sulfurreducens in the absence of quinone compounds could reach the OD660 of 0.615 after incubation of 4 d with the growth rate of 0.235 d 1. Addition of quinone compounds decreased the OD660 to 0.081–0.518 after 14 d and the growth rate of G. sulfurreducens was 0.117 d 1 for AQDS, 0.099 d 1 for LQ, and 0.021 d 1 for NQ. No obvious growth of G. sulfurreducens was observed when UQ or JQ was added. These results clearly show that addition of naturally occurring quinone compounds would inhibit the growth of G. sulfurreducens, which may also influence the reductive dissolution efficiency and rate of ferrihydrite by G. sulfurreducens. Fig. 1 shows the biotransformation of 10 mM ferrihydrite by G. sulfurreducens in the presence of various concentrations of quinone compounds in bicarbonate buffer at pH 7.0 ± 0.2. In the absence of electron shuttling compounds, the biotransformation extent of ferrihydrite by G. sulfurreducens alone was low, and only 0.39 ± 0.02 mM of Fe(II) were produced after incubation of 80 h. Addition of quinone compounds has different effects on the bioreduction of ferrihydrite by G. sulfurreducens. Addition of AQDS significantly enhanced the production of Fe(II) within 80 h, and the maximum production concentration of Fe(II) was (4.4 ± 0.2)– (5.0 ± 0.1) mM at 10–200 lM AQDS (Fig. 1(a)). This result clearly indicates that the generation of Fe(II) is mainly attributed to the microbial Fe(III) reduction in the presence of AQDS and 10 lM AQDS is sufficient for reductive dissolution of ferrihydrite to ferrous ions. Similar to the enhanced behaviors of AQDS, the concentration of Fe(II) produced from the reductive dissolution of ferrihydrite by G. sulfurreducens increased with incubation time after addition of LQ and NQ, and the enhanced effect is concentration-dependent. Although 10 lM LQ could produce 7.3 ± 0.4 mM Fe(II) after incubation of 9 d, the addition of high concentration of LQ decreased the production efficiency of Fe(II), and only 5.4 ± 0.3 and 1.4 ± 0.2 mM Fe(II) were produced at 50 and 100 lM LQ, respectively (Fig. 1(b)). This result may be attributed to the inhibition of growth of G. sulfurreducens at high concentration of LQ. As shown in Fig. SM-3 (see Supplementary data), the optical density of G. sulfurreducens started to increase after incubation of 8 d, and the growth rates were 0.058 and 0.037 d 1 when LQ concentrations were 50 and 100 lM, respectively, which are lower than that at 10 lM LQ. It is noted that the produced Fe(II) concentrations at 10 lM LQ decreased to 5.6 ± 0.3 mM after incubation of 24 d, presumably

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Fig. 1. The reductive dissolution of 10 mM ferrihydrite by G. sulfurreducens in the presence of 10–200 lM (a) AQDS, (b) LQ, (c) NQ, and (d) UQ under iron-reducing conditions in bicarbonate buffer at pH 7.2 ± 0.1. The error bar indicates one standard deviation of the mean value.

attributed to the formation of precipitates (Zachara et al., 2002). Similarly, low concentration of NQ enhanced the production of Fe(II) from 0.39 ± 0.02 mM in the absence of electron mediator to 4.4 ± 0.2 mM at 10 lM NQ, and then decreased to (1.6 ± 0.1)– (2.5 ± 0.2) mM at 50–200 lM NQ (Fig. 1(c)). However, addition of UQ (Fig. 1(d)) and JQ (Fig. SM-4, see Supplementary data) in the concentration range 10–200 lM had little effect on the enhancement of reductive dissolution of ferrihydrite, and the produced Fe(II) concentrations were 0.05–0.41 mM for UQ and 0.03– 0.07 mM for JQ. Both the optical density of G. sulfurreducens in the presence of UQ and JQ were lower than 0.1 after incubation of 14 d (Fig. SM-2), clearly indicating that the low production efficiency of Fe(II) is mainly attributed to the low growth rate of G. sulfurreducens. The secondary minerals of iron oxides after reductive dissolution in the presence of quinone compounds were different. Fig. 2 shows the SEM images and XRD patterns of secondary minerals after reductive dissolution of ferrihydrite in the presence of quinone compounds. After bioreduction of ferrihydrite by G. sulfurreducens in the presence of AQDS and NQ, the original dark-brown color of ferrihydrite gradually transformed to a black-colored product and could be collected magnetically, indicting the formation of magnetite (Fe3O4) (Fig. 2(a)). The XRD patterns of secondary minerals after bioreduction of ferrihydrite showed peaks at 30.08°, 35.46°, 43.31°, 56.96°, and 62.79° 2h, which are consistent with the standard patterns of magnetite (Fig. 2(c)). Different from the AQDS and NQ, the rhombohedra-shaped crystals were formed in the LQ-amended batches after incubation of 10 d (Fig. 2(b)). The

XRD patterns showed peaks at 24.46°, 31.86°, 38.10°, 42.06°, 45.84°, and 52.44°, which can be assigned as the (0 1 2), (1 0 4), (1 1 0), (1 1 3), (2 0 2), and (0 1 8) orientations of siderite (FeCO3) (Fig. 2(d)). This may be attributed to the fact that high concentration of Fe(II) was produced in the presence of LQ, and then reacted with bicarbonate buffer to form siderite. Fig. 3 shows the bioreduction ratios and dissolution rate constant of ferrihydrite as a function of quinone concentration. The bioreduction ratio of ferrihydrite, defined as the ratio of produced Fe(II) concentration to the added concentration of ferrihydrite, in the presence of AQDS maintained at a relatively stable value of 44–50%. Several studies have investigated the reductive dissolution of crystalline Fe(III) (hydr)oxides and ferrihydrite by Shewanella putrefaciens CN32 in the absence of electron shuttling compounds and found that 25–43.5% of ferrihydrite can be bioreduced (Langley et al., 2009; Ekstrom et al., 2010). Maithreepala and Doong (2009) have investigated the reductive dissolution of ferrihydrite by G. sulfurreducens in the 10 mM HEPES buffer using 10 lM AQDS as the electron shuttling compound, and found that the conversion ratio was 48.3 ± 1.2% after incubation of 9 d, which is in good agreement with the results obtained in this study. In addition, the conversion ratios of ferrihydrite by G. sulfurreducens were 30–73% and 16–44% when 10–200 lM LQ and NQ, respectively, were amended as the electron mediator, clearly showing that the conversion efficiency of ferrihydrite is highly dependent on chemical structure and concentration of quinone compounds. The production of Fe(II) from the reductive dissolution of ferrihydrite by G. sulfurreducens in the presence of quinone

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compounds followed the pseudo-first-order kinetics (Doong and Schink, 2002), and the pseudo-first-order rate constant (kobsFe) for Fe(II) production increased from 0.62 ± 0.03 to 0.72 ± 0.01 d 1 as the AQDS concentration increased from 10 to 200 lM, which is 17–20 times higher than that in the control system without the addition of electron shuttling compound (0.036 ± 0.001 d 1). In addition, the kobsFe for Fe(II) production from the dissolution of ferrihydrite in the presence of 10–200 lM LQ and NQ were (0.061 ± 0.006)–(0.155 ± 0.008) d 1 and (0.048 ± 0.008)– (0.137 ± 0.008) d 1, respectively. Langley et al. (2009) investigated the biological reduction of putative biogenic iron oxides in volcano sediments from Tonga-Kermadec Arc, southwest Pacific Ocean by S. putrefaciens CN32 and found that the bioreduction rate was 0.024 d 1 for ferrihydrite and 0.047–0.052 d 1 for volcano sediments. Glasauer et al. (2003) used S. putrefaciens CN32 for bioreduction of hydrous ferric oxides in a mineral growth medium and the bioreduction rates were in the range 0.021–0.036 d 1. In addition, Doong and Schink (2002) used 0.5–2 mM cysteine as the electron mediation for reductive dissolution of ferrihydrite by G. sulfurreducens. The efficiency of ferrihydrite bioreduction was significantly enhanced in the presence of cysteine and the bioreduction rates of ferrihydrite increased from 0.21 d 1 at 0.5 mM cysteine to 0.84 d 1 at 2 mM cysteine. In this study, the bioreduction rate of ferrihydrite by G. sulfurreducens in the presence of different quinone moieties were in the range 0.036–0.72 d 1, which is in good agreement with those reported data. It is noteworthy that the kobsFe for Fe(II) production followed the order AQDS  LQ > NQ. However, the produced Fe(II) concentration in the presence of AQDS was lower than that of low concentration (10–50 lM) of LQ for a long period of time. AQDS has an excellent electron-accepting capacity which can produce higher organic radicals than those of naturally occurring humic acids after microbial reduction (Scott et al., 1998). Therefore, the production rate of Fe(II) in AQDS-amended batched is higher than those of naturally occurring quinones. In contrast, the dissolution of

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(a) 5 CT concentration (μM)

ferrihydrite in the presence of LQ and NQ is relatively low and the aqueous Fe(II) would replace the hydrogen ions in surface-OH groups of ferric oxides to form inner-sphere surface complexes, resulting in the formation of surface-bound ferroferric oxides (Maithreepala and Doong, 2004, 2005). Results obtained in this study clearly show that addition of AQDS could produce higher Fe(II) concentration than naturally occurring quinone compounds in a short time. Since AQDS is an artificial analogue for the investigation of reductive dissolution of iron oxides and biotransformation of priority pollutants, this result means that the naturally occurring quinone compounds may be the more suitable electron mediator than AQDS for understanding the production of Fe(II) and reactivity of iron species under iron-reducing conditions.

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Time (day)

(c) 5 CT concentration (μM)

The dechlorination of CT by the produced biogenic iron species in the presence of G. sulfurreducens and quinone compounds under iron-reduction conditions was further investigated. Fig. 4 shows the biotransformation of CT in the presence of G. sulfurreducens and various concentrations of quinone compounds in the bicarbonate buffer under iron-reducing conditions. In the absence of quinone compounds, 22% of the original CT was biotransformed in the presence of 10 mM ferrihydrite and G. sulfurreducens after the incubation of 28 d (Fig. 4(a)). In addition, trace amount of chloroform and dichloromethane with carbon mass balance of 87% was observed. Less than 10% of CT was removed within 28 d when solution only contained G. sulfurreducens or 10 mM ferrihydrite, showing that the reduction of CT was mainly facilitated by the biological reaction. It is noted that no obvious decrease in CT concentration was observed in the sterilized control, indicating that no leakage or adsorption occurred during the experimental course. Addition of quinone compounds significantly enhanced the dechlorination efficiency and rate of CT mediated by G. sulfurreducens and biogenic species. The biotransformation efficiencies of CT, defined as the percentage of CT consumed to the added CT concentration, were 72–76% in the presence of 10–100 lM AQDS, which indicate that the change in AQDS concentration has little influence on the dechlorination efficiency and rate of CT under anaerobic conditions. Several studies have depicted that the biogenic iron concentration and species play important roles in dechlorination of chlorinated hydrocarbons (Elsner et al., 2004; McCormick and Adriaens, 2004; Maithreepala and Doong, 2009). In this study, addition of 10–100 lM AQDS produced similar biogenic Fe(II) concentration and species, resulting in the similar dechlorination efficiency and rate of CT. Different from the AQDS-amended batches, the dechlorination efficiency and rate of CT was dependent on the added concentration of LQ and NQ. As shown in Fig. 4b, the dechlorination efficiency of CT mediated by G. sulfurreducens and 10 lM LQ was 80% after incubation of 20 d, and then decreased to 70% and 45% when 50 and 100 lM LQ, respectively, was added. The decrease in dechlorination efficiency of CT at high LQ concentration is mainly attributed to the low production rate of Fe(II). Similar to the LQ–amended batches, the biotransformation efficiency of CT decreased upon increasing NQ concentration and 40–74% of CT was dechlorinated by G. sulfurreducens in 30 mM bicarbonate buffer solutions containing 10–100 lM NQ (Fig. 4(c)). Biotransformation of CT is a complex process and several pathways including reductive dechlorination, a-elimination, C–C coupling reaction, and dihaloelimination are involved in the dechlorination of CT (Penny et al., 2010). Several products such as chloroform(CHCl3), dichloromethane (CH2Cl2), methane (CH4), CO, CS2, and cell-bound products have been identified as the intermediates for CT biotransformation under anaerobic conditions (McCormick and Adriaens, 2004; Maithreepala and Doong, 2009;

5

4

3

2

Sterilized control 10 μM NQ only 10 μM NQ 50 μM NQ 100 μM NQ

1

0

0

5

10

15

20

25

30

Time (day) Fig. 4. The dechlorination of CT in the presence of G. sulfurreducens and ferrihydrite in bicarbonate buffer at pH 7.2 ± 0.1. The error bar indicates one standard deviation of the mean value.

Penny et al., 2010; Bae and Lee, 2012). In this study, chloroform and dichloromethane were found to be the predominant biotransformation products for CT dechlorination in the presence of various quinone compounds, indicating that reductive dechlorination is the major reaction mechanism for CT biotransformation in the presence of G. sulfurreducens and quinone compounds under iron-reducing conditions. It is noteworthy that the carbon mass balance in all batches was only 42–56% after the incubation of 28 d, suggesting the formation of non-volatile compounds that cannot be detected by GC-ECD or GC-FID in the headspace. Several studies have reported that the biotransformation of CT by G. metallireducens and S. putrefaciens produced low concentrations of chloroform, dichloromethane, methane, cell-bound products, and other

R.-a. Doong et al. / Chemosphere 97 (2014) 54–63

-1

kobsCT for CT dechlorination (d )

(a) 0.07 AQDS LQ NQ

0.06 0.05 0.04 0.03

y = 0.00008x + 0.0055 2 r = 0.961

0.02 0.01 0.00

0

10

20

30

40

50

60

70

80

Bioreduction ratio of ferrihydrite (%)

(b) 0.07 -1

unidentified products (McCormick and Adriaens, 2004; Maithreepala and Doong, 2009; Bae and Lee, 2012). McCormick and Adriaens (2004) investigated the dechlorination of 13 lM CT by nanoscale biogenic magnetite/maghemite particles produced from G. metallireducens and chloroform, CO, CH4, and a trace amount of tetrachloroethylene were identified as the products. Bae and Lee (2012) used S. putrefaciens CN32 to dechlorinate CT in the presence of iron-bearing soil minerals, and found that chloroform, carbon monoxide, and formate were the main products for the degradation of CT by magnetite and lepidocrocite. These results imply that other dechlorination mechanisms may also be involved in CT transformation mediated by G. sulfurreducens and quinone compounds under iron-reducing conditions. Table 1 summaries the pseudo-first-order rate constant (kobsCT) for CT dechlorination in batches containing G. sulfurreducens and various concentrations of quinone compounds under iron-reducing conditions. It is clearly that the kobsCT values for CT dechlorination in the presence of 10–100 lM AQDS were in the range (0.042 ± 0.005)–(0.044 ± 0.006) d 1, which were 5.4–5.8 times higher than that in the absence of quinone compound (0.008 ± 0.001 d 1). The kobsCT for CT dechlorination in LQ- and NQ-amended batches were (0.037 ± 0.003)–(0.061 ± 0.004) and (0.019 ± 0.002)–(0.044 ± 0.002) d 1, respectively, which were 4.6– 7.4 and 2.4–5.8 times higher than those in the absence of quinones. These results clearly show that addition of naturally occurring quinone compounds including LQ and NQ is effective on the enhancement of dechlorination efficiency and rate of CT in the presence of G. sulfurreducens under iron-reducing conditions. The chlorinated hydrocarbons can undergo both biotic and abiotic dechlorination in natural environments as well as in wastewater treatment facilities. The biogenic and surface-bound iron species have been shown to play a crucial role in dechlorination of chlorinated hydrocarbon under iron-reducing conditions (Elsner et al., 2004; Tao et al., 2010; Bae and Lee, 2012). The biogenic Fe(II) can be produced from the reductive dissolution of amorphous and crystalline iron oxides by DIRB and then adsorbed onto the surface of secondary iron oxide minerals, resulting in the formation of reactive surface-bound Fe(II) species for the dechlorination of CT (Maithreepala and Doong, 2009). Fig. 5 shows the kobsCT for CT dechlorination as a function of bioreduction ratio of ferrihydrite and kobsFe for Fe(II) production. In Fig. 5(a), the maximum bioreduction ratio of ferrihydrite was used for correlation to avoid the interference of siderite produced in LQ-amended solutions. A good linear relationship between kobsCT for CT dechlorination and maximum conversion ratio of ferrihydrite to Fe(II) was observed (r2 = 0.961). On the contrary, no obvious relationship between kobsCT for CT dechlorination and kobsFe for Fe(II) production was observed in the presence of various quinone compounds. However, a satisfactory linear relationship between kobsCT for CT dechlorination and kobsFe for production of Fe(II) was obtained when only LQ and NQ were encountered (r2 = 0.831) (Fig. 5(b)). This result clearly indicates that the dechlorination of CT in the presence of G. sulfurreducens and ferrihydrite is mainly from the production of biogenic iron species. The slow production rate of Fe(II) by G.

kobsCT for CT dechlorination (d )

60

0.06 0.05

y = 0.321x + 0.0007 2 r = 0.831

0.04 0.03

AQDS LQ NQ

0.02 0.01 0.00

0.0

0.1

0.2

0.3

0.4

0.5

0.6

0.7

-1

kobsFe for Fe(II) production (d ) Fig. 5. The pseudo-first-order rate constant (kobsCT) for CT dechlorination as a function of (a) bioreduction ratio of ferrihydrite and (b) kobsFe for Fe(II) production. The error bar indicates one standard deviation of the mean value.

sulfurreducens in the presence of naturally occurring quinones such as LQ and NQ would provide new reactive iron species for biotransformation over a long period of time, and subsequently enhances the dechlorination efficiency and rate of CT by G. sulfurreducens. These results clearly demonstrate the feasibility of using naturally occurring quinone compounds to constantly produce biogenic Fe(II). The biogenic Fe(II) would then adsorb onto the surface of ferric oxides to form surface-bound iron species, resulting in the acceleration of the dechlorination efficiency and rate of CT under anaerobic conditions. This process could facilitate the development of treatment processes that is useful for the enhanced biotransformation of chlorinated hydrocarbons containing ferric oxides and DIRB. It should be noted that the acetate concentration used in this study was 20 mM, which is higher than what is practical in relation to the concentration of CT in the contaminated aquifers. Further study using low concentrations of acetate is required to show the feasibility of using naturally occurring quinine for natural attenuation of chlorinate hydrocarbons in the contaminated sites.

Table 1 The pseudo-first-order rate constant (kobsCT) for CT dechlorination by G. sulfurreducens in the presence of various concentrations of quinone compounds under iron-reducing conditions at pH 7.2 ± 0.1. The medium contained 30 mM bicarbonate buffer, 20 mM acetate, and 10 mM ferrihydrite. Concentration (lM)

AQDS kobsCT (d

0 10 50 100

LQ 1

)

0.008 ± 0.001 0.044 ± 0.004 0.044 ± 0.006 0.042 ± 0.005

NQ

r2

kobsCT (d

1

0.991 0.993 0.994 0.991

0.008 ± 0.001 0.061 ± 0.007 0.044 ± 0.005 0.037 ± 0.003

)

r2

kobsCT (d

1

0.991 0.987 0.989 0.990

0.008 ± 0.001 0.044 ± 0.002 0.026 ± 0.003 0.019 ± 0.002

)

r2 0.991 0.994 0.984 0.976

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(a)

(d)8000

100 μM AQDS

3

0.5 mM AQDS

6000

Intensity (a.u.)

Absorbance (a.u.)

4000

AQDS+ Pd + H2

2

AQDS alone AQDS +G. sulfurreducens

1

2000 0 -2000 -4000 -6000 -8000

0 200

300

400

500

3480

600

3482

3484

3486

3488

3490

3492

3494

Magnetic field (G)

Wavelength (nm)

(b)4

(e) 1500

100 μM NQ

100 μM NQ

1000

Absorbance (a.u.)

Intensity (a.u.)

NQ + Pd + H2

3

NQ + G. sulfurreducens NQ alone 2

1

500 0 -500 -1000 -1500

0 200

300

400

500

3478

600

3480

3482

(c)

3486

3488

(f) 1500

100 μM LQ

3

3484

3490

3492

3494

Magnetic field (G)

Wavelength (nm)

100 μM LQ

Intensity (a.u.)

Absorbance (a.u.)

1000

2

LQ+ Pd + H2 LQ + G. sulfurreducens LQ alone

500 0 -500

1 -1000 -1500 0 200

300

400

500

600

Wavelength (nm)

3478

3480

3482

3484 3486

3488 3490

3492

3494

Magnetic field (G)

Fig. 6. The UV–visible and EPR spectra of quinone compounds at pH 7.2 ± 0.1. (a), (b), and (c) are UV–Vis spectra of AQDS, NQ, and LQ, while (d), (e), and (f) are EPR spectra of AQDS, NQ, and LQ, respectively.

In addition, the final levels of CT achieved in the experiments after incubation of 30 d are much higher than the regulatory limit of 5 lg L 1. According to the kobsCT for CT dechlorination obtained in this study, a period of 80–272 d is needed to lower the CT concentration from 5 lM to the regulatory limit in the presence of various quinone compounds under iron-reducing conditions. 3.3. Identification of active quinone species To further elucidate the change in chemical structures of quinone compounds, UV–Vis spectroscopy and EPR were used to

characterize the active species of quinone compounds in the presence of G. sulfurreducens under anaerobic condition. Fig. 6 shows the UV–visible and EPR spectra of 100 lM quinone compounds in solutions containing G. sulfurreducens at pH 7.2 ± 0.1. In addition, palladium was used as a catalyst to chemically reduce quinone compounds in the presence of hydrogen gas for comparison. As shown in Fig. 6(a), an absorbance peak at 329 nm was clearly observed for original AQDS curve. A bathochromic shift from 329 to 406 nm was observed in AQDS spectrum when G. sulfurreducens was added into the solution. This shift is in good agreement with the spectrum of chemically reduced AQDS in the presence of

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R.-a. Doong et al. / Chemosphere 97 (2014) 54–63

palladium and H2, which indicates that G. sulfurreducens can reduce AQDS to hydroquinone (AHQDS) under neutral conditions (Jiang et al., 2009). In addition, the bathochromic shift in the UV– Vis spectra may be attributed to the formation of semiquinones (Doong and Chiang, 2005) or species derived from substitution (Perlinger et al., 2002). Two absorbance peaks at 387 and 406 nm were observed in the AQDS spectra, which are similar to the semiquinone radical absorption of anthraquinone-1-sulfonate (AQ-1-S) (400 nm) (Doong and Chiang, 2005). Similar to the AQDS, the UV–visible spectra of original NQ showed a peak at 251 nm, and a bathochromic shift to 283, 335, and 400 nm was observed in the presence of G. sulfurreducens under anaerobic conditions, which is in good agreement with the spectrum of chemically reduced NQ by palladium in the presence of hydrogen gas. For LQ-amended batches, a hypsochromic shift occurred when G. sulfurreducens was added to solutions. The measured spectrum of original LQ showed peaks at 265 nm, and slightly shifted to 241 nm after reduction by G. sulfurreducens or palladium/H2. Zhou and Rieker (1997) reported that hydroquinone compounds displayed a hypsochromic shift when they possess aromatic rings. This observation reflects the fact that LQ may be partially reduced to hydroquinones by G. sulfurreducens under anaerobic conditions. ERP was further used to characterize the organic radicals produced from quinones in the presence of G. sulfurreducens under neutral conditions. As indicated in Fig. 6(d), addition of 100 lM AQDS to the solution containing G. sulfurreducens showed signals in the EPR spectrum, suggesting the formation of organic radicals in the presence of G. sulfurreducens at pH 7.2. A small EPR signal was also observed for NQ-amended solution containing G. sulfurreducens, suggesting the possible generation of semiquinones from NQ. In contrast, no signal was observed in the EPR spectrum of LQ in the presence of G. sulfurreducens under anaerobic conditions, clearly showing that the reduced species of LQ does not produce organic radicals. These results clearly indicate that AQDS and NQ produced organic radicals and hydroquione species, while only hydroquinones were generated in the presence of LQ and G. sulfurreducens to serve as the redox mediators for biotransformation of CT.

4. Conclusions In this study, we have demonstrated the enhancement effect of naturally occurring quinones on the reductive dissolution of ferrihydrite and the dechlorination of CT under iron-reducing conditions. CT is a persistent compound under iron-reducing conditions and AQDS is an active electron mediator which can rapidly bioreduce ferrihydrite by G. sulfurreducens in a short period. Different from the enhancement effect of AQDS, the dissolution efficiency of ferrihydrite by naturally occurring quinones is highly dependent on the added concentration and structural property. In addition, the bioreduction efficiency of ferrihydrite in the presence of low concentration of LQ is higher than that of AQDS, and the slow and constant production of Fe(II) by G. sulfurreducens in the presence of LQ and NQ provides new reactive iron species for enhanced biotransformation of CT over a long period of time under iron-reducing conditions. A good relationship between kobsCT for CT dechlorination and conversion efficiency of ferrihydrite was observed, indicating that the biogenic Fe(II) plays an important role in dechlorination of CT under iron-reducing conditions. The UV– visible and EPR spectroscopic results showed that AQDS and NQ produced semiquinones and hydroquinones for dechlorination of CT, while only hydroquinones were produced in LQ-amended batched in the presence of G. sulfurreducens. The naturally occurring quinone moiety is widespread in both terrestrial and aquatic

environments and the utilization of quinones as electron transfer mediators has been investigated for the transformation of a number of organic contaminants. Results obtained in this study clearly indicate the feasibility of using naturally occurring quinone compounds to facilitate the natural attenuation of CT and other chlorinated hydrocarbons in the subsurface environments. Acknowledgement The authors thank the National Science Council, Taiwan for financial support under Contract No. NSC 101-2221-E-007-084MY3. Appendix A. Supplementary material Supplementary data associated with this article can be found, in the online version, at http://dx.doi.org/10.1016/ j.chemosphere.2013.11.004. References Amir, A., Lee, W.J., 2012. Enhanced reductive dechlorination of tetrachloroethene during reduction of cobalamin (III) by nano-mackinawite. J. Hazard. Mater. 235–236, 359–366. Aulenta, F., Di Maio, V., Ferri, T., Majone, M., 2010. The humic acid analogue antraquinone-2,6-disulfonate (AQDS) serves as an electron shuttle in the electricity-driven microbial dechlorination of trichloroethene to cisdichloroethene. Bioresour. Technol. 101, 9728–9733. Bae, S., Lee, W.J., 2012. Enhanced reductive degradation of carbon tetrachloride by biogenic vivianite and Fe(II). Geochim. Cosmochim. Acta 85, 170–186. Borch, T., Kretzschmar, R., Kappler, A., van Cappellen, P., Ginder-Vogel, M., Voegelin, A., Campbell, K., 2010. Biogeochemical redox processes and their impact on contaminant dynamics. Environ. Sci. Technol. 44, 15–23. Bose, S., Hochella, M.F., Gorby, Y.A., Kennedy, D.W., McCready, D.E., Madden, A.S., Lower, B.H., 2009. Bioerduction of hematite nanoparticles by the dissimilatory iron reducing bacterium Shewanella oneidensis MR-1. Geochim. Cosmochim. Acta 73, 962–976. Buchholz, A., Laskov, C., Haderlein, S.B., 2011. Effects of zwitterionic buffers on sorption of ferrous iron at goethite and its oxidation by CCl4. Environ. Sci. Technol. 45, 3355–3360. Cao, F., Liu, T.X., Wu, C.Y., Li, F.B., Li, X.M., Yu, H.Y., Hong, H., Chen, M.J., 2012. Enhanced biotransformation of DDTs by an iron- and humic-reducing bacteria Aeromonas hydrophila HS01 upon addition of goethite and anthraquinone-2,6disulphonic disodium salt (AQDS). J. Agric. Food Chem. 60, 11238–11244. Cory, R.M., McKnight, D.M., 2005. Fluorescence spectroscopy reveals ubiquitous presence of oxidized and reduced quinones in dissolved organic matter. Environ. Sci. Technol. 39, 8142–8149. Doong, R.A., Chiang, H.C., 2005. Dechlorination of carbon tetrachloride by thiol compounds in the presence of quinone compounds. Environ. Sci. Technol. 39, 7460–7568. Doong, R.A., Schink, B., 2002. Cysteine-mediated reductive dissolution of poorly crystalline iron(III) oxides by Geobacter sulfurreducens. Environ. Sci. Technol. 36, 2939–2945. Doong, R.A., Chen, K.T., Tsai, H.C., 2003. Reductive dechlorination of carbon tetrachloride and tetrachloroethylene by zerovalent silicon-iron reductants. Environ. Sci. Technol. 37, 2575–2581. Ekstrom, E.B., Learman, D.R., Madden, A.S., Hansel, C.M., 2010. Contrasting effects of Al substitution on microbial reduction of Fe(III) (hydr)oxides. Geochim. Cosmochim. Acta 74, 7086–7099. Elsner, M., Schwarzenbach, R.P., Haderlein, S.B., 2004. Reactivity of Fe(II)-bearing minerals toward reductive transformation of organic contaminants. Environ. Sci. Technol. 38, 799–807. Garcia-Balboa, C., Bedoya, I.C., Gonzalez, F., Blazquez, M.L., Munoz, J.A., Ballester, A., 2010. Bioreduction of Fe(III) ores using three pure strains of Aeromonas hydrophila, Serratia fonticola and Clostridium celerecrescens and a natural consortium. Bioresour. Technol. 101, 7864–7871. Glasauer, S., Weidler, P.G., Langley, S., Beveridge, T.J., 2003. Controls on Fe reduction and mineral formation by a subsurface bacterium. Geochim. Cosmochim. Acta 67, 1277–1288. Jeong, H.Y., Hayes, K.F., 2007. Reductive dechlorination of tetrachloroethylene and trichloroethylene by mackinawite (FeS) in the presence of metals: Reaction rates. Environ. Sci. Technol. 41, 6390–6396. Jiang, J., Bauer, I., Paul, A., Kappler, A., 2009. Arsenic redox changes by microbially and chemically formed semiquinone radicals and hydroquinones in a humic substance model quinone. Environ. Sci. Technol. 43, 3639–3645. Kwon, M.J., Finneran, K.T., 2006. Microbially mediated biodegradation of hexahydro-1,3,5- trinitro-1,3,5-triazine by extracellular electron shuttling compounds. Appl. Environ. Microbiol. 72. 5933–5741.

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