Water Air Soil Pollut (2014) 225:1904 DOI 10.1007/s11270-014-1904-2

Characterization of Five Chromium-Removing Bacteria Isolated from Chromium-Contaminated Soil Zhiguo He & Shuzhen Li & Lisha Wang & Hui Zhong

Received: 21 January 2013 / Accepted: 10 February 2014 / Published online: 21 February 2014 # Springer International Publishing Switzerland 2014

Abstract The potential for bioremediation of chromium pollution using bacteria was investigated in this study. Five chromium-removing bacteria strains were successfully isolated from Cr(VI)contaminated soils and identified by their 16S rRNA gene sequences. The optimum growth temperature (30–40 °C) and pH (8.5– 11) for the five isolates were investigated. The effect of initial Cr(VI) concentrations (0–1,575 mg L−1) on bacterial growth was also studied. Results showed that Pseudochrobactrum saccharolyticum strain W1 had high chromium-removing ability and could grow at Cr(VI) concentrations from 0 to 1,225 mg L−1. To our knowledge, this is the first report of chromium removal by a member of the Pseudochrobactrum genus. Sporosarcina saromensis W5 had the highest chromium-removing rate of 0.79 mg h−1 mg−1 biomass. Exopolysaccharide (EPS) production and components of the five bacteria strains were also investigated, and a positive relationship was found between the bacterial chromium removal and EPS production. Z. He : S. Li : L. Wang School of Minerals Processing and Bioengineering, Central South University, Changsha 410083, People’s Republic of China Z. He : S. Li : L. Wang Key Laboratory of Biohydrometallurgy of Ministry of Education, Central South University, Changsha, Hunan, China 410083 H. Zhong (*) School of Life Science, Central South University, Changsha, Hunan, China 410012 e-mail: [email protected]

Keywords Chromium-removing bacteria . Pseudochrobactrum saccharolyticum . Aerobic process . Biotransformations . Bioremediation . Waste treatment

1 Introduction Among heavy-metal pollutants, chromium (Cr) is considered to be toxic and one of the main pollutants (Yewalkar et al. 2007). Chromium is widely used in industrial operations such as leather tanning, pigment production, electroplating, paints, steel manufacture, and automobile production (Wang and Xiao 1995; Pattanapipitpaisal et al. 2001). Intensive industrial applications of chromium and releases of associated waste have caused substantial soil contamination. Chromium exists in several oxidation states, from −2 to 6 (Jacobs and Testa 2005). However, in the environment, the most stable and common forms are the trivalent [Cr(III)] and hexavalent [Cr(VI)] species (Fendorf 1995). The Cr(VI) form is more reactive and harmful than the trivalent (Francisco et al. 2002) which is, in comparison, less toxic, less soluble, and less mobile than the hexavalent form (Stanin 2005). In recent years, more and more attention has been focused on the bioremediation of Cr(VI) contamination with chromate-resistant bacteria (Zhu et al. 2008). In this work, plating method was used to isolate Cr(VI)-removing bacteria from chromiumcontaminated soil samples adjacent to a chromium landfill in Changsha, Hunan province, China. The growth conditions, such as pH, temperature, and the initial

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A number of previous studies have shown that many microorganisms possess Cr(VI) tolerance/resistance. As outlined in Table 1, P. saccharolyticum strain W1, Oceanobacillus sp. W4, and S. saromensis W5 had higher Cr(VI)-resistance ability than most other previously identified strains, which highlight the potential of these three isolates as bioremediators of Cr(VI) from chromium-polluted areas. Pseudochrobactrum was recently proposed by Kämpfer et al. (2006) and comprises five species to date: P. saccharolyticum (Kampfer et al. 2006), Pseudochrobactrum asaccharolyticum (Kampfer et al. 2006), Pseudochrobactrum kiredjianiae (Kampfer et al. 2007), Pseudochrobactrum lubricantis (Kampfer et al. 2009), and Pseudochrobactrum glaciei (Romanenko et al. 2008). Pseudochrobactrum sp. (Kampfer et al. 2006; Kampfer et al. 2007; Kampfer et al. 2009; Romanenko et al. 2008) was able to grow at 15–40 °C (optimum temperature was 25–30 °C) and the optimum pH value was about 7.1–7.5. In this study, the optimum temperature for P. saccharolyticum strain W1 was 35 °C and it could grow at a pH range of 8.0–11.0, with optimum pH value of 9.5. To our knowledge, this is the first report of chromium resistance by a strain from Pseudochrobactrum sp. Oceanobacillus sp. was initially reported to grow at pH 9–10 and at 15–40 °C with the optimum temperature at 30–36 °C and was identified as a halotolerant obligate alkaliphile isolated from the skin of a rainbow trout (Yumoto et al. 2005). Molokwane et al. (2008) also reported that a mixed culture of bacteria, containing Oceanobacillus sp., could remove Cr(VI) under anaerobic condition. In this study, Oceanobacillus sp. W4 was observed to have the ability to remove chromate under aerobic conditions and could grow at a wider range of temperature and pH value than those previously reported. As described by Ilhan et al. (2004), the optimum temperature for chromium removal by a strain of S. saprophyticus was 27 °C and the optimum pH value was found to be at 2.0. Mistry et al. (2010) reported the optimum pH value was 7.0 for the chromate removal of a strain of S. saprophyticus. In this study, S. saprophyticus W2 could grow at a pH range of 8.0– 10.0 even though the 16S rRNA gene sequences share 99 % identity, which may indicate that strain W2 and the strain of S. saprophyticus described in Ilhan et al. (2004) may belong to different subspecies. S. saromensis W5 had the highest resistance to chromium among the five isolates, as it reached the maximum cell density at

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Cr(VI) concentration of 1,050 mg L−1. In very limited previous studies, Sporosarcina sp. have been found capable of growing in Cr(VI) concentration of 5 ppm (5 mg L−1) (Fein et al. 2002) and tolerate 2,900 μM (=150.8 mg L − 1 ) of Cr (Bafana 2011), while S. saromensis W5 could tolerate Cr(VI) concentration of up to 1, 400 mg L−1. Most previous studies have found that Cr(VI) inhibits bacterial growth at any Cr(VI) concentration (He et al. 2009; Middleton et al. 2003; Camargo et al. 2003). It has been reported that the growth of Arthrobacter sp. and Bacillus sp. was stimulated by Cr(VI) concentrations of 50 and 5 mg L−1, respectively (Megharaj et al. 2003). In this study, growth of the five isolates was promoted by Cr(VI) when under certain Cr(VI) concentrations. P. saccharolyticum strain W1 and S. saromensisW5 were stimulated by Cr(VI) concentration from 0 to 875 mg L−1 and from 175 to 1,400 mg L−1, respectively. This mechanism was not explicit and need to be further studied. 3.3 Cr(VI) Removal by the Five Isolates The ability of the five isolates to remove Cr(VI) when incubated in their respective optimum initial Cr(VI) concentrations for growth was studied. Abiotic removal of Cr(VI) has also been evaluated and was found to be neglectable. The changes in Cr(VI) concentration in the medium with time for the five isolates were shown in Fig. 3. A decrease of the chromium concentration in a solution was observed with time, with the maximum change generally observed within the initial 15 h for all isolates. Among the five isolates, S. saromensis W5 had the highest chromium-removal rate of 0.79 mg h−1 mg−1 biomass, as the Cr(VI) concentration decreased from 1,050 to 165 mg L−1 within the first 15 h. The concentration of Cr(VI) decreased from 350 to 111.76 mg L−1 (a rate of 0.21 mg h−1 mg−1 biomass) in incubations with P. saccharolyticum strain W1 within the first 15 h. The removal rate of Cr(VI) by S. saprophyticus strain W2 and Lysinibacillus sp. strain W3 was both about 0.07 mg h−1 mg−1 biomass within 15 h at the initial Cr(VI) concentration of 175 mg L−1. According to Fein et al. (2002), a strain of Sporosarc inaureae was capable of removing Cr from a medium containing 5 mg L−1 Cr(VI); calculation based on data got in the first 20 h showed a removal rate of 1.15×10−5 mg h−1 mg−1 biomass. S. saromensis W5 had much greater ability to remove Cr(VI) when compared with that. As described by Ilhan et al. (2004), the Cr(VI) bio-removing

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removal and EPS production was investigated, and the results are shown in Table 2. The EPS content of the five isolates ranged from 68.03 to 189.19 mg in 1 g of dried cell material. P. saccharolyticum strain W1 and S. saromensis W5, which had higher chromiumremoving ability as compared with the other isolates, produced much more EPS, with 178.60 and 189.19 mg g−1, respectively. S. saprophyticus strain W2 (82.15 mg g−1), Lysinibacillus sp. strain W3 (68.03 mg g − 1 ), and Oceanobacillus sp. W4 (140.40 mg g −1 ) had lower EPS contents which corresponded to a relatively low chromium-removing ability. The positive correlation is consistent with the report by Ozturk and Aslim (2008). As seen in Table 2, the contents of protein and carbohydrate of bacterial EPS were analyzed. The EPS of P. saccharolyticum strain W1 and S. saromensis W5 had the highest carbohydrate contents when compared with the rest of the isolates. Chromium was also found in the EPS with contents ranged from 20.7 to 5.90 mg g−1, which was similar to the reports of Priester et al. (2006), FreireNordi et al. (2005), and Kiran and Kaushik (2008). Additionally, it was found that more chromium was removed when more EPS was produced, suggesting that the bacterial EPS did contribute to the Cr(VI) removal.

4 Conclusion Five Cr(VI)-removing bacterial strains were successfully isolated from the Cr(VI)-contaminated soil samples by using plating method, and these strains were found to grow at a wide range of initial Cr(VI) concentration. Results showed that both P. saccharolyticum strain W1 and S. saromensis W5 had high chromium-removing rates of up to 0.21 and 0.79 mg h−1 mg−1 biomass, respectively. To our knowledge, this is the first report on chromium removal by Pseudochrobactrum sp. A positive relationship was found between bacterial Cr(VI) removal and EPS production, suggesting that EPS may be important in chromium removal; however, the exact mechanism of how the isolated strains are involved in chromium removal needs further study.

Acknowledgment The authors would like to thank Dr. Allyson Brady at the University of Calgary for her help in improving the paper. This work was financially supported by the National Natural Science Foundation of China (no. 31370053).

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References Badar, U., Ahmed, N., Beswick, A., Pattanapipitpaisal, P., & Macaskie, L. (2000). Reduction of chromate by microorganisms isolated from metal contaminated sites of Karachi, Pakistan. Biotechnology Letters, 22(10), 829–836. Bafana, A. (2011). Mercury resistance in Sporosarcina sp. G3. Biometals, 24(2), 301–309. Camargo, F., Okeke, F., & Frankenberger, B. (2003). Chromate reduction by chromium-resistant bacteria isolated from soils contaminated with dichromate. Journal of Environmental Quality, 32(4), 1228–1233. Camargo, F. A. O., Okeke, B. C., Bento, F. M., & Frankenberger, W. T. (2005). Diversity of chromium-resistant bacteria isolated from soils contaminated with dichromate. Applied Soil Ecology, 29(2), 193–202. Chen, J., Tang, Y.-Q., & Wu, X.-L. (2012). Bacterial community shift in two sectors of a tannery plant and its Cr (VI) removing potential. Geomicrobiology Journal, 29(3), 226–235. Desai, C., Parikh, R. Y., Vaishnav, T., Shouche, Y. S., & Madamwar, D. (2009). Tracking the influence of long-term chromium pollution on soil bacterial community structures by comparative analyses of 16S rRNA gene phylotypes. Research In Microbiology, 160(1), 1–9. Dubois, M., Gilles, K. A., Hamilton, J. K., Rebers, P., & Smith, F. (1956). Colorimetric method for determination of sugars and related substances. Analytical Chemistry, 28(3), 350–356. Fein, J. B., Fowle, D. A., Cahill, J., Kemner, K., Boyanov, M., & Bunker, B. (2002). Nonmetabolic reduction of Cr (VI) by bacterial surfaces under nutrient-absent conditions. Geomicrobiology Journal, 19(3), 369–382. Fendorf, S. E. (1995). Surface reactions of chromium in soils and waters. Geoderma, 67(1–2), 55–71. Francisco, R., Alpoim, M., & Morais, P. (2002). Diversity of chromium resistant and reducing bacteria in a chromium contaminated activated sludge. Journal of Applied Microbiology, 92(5), 837–843. Freire-Nordi, C. S., Vieira, A. A. H., & Nascimento, O. R. (2005). The metal binding capacity of Anabaena spiroides extracellular polysaccharide: an EPR study. Process Biochemistry, 40(6), 2215–2224. Gehrke, T., Telegdi, J., Thierry, D., & Sand, W. (1998). Importance of extracellular polymeric substances from Thiobacillus ferrooxidans for bioleaching. Applied And Environmental Microbiology, 64(7), 2743–2747. He, Z., Gao, F., Sha, T., Hu, Y., & He, C. (2009). Isolation and characterization of a Cr (VI)-reduction Ochrobactrum sp. strain CSCr-3 from chromium landfill. Journal of Hazardous Materials, 163(2–3), 869–873. Ilhan, S., Nurbas, M., Kiliarslan, S., & Ozdag, H. (2004). Removal of chromium, lead and copper ions from industrial waste waters by Staphylococcus saprophyticus. Turkish Electronic Journal Biotechnology, 2, 50–57. Jacobs, J. A., & Testa, S. M. (2005). Overview of chromium (VI) in the environment: background and history. In J. Guertin, J. A. Jacobs, & C. P. Avakian (Eds.), Chromium (VI) Handbook (pp. 1–21). Boca Raton: CRC Press. Kampfer, P., Rossello-Mora, R., Scholz, H. C., Welinder-Olsson, C., Falsen, E., & Busse, H. J. (2006). Description of

REVIEW ARTICLE

Molecular mechanisms of Cr(VI) resistance in bacteria and fungi Carlo Viti, Emmanuela Marchi, Francesca Decorosi & Luciana Giovannetti Dipartimento di Scienze delle Produzioni Agroalimentari e dell’Ambiente – sezione di Microbiologia, Universit
a degli Studi di Firenze, Florence, Italy

Correspondence: Carlo Viti, Piazzale delle Cascine 24, 50144 Florence, Italy. Tel.: +39 0553288307; fax: +39 0553288272; e-mail: [email protected] Received 7 May 2013; revised 13 September 2013; accepted 28 October 2013. Final version published online 3 December 2013. DOI: 10.1111/1574-6976.12051 Editor: Bernardo Gonz alez

MICROBIOLOGY REVIEWS

Keywords Cr(VI) toxicity; chromate; dichromate; genomics; proteomics; transcriptomics.

Abstract Hexavalent chromium [Cr(VI)] contamination is one of the main problems of environmental protection because the Cr(VI) is a hazard to human health. The Cr(VI) form is highly toxic, mutagenic, and carcinogenic, and it spreads widely beyond the site of initial contamination because of its mobility. Cr(VI), crossing the cellular membrane via the sulfate uptake pathway, generates active intermediates Cr(V) and/or Cr(IV), free radicals, and Cr(III) as the final product. Cr(III) affects DNA replication, causes mutagenesis, and alters the structure and activity of enzymes, reacting with their carboxyl and thiol groups. To persist in Cr(VI)-contaminated environments, microorganisms must have efficient systems to neutralize the negative effects of this form of chromium. The systems involve detoxification or repair strategies such as Cr(VI) efflux pumps, Cr(VI) reduction to Cr(III), and activation of enzymes involved in the ROS detoxifying processes, repair of DNA lesions, sulfur metabolism, and iron homeostasis. This review provides an overview of the processes involved in bacterial and fungal Cr(VI) resistance that have been identified through ‘omics’ studies. A comparative analysis of the described molecular mechanisms is offered and compared with the cellular evidences obtained using classical microbiological approaches.

Introduction Chromium, which belongs to the group VI-B transition metals of the periodic table, has an atomic number of 24, is the most abundant heavy metal, together with zinc, in the lithosphere (69 lg g 1; Li, 2000) and the 21st most abundant element in the Earth’s crust (ranging from 100 to 300 lg g 1; Cervantes et al., 2001). This metal is introduced into the environment from natural sources such as volcanic eruptions, forest fires, and weathering, but the largest contribution to the deposition of chromium in the biosphere is the result of anthropogenic activities. Chromium, due to its hardness, sheen, high melting point, odorlessness, and anti-corrosiveness, is utilized in various industrial activities, including electroplating, steel, and automobile manufacturing, wood treatment, leather tanning, pigments in dyes, paints, inks, plastics, and military defense applications (Lang ard, 1980; James, 1996; Viti & Giovannetti, 2007). Chromium exists in different oxidation states but its two most stable oxidation forms in the environment are FEMS Microbiol Rev 38 (2014) 633–659

the hexavalent [Cr(VI)] and trivalent [Cr(III)] forms (Bartlett, 1991; Zayed & Terry, 2003). These oxidation states have different chemical features and affect organisms in different ways. Cr(III) is conventionally considered an essential micro-nutrient in the diet of animals and humans. Nevertheless, it has been reported recently that chromium can no longer be considered an essential element because rats on a diet with low-Cr(III) suffered no adverse consequences to body composition, glucose metabolism or insulin sensitivity compared with rats on a diet with a sufficient dose of Cr(III) (Di Bona et al., 2011). On the other hand, a high dose (supra-nutritional level) of Cr(III) in the diet improved insulin sensitivity (Di Bona et al., 2011). Cr(III) complexes accumulating in the body are potentially genotoxic (Levina et al., 2003) and therefore their use in micro-nutrient or antidiabetic treatments should be reconsidered after the accurate analysis of available and/or emerging data (Levina & Lay, 2008; Di Bona et al., 2011). Cr(III) is relatively insoluble under environmental conditions (Bartlett & Kimble, 1976; Sass & Rai, 1987) and is considered less toxic than Cr(VI) ª 2013 Federation of European Microbiological Societies. Published by John Wiley & Sons Ltd. All rights reserved

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The reaction mechanism of YieF is different from that described for ChrR of P. putida and involves an obligatory four-electron reduction of chromate in which the enzyme simultaneously transfers three electrons to chromate to produce Cr(III) and one electron to molecular oxygen, generating ROS (Ackerley et al., 2004b) (Fig. 3). This reaction mechanism generates less ROS than that described for ChrR of P. putida, based on the combination of two- and one-electron reduction, thus YieF should be a more suitable enzyme for chromate detoxification than the P. putida ChrR (Ramirez-Diaz et al., 2008). Although a large number of studies have demonstrated the role of ChrR in Cr(VI) reduction, proteomics revealed that this protein in P. putida F1, possessing a ChrR with a 100% amino acid identity to that of P. putida KT2404 (Barak et al., 2006), was down-regulated in response to acute chromate exposure in all conditions tested (Thompson et al., 2010). On the other hand, temporal genomic and proteomic studies of S. oneidensis MR-1 indicated that a NADPH-dependent FMN reductase [SO3585, incorrectly annotated as putative azoreductase (Mugerfeld et al., 2009)], sharing approximately 28% of identity with ChrR of P. putida, was significantly upregulated in Cr(VI)-exposed cells (Brown et al., 2006; Thompson et al., 2007), especially at the highest chromate doses used (Thompson et al., 2007). The deletion of the so3585 gene was not critical for cell survival in the presence of chromate, and only an initial decrease of Cr (VI) reduction rate was observed (Mugerfeld et al., 2009) and therefore more studies are need to understand the role of so3585 in chromate resistance. Microbial respiration with Cr(VI) as the terminal electron acceptor has never been rigorously shown (Richter et al., 2012). Nevertheless, the global transcriptomic analysis of S. oneidensis MR-1, treated with 100 lM Cr (VI) as the sole electron acceptor, revealed the up-regulation of genes encoding MtrA, MtrB, MtrC, and OmcA (Bencheikh-Latmani et al., 2005), which are involved in the dissimilatory extracellular reduction of solid ferric iron [Fe(III)] (hydr)oxides, uranium [U(VI)] and technetium [Tc(VII)] (Belchik et al., 2011). The cytochromes MtrC and OmcA of S. oneidensis MR-1 were deeply characterized to understand their role in Cr(VI) reduction. The data obtained supported the idea that MtrC and OmcA are the terminal reductases of Cr(VI) in S. oneidensis MR-1 (Belchik et al., 2011). Chromate reduction has been also associated with biosorption. Fein et al. (2002) showed nonmetabolic reduction of Cr(VI) to Cr(III) by bacterial surfaces under nonutrient conditions as probable results of the oxidation of organic molecules within the cell wall that serve as electron donors for Cr(VI) reduction to Cr(III). Nancharaiah et al. (2010), studying Cr(VI) reduction by aerobically ª 2013 Federation of European Microbiological Societies. Published by John Wiley & Sons Ltd. All rights reserved

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grown granular bacterial biofilms, found that there was not reduction of Cr(VI) to Cr(III) under nonutrient conditions, whereas they efficiently reduced Cr(VI) from minimal media in the presence of acetate. In both studies, X-ray absorption near-edge structure (XANES) spectroscopy and extended X-ray adsorption fine structure (EXAFS) were used, demonstrating that these approaches produce useful information about the speciation and association of the Cr immobilized on microbial biomass. Very little is known about the mechanisms mediating Cr(VI) reduction in fungi. However, fungi have the ability to reduce Cr(VI), and many studies have been performed to exploit this capability for environmental bioremediation. Filamentous fungi, such as Aspergillus sp., Penicillium sp., and Trichoderma inhamatum, reduce Cr (VI) to Cr(III) by exploiting the reducing power generated by carbon metabolism as mechanism of Cr (VI) detoxification (Acevedo-Aguilar et al., 2006; Morales-Barrera & Cristiani-Urbina, 2008). Paecilomyces lilacinus has demonstrated the ability to both biotransform Cr (VI) and accumulate it in the biomass, exerting the maximum reduction activity during the log phase of growth, when cellular metabolic activity is maximized, and maximum accumulation during the stationary phase (Sharma & Adholeya, 2011). Aspergillus niger strains have been described as coping with chromium mainly via the biosorption of the metal into the cells, rather than via the use of reducing activity (Sandana Mala et al., 2006). The Ed8 strain of Aspergillus tubingensis, included in the A. niger species complex, demonstrated the ability to decrease Cr(VI) concentration in the medium via a reduction mechanism stimulated by carboxylic acids and metal-chelating agents (Coreno-Alonso et al., 2009). Extracellular reduction of Cr(VI) to Cr(III) was observed during the growth of Candida utilis by mechanisms independent from the intensity of culture growth or initial chromium concentration (Muter et al., 2001). On the basis of their results Muter et al. (2001) hypothesized that Cr(VI) reduction in C. utilis could be partly dependent on pH changes of broth during the exponential phase or on exo-enzymatic activities during stationary phase. Candida maltosa, isolated from tanning liquors from a leather factory and characterized by a high tolerance level of chromate in comparison with the yeast laboratory strains C. albicans, S. cerevisiae, and Yarrowia lipolytica, demonstrated the ability to reduce Cr(VI) both in the presence of viable intact cells and in cell-free extracts (Ramirez-Ramirez et al., 2004). This ability was related to NADH-dependent chromate reductase activity associated with soluble proteins and, to a lesser extent, with the membrane fraction (Ramirez-Ramirez et al., 2004). Recently, the reduction of Cr(VI) to Cr(III) through an enzymatic mechanism has been observed in FEMS Microbiol Rev 38 (2014) 633–659

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mutants of Schizosaccharomyces pombe. FEMS Microbiol Lett 178: 109–115. Damron FH & Yu HWD (2011) Pseudomonas aeruginosa MucD regulates the alginate pathway through activation of MucA degradation via MucP proteolytic activity. J Bacteriol 193: 286–291. Das SK, Mukherjee M & Guha AK (2008) Interaction of chromium with resistant strain Aspergillus versicolor: investigation with atomic force microscopy and other physical studies. Langmuir 24: 8643–8650. De Flora S & Wetterhahn KE (1989) Mechanisms of chromium metabolism and genotoxicity. Life Chem Rep 7: 169–244. Decorosi F (2010) Studio di ceppi batterici per il biorisanamento di suoli contaminati da Cr(VI). Firenze University Press, Florence. Decorosi F, Tatti E, Mini A, Giovannetti L & Viti C (2009) Characterization of two genes involved in chromate resistance in a Cr(VI)-hyper-resistant bacterium. Extremophiles 13: 917–923. Decorosi F, Lori L, Santopolo L, Tatti E, Giovannetti L & Viti C (2011) Characterization of a Cr(VI)-sensitive Pseudomonas corrugata 28 mutant impaired in a pyridine nucleotide transhydrogenase gene. Res Microbiol 162: 747– 755. Di Bona KR, Love S, Rhodes NR et al. (2011) Chromium is not an essential trace element for mammals: effects of a ‘low-chromium’ diet. J Biol Inorg Chem 16: 381–390. Diaz-Perez C, Cervantes C, Campos-Garcia J, Julian-Sanchez A & Riveros-Rosas H (2007) Phylogenetic analysis of the chromate ion transporter (CHR) superfamily. FEBS J 274: 6215–6227. Dillon CT, Lay PA, Cholewa M, Legge GJF, Bonin AM, Collins TJ, Kostka KL & SheaMcCarthy G (1997) Microprobe X-ray absorption spectroscopic determination of the oxidation state of intracellular chromium following exposure of V79 Chinese hamster lung cells to genotoxic chromium complexes. Chem Res Toxicol 10: 533–535. Eleutherio ECA, Ribeiro MDS, Pereira MD, Maia FM & Panek AD (1995) Effect of trehalose during stress in a heat-shock resistant mutant of Saccharomyces cerevisiae. Biochem Mol Biol Int 36: 1217–1223. Fahey RC (2013) Glutathione analogs in prokaryotes. Biochim Biophys Acta 1830: 3182–3198. Faisal M & Hasnain S (2004) Comparative study of Cr(VI) uptake and reduction in industrial effluent by Ochrobactrum intermedium and Brevibacterium sp. Biotechnol Lett 26: 1623–1628. Fauchon M, Lagniel G, Aude JC, Lombardia L, Soularue P, Petat C, Marguerie G, Sentenac A, Werner M & Labarre J (2002) Sulfur sparing in the yeast proteome in response to sulfur demand. Mol Cell 9: 713–723. Fedorovych D, Kszeminska H, Babjak L, Kaszycki P & Koloczek H (2001) Hexavalent chromium stimulation of riboflavin synthesis in flavinogenic yeast. Biometals 14: 23–31. Fein JB, Fowle DA, Cahill J, Kemner K, Boyanov M & Bunker B (2002) Nonmetabolic reduction of Cr(VI) by bacterial

ª 2013 Federation of European Microbiological Societies. Published by John Wiley & Sons Ltd. All rights reserved

C. Viti et al.

surfaces under nutrient-absent conditions. Geomicrobiol J 19: 369–382. Felice MR, De Domenico I, Li LT, Ward DM, Bartok B, Musci G & Kaplan J (2005) Post-transcriptional regulation of the yeast high affinity iron transport system. J Biol Chem 280: 22181–22190. Flores-Alvarez LJ, Corrales-Escobosa AR, Cortes-Penagos C, Martinez-Pacheco M, Wrobel-Zasada K, Wrobel-Kaczmarczyk K, Cervantes C & Gutierrez-Corona F (2012) The Neurospora crassa chr-1 gene is up-regulated by chromate and its encoded CHR-1 protein causes chromate sensitivity and chromium accumulation. Curr Genet 58: 281–290. Francisco R, Alpoim MC & Morais PV (2002) Diversity of chromium-resistant and -reducing bacteria in a chromiumcontaminated activated sludge. J Appl Microbiol 92: 837–843. Frederick TM, Taylor EA, Willis JL, Shultz MS & Woodruff PJ (2013) Chromate reduction is expedited by bacteria engineered to produce the compatible solute trehalose. Biotechnol Lett 35: 1291–1296. Fucci L, Oliver CN, Coon MJ & Stadtman ER (1983) Inactivation of key metabolic enzymes by mixed-function oxidation reactions: possible implication in protein turnover and ageing. P Natl Acad Sci USA 80: 1521–1525. Fujs S, Ekert M, Scancar J & Raspor P (2007) Induced cross-protection responses against Cr(III) and Fe(III) ions in Saccharomyces cerevisiae. J Basic Microbiol 47: 301–308. Gancedo C & Flores CL (2004) The importance of a functional trehalose biosynthetic pathway for the life of yeasts and fungi. FEMS Yeast Res 4: 351–359. Gasch AP & Wemer-Washburne M (2002) The genomics of yeast responses to environmental stress and starvation. Funct Integr Genomics 2: 181–192. Gazdag Z, Pocsi I, Belagyi J, Emri T, Blasko A, Takacs K & Pesti M (2003) Chromate tolerance caused by reduced hydroxyl radical production and decreased glutathione reductase activity in Schizosaccharomyces pombe. J Basic Microbiol 43: 96–103. Gharieb MM & Gadd GM (1998) Evidence for the involvement of vacuolar activity in metal(loid) tolerance: vacuolar-lacking and -defective mutants of Saccharomyces cerevisiae display higher sensitivity to chromate, tellurite and selenite. Biometals 11: 101–106. Gharieb MM & Gadd GM (2004) Role of glutathione in detoxification of metal(loid)s by Saccharomyces cerevisiae. Biometals 17: 183–188. Giaever G, Shoemaker DD, Jones TW, Liang H, Winzeler EA, Astromoff A & Davis RW (1999) Genomic profiling of drug sensitivities via induced haploinsufficiency. Nat Genet 21: 278–283. Gibb HJ, Lees PSJ, Pinsky PF & Rooney BC (2000) Lung cancer among workers in chromium chemical production. Am J Ind Med 38: 115–126. Gitan RS, Luo H, Rodgers J, Broderius M & Eide D (1998) Zinc-induced inactivation of the yeast ZRT1 zinc transporter occurs through endocytosis and vacuolar degradation. J Biol Chem 273: 28617–28624.

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Article pubs.acs.org/JPCC

Interaction of Cr(III) and Cr(VI) with Hematite Studied by Second Harmonic Generation Julianne M. Troiano, David S. Jordan, Christopher J. Hull, and Franz M. Geiger* Department of Chemistry, Northwestern University, 2145 Sheridan Road, Evanston, Illinois 60208, United States S Supporting Information *

ABSTRACT: The fate of chromium in the environment relies heavily on its redox chemistry and interaction with iron oxide surfaces. Atomic layer deposition was used to deposit a 10 nm film of polycrystalline α-Fe2O3 (hematite) onto a fused silica substrate which was analyzed using second harmonic generation (SHG), a coherent, surface-specific, nonlinear optical technique. Specifically, the χ(3) technique was used to investigate the adsorption of Cr(III) and Cr(VI) to the hematite/ water interface under flow conditions at pH 4 with 10 mM NaCl. We observed partially irreversible adsorption of Cr(III), the extent of which was found to be dependent on the concentration of Cr(III) ions in solution. This result was confirmed using X-ray photoelectron spectroscopy. The interaction of Cr(III) with hematite is compared with the adsorption of Cr(III) to the silica/water interface, which is the substrate for the ALD-prepared hematite films, and found to be fully reversible under the same experimental conditions. The observed binding constant for Cr(III) interacting with the silica surface was found to be 4.0(6) × 103 M−1, which corresponds to an adsorption free energy of −30.5(4) kJ/mol when referenced to 55.5 M water. The surface charge density at maximum metal ion surface coverage was found to be 0.005(1) C/m2, which corresponds to 1.0 × 1012 ions/cm2 assuming a +3 charge for chromium. In contrast, the observed binding constant for Cr(III) interacting reversibly with the hematite surface was calculated to be 2(2) × 104 M−1, corresponding to an adsorption free energy of −35(2) kJ/mol when referenced to 55.5 M water. The surface charge density at maximum metal ion surface coverage was found to be 0.004(5) C/m2 for the reversibly bound chromium species, which corresponds to 8.3 × 1011 reversibly bound ions per cm2, again assuming a +3 charge of chromium. The data also allows us to estimate that about 6.7 × 1012 Cr(III) ions are irreversibly bound per cm2 hematite at saturation coverage. The results of this investigation suggest that the use of hematite in permeable reactive barriers, for cost-effective chromium remediation, allows for Cr(III) remediation at very low concentrations through adsorptive and redox processes but quickly renders the barriers ineffective at high chromium concentrations due to surface saturation. barriers,29 which is often a less expensive and more effective alternative to pumping treatments or bioremediation;30−33 however, those processes are currently limited by surface passivation of the iron-bearing solid phase.34 Reactive barriers used for remediating chromium contain iron-bearing materials such as magnetite,35−37 goethite,38 and iron sulfides,20 as well as zerovalent iron20,39,40 that are placed in the path of flowing groundwater.41−43 Contaminants are then removed at the interface between the reactive barrier material and the chromium-containing aqueous phase by in situ transformations, including sorption and redox chemistry. Specifically, the reduction of Cr(VI) by Fe(II) solutions, iron-bearing minerals, or zerovalent iron allows Cr(VI) to be removed from groundwater as the less mobile, less toxic Cr(III). Once generated, aqueous Cr(III) commonly adsorbs to iron oxides and oxyhydroxides,15,44 forming strongly bound inner-sphere complexes and/or precipitates45 that effectively remove Cr(III) from the aqueous phase. While chromium remediation via permeable reactive barriers is widely used, a significant problem

I. INTRODUCTION Chromium is a common contaminant in groundwater and is released into the environment primarily through industrial activities,1−6 such as through its use in wood preservatives, refractory bricks, leather tanning processes, chromate plating, and steel manufacturing. Chromium exists in the environment in two stable oxidation states, namely Cr(III) and Cr(VI).7 Hexavalent chromium, CrO42−, is a highly toxic heavy metal ion species and a known carcinogen1,2,7−10 which is highly mobile11−14 in the environment. On the other hand, Cr(III), the reduced form of Cr(VI), is much less mobile in the environment due to its propensity to form insoluble oxyhydroxides and adsorb to mineral surfaces.7,15,16 Cr(III) is also an essential nutrient for humans and animals.1,2 Given that redox chemistry in the environment allows these two oxidation states to easily interconvert, they are both considered potentially carcinogenic,1,17,18 placing chromium on the short list of EPA priority pollutants.19 Because Cr(III) is less mobile and less toxic than Cr(VI), many Cr(VI) remediation techniques involve its reduction to Cr(III) through the use of readily available, inexpensive, and strong reducing agents, such as iron.17,20−28 One way to remediate redox-active contaminants is the use of reactive © 2013 American Chemical Society

Received: December 13, 2012 Revised: February 14, 2013 Published: February 18, 2013 5164

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The Journal of Physical Chemistry C

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(32) Wielinga, B.; Mizuba, M. M.; Hansel, C. M.; Fendorf, S. Environ. Sci. Technol. 2001, 35, 522−527. (33) Fein, J. B.; Fowle, D. A.; Cahill, J.; Kemner, K.; Boyanov, M.; Bunker, B. C. Geomicrobiol. J. 2002, 19, 369−82. (34) Mackay, D. M.; Cherry, J. A. Environ. Sci. Technol. 1989, 23, 630−6. (35) Kendelewicz, T.; Liu, P.; Doyle, C. S.; Brown, G. E.; Nelson, E. J.; Chambers, S. A. Surf. Sci. 1999, 424, 219−31. (36) Eary, L. E.; Rai, D. Am. J. Sci. 1989, 289, 180−213. (37) Ilton, E. S.; Veblen, D. R. Geochim. Cosmochim. Acta 1994, 58, 2777−88. (38) Bidoglio, G.; Gibson, P. N.; Ogorman, M.; Roberts, K. J. Geochim. Cosmochim. Acta 1993, 57, 2389−94. (39) Pratt, A. R.; Blowes, D. W.; Ptacek, C. J. Environ. Sci. Technol. 1997, 31, 2492−8. (40) Powell, R. M.; Puls, R. W.; Hightower, S. K.; Sabatini, D. A. Environ. Sci. Technol. 1995, 29, 1913−22. (41) Artiole, J.; Fuller, W. H. J. Environ. Qual. 1979, 8, 503−10. (42) Morrison, S. J.; Spangler, R. R. Environ. Prog. 1993, 12, 175−81. (43) Robertson, W. D.; Cherry, J. A. Ground Water 1995, 33, 99− 111. (44) Dzombak, D. A.; Morel, F. M. M. Surface Complexation Modeling: Hydrous Ferric Oxide; Wiley-Interscience: New York, 1990. (45) Charlet, L.; Manceau, A. J. Colloid Interface Sci. 1992, 148, 443− 58. (46) Melitas, N.; Chuffe-Moscoso, O.; Farrell, J. Environ. Sci. Technol. 2001, 35, 3948−53. (47) Williams, A. G. B.; Scherer, M. M. Environ. Sci. Technol. 2001, 35, 3488−94. (48) Manning B. A.; Kiser J.; Kwon, H. Abstr. Pap. Am. Chem. Soc. 2006, 231. (49) Jordan, D. S.; Hull, C. J.; Troiano, J. M.; Riha, S. C.; Martinson, A. B. F.; et al. J. Phys. Chem. C 2013. (50) Martinson, A. B. F.; DeVries, M. J.; Libera, J. A.; Christensen, S. T.; Hupp, J. T.; et al. J. Phys. Chem. C 2011, 115, 4333−9. (51) Hayes, P. L.; Malin, J. N.; Konek, C. T.; Geiger, F. M. J. Phys. Chem. A 2008, 112, 660−8. (52) Malin, J. N.; Hayes, P. L.; Geiger, F. M. J. Phys. Chem. C 2008, 113, 2041−52. (53) Mifflin, A. L.; Musorrafiti, M. J.; Konek, C. T.; Geiger, F. M. J. Phys. Chem. B 2005, 109, 24386−90. (54) Al-Abadleh, H. A.; Mifflin, A. L.; Bertin, P. A.; Nguyen, S. T.; Geiger, F. M. J. Phys. Chem. B 2005, 109, 9691−702. (55) Al-Abadleh, H. A.; Mifflin, A. L.; Musorrafiti, M. J.; Geiger, F. M. J. Phys. Chem. B 2005, 109, 16852−9. (56) Hayes, P. L.; Gibbs-Davis, J. M.; Musorrafiti, M. J.; Mifflin, A. L.; Scheidt, K. A.; Geiger, F. M. J. Phys. Chem. C 2007, 111, 8796−804. (57) Hayes, P. L.; Chen, E. H.; Achtyl, J. A.; Geiger, F. M. J. Phys. Chem. A 2009, 113, 4269−80. (58) Konek, C. T.; Illg, K. D.; Al-Abadleh, H. A.; Voges, A. B.; Yin, G.; et al. J. Am. Chem. Soc. 2005, 127, 15771−7. (59) Hayes, P. L.; Gibbs-Davis, J. M.; Musorrafiti, M. J.; Mifflin, A. L.; Scheidt, K. A.; Geiger, F. M. J. Phys. Chem. C 2007, 111, 8796−804. (60) Boyd, R. W. Nonlinear Optics; Academic Press: Burlington, MA, 2008; p xix, 613 pp. (61) Salafsky, J. S.; Eisenthal, K. B. J. Phys. Chem. B 2000, 104, 7752− 5. (62) Chen, E. H.; Saslow, S. A.; Nguyen, S. T.; Geiger, F. M. J. Phys. Chem. C 2012, 116, 19886−92. (63) Xiao, X. D.; Vogel, V.; Shen, Y. R. Chem. Phys. Lett. 1989, 163, 555−9. (64) Zhao, X. L.; Ong, S. W.; Eisenthal, K. B. Chem. Phys. Lett. 1993, 202, 513−20. (65) Morel, F.; Hering, J. G. Principles and Applications of Aquatic Chemistry; Wiley: New York, 1993; p xv, 588 pp. (66) Langmuir, D. Aqueous Environmental Chemistry; Prentice Hall: Upper Saddle River, NJ, 1997; p 373. (67) Brown, G. E.; Henrich, V. E.; Casey, W. H.; Clark, D. L.; Eggleston, C.; et al. Chem. Rev. 1999, 99, 77−174.

(68) Koopal, L. K. Electrochim. Acta 1996, 41, 2293−305. (69) Hayes, K. F.; Papelis, C.; Leckie, J. O. J. Colloid Interface Sci. 1988, 125, 717−26. (70) Jena, K. C.; Covert, P. A.; Hore, D. K. J. Phys. Chem. Lett. 2011, 2, 1056−61. (71) Atkins, P. W.; De Paula, J. Elements of Physical Chemistry; Oxford University Press: New York, 2009; p xx, 578 pp. (72) Holland, J. G.; Jordan, D. S.; Geiger, F. M. J. Phys. Chem. B 2011, 115, 8338−45. (73) Holland, J. G.; Geiger, F. M. J. Phys. Chem. B 2012, 116, 6302− 10. (74) Jordan, D. S.; Saslow, S. A.; Geiger, F. M. J. Phys. Chem. A 2011, 115, 14438−45. (75) Das, M. R.; Bordoloi, D.; Borthakur, P. C.; Mahiuddin, S. Colloids Surf., A 2005, 254, 49−55. (76) Ahmed, S. M.; Maksimov, D. Can. J. Chem. 1968, 46, 3841. (77) Song, Q. Y.; Xu, F.; Tsai, S. C. Int. J. Miner. Process. 1992, 34, 219−29. (78) Wang, Y. M.; Pugh, R. J.; Forssberg, E. Colloids Surf., A 1994, 90, 117−33. (79) Yuhua, W.; Jianwei , R. Int. J. Miner. Process. 2005, 77, 116−122. (80) Kosmulski, M. Chemical Properties of Material Surfaces; Marcel Dekker: New York, 2001; p viii, 753 pp. (81) Joy, A. S.; Watson, D.; Cropton, R. W. G. Trans. Soc. Min. Eng. 1964, 229, 5−7. (82) Zhao, X.; Ong, S.; Eisenthal, K. B. Chem. Phys. Lett. 1993, 202, 513−20. (83) Higgins, S. R.; Stack, A. G.; Knauss, K. G.; Eggleston, C. M.; Jordan, G. Geochem. Soc. 2002, No. 7, 111. (84) Du, Q.; Freysz, E.; Shen, Y. R. Phys. Rev. Lett. 1994, 72, 238−41. (85) Hayes, P. L.; Malin, J. N.; Jordan, D. S.; Geiger, F. M. Chem. Phys. Lett. 2010, 499, 183−92. (86) O’Connor, D. J.; Sexton, B. A.; Smart, R. S. C. Surface Analysis Methods in Materials Science; Springer: Berlin, 2003; p xxvi, 585 pp. (87) Astrup, T.; Stipp, S. L. S.; Christensen, T. H. Environ. Sci. Technol. 2000, 34, 4163−8. (88) Jordan, D. S.; Malin, J. N.; Geiger, F. M. Environ. Sci. Technol. 2010, 44, 5862−7. (89) EPA. An In-Situ Permeable Reactive Barrier for the Treatment of Hexavalent Chromium and Trichloroethylene in Ground Water: Vol. 3 Multicomponent Reactive Transport Modeling, Washington, DC, 1999. (90) Hoch, L. B.; Mack, E. J.; Hydutsky, B. W.; Hershman, J. M.; Skluzacek, J. M.; Mallouk, T. E. Environ. Sci. Technol. 2008, 42, 2600− 5. (91) Ponder, S. M.; Darab, J. G.; Bucher, J.; Caulder, D.; Craig, I.; et al. Chem. Mater. 2001, 13, 479−86. (92) Ponder, S. M.; Darab, J. G.; Mallouk, T. E. Environ. Sci. Technol. 2000, 34, 2564−9.

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Journal of Hazardous Materials 256–257 (2013) 24–32

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Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat

Cr(VI) reduction by a potent novel alkaliphilic halotolerant strain Pseudochrobactrum saccharolyticum LY10 Dongyan Long a , Xianjin Tang a,∗ , Kuan Cai b , Guangcun Chen a,a , Linggui Chen a , Dechao Duan a , Jun Zhu c , Yingxu Chen a a

Institute of Environmental Science and Technology, Zhejiang University, Yuhangtang Road 388, Hangzhou, 310058, Zhejiang, PR China Global Environmental Technology Co., Ltd., Gudun Road 656, Hangzhou, 310058, Zhejiang, PR China c Southern Research Outreach Center, University of Minnesota, 35838 120th Street, Waseca, MN 56093, USA b

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

• A novel P. saccharolyticum strain LY10 • • • •

was isolated from Cr contaminated soil. The alkaliphilic and halotolerant versatilities of the strain were characterized. Strain LY10 could accumulate Cr both extracellularly and intracellularly. XANES confirmed that the chromium immobilized by the cells was in the Cr(III) state. P. saccharolyticum was for the first time reported as the Cr(VI) reducing bacteria.

a r t i c l e

i n f o

Article history: Received 19 November 2012 Received in revised form 21 March 2013 Accepted 15 April 2013 Available online 20 April 2013 Keywords: Bioremediation Cr(VI) reduction Pseudochrobactrum saccharolyticum Alkaliphilic Halotolerant X-ray absorption near-edge structure

a b s t r a c t A novel Cr(VI)-reducing strain, Pseudochrobactrum saccharolyticum LY10, was isolated and characterized for its high Cr(VI)-reducing ability. Strain LY10 had typical characteristics of alkali-tolerance and halotolerance. Kinetic analysis indicated that the maximum reduction rate was achieved under optimum conditions with initial pH 8.3, 20 g L−1 NaCl, 55 mg L−1 Cr(VI), and 1.47 × 109 cells mL−1 of cell concentration. Further mechanism studies verified that the removal of Cr(VI) was mainly achieved by a metabolism-dependent bioreduction process. Strain LY10 accumulated chromium both in and around the cells, with cell walls acting as the major binding sites for chromium. X-ray absorption near-edge structure (XANES) analysis further confirmed that the chromium immobilized by the cells was in the Cr(III) state. In the present study, Pseudochrobactrum saccharolyticum was, for the first time, reported to be a Cr(VI)-reducing bacteria. Results from this research would provide a potential candidate for bioremediation of Cr(VI)-contaminated environments, especially alkaline and saline milieus with Cr(VI) at low-to-mid concentrations. © 2013 Elsevier B.V. All rights reserved.

1. Introduction

Abbreviations: TEM, Transmission electron microscopy; EDS, Energy dispersive X-ray spectroscopy; XANES, X-ray absorption near-edge structure; XAS, X-ray absorption spectroscopy. ∗ Corresponding author. Tel.: +86 571 8898 2013; fax: +86 571 8898 2010. E-mail address: [email protected] (X. Tang). 0304-3894/$ – see front matter © 2013 Elsevier B.V. All rights reserved. http://dx.doi.org/10.1016/j.jhazmat.2013.04.020

Chromium, a priority pollutant in the United States and many other countries, has caused great public concern in recent years because of its wide usage, extensive distribution, and hazardous potential [1,2]. However, the toxicity, solubility and bioavailability of Cr, depend primarily on its chemical form [3]. Although chromium can exist in a variety of valence states, Cr(VI) and Cr(III)

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cell envelopes of strain LY10 treated with a low concentration of Cr(VI) (110 mg L−1 ) (Fig. 7D). The EDS spectrum showed an obvious signal of chromium precipitation (Fig. 7c), and the content of chromium was as high as 15.12%. In order to verify whether chromium accumulated only on the cell wall or throughout the whole cell, the intracellular electron dense region was also submitted to EDS analysis. The spectrum clearly showed the presence of chromium within the cell, although the percentage (8.88%) was much lower than that observed on the cell wall (Fig. 7b). Results indicated that chromium precipitated both in and around the cells, with the cell surface acting as the main binding site. Moreover, neither the cell wall nor the cytosolic region showed a detectable chromium signal in cells grown in the absence of Cr(VI) (Fig. 7a). 3.4. Speciation of cell-associated chromium

Fig. 6. The Cr(VI) reduction percentage (A) and reduction rate (B) under optimal conditions. Experiments were conducted with initial pH 8.3, 20 g L−1 NaCl, 55 mg L−1 Cr(VI) and 1.47 × 109 cells mL−1 of cell concentration.

B). However, when exposed to Cr(VI) from 110 mg L−1 (Fig. 7C and D) to 220 mg L−1 (Fig. 7E and F), the shapes of cells became more and more irregular and a few of them even lost their shapes. Moreover, a cluster of round globules encrusted on the surface of cells treated with 220 mg L−1 Cr(VI) (red circle in Fig. 7F), while no precipitation was observed in the control (Fig. 7B), nor on the

To confirm that P. saccharolyticum LY10 has a strong Cr(VI) reducing capacity rather than biosorption ability, the speciation of cell-associated chromium was analyzed after 96 h of exposure to 220 mg L−1 Cr(VI). As Cr(VI) has a characteristic sharp pre-edge feature, which is absent in the Cr(III) spectrum, the deconvolution of sample XANES spectra with known Cr(VI) and Cr(III) standards is regarded as a relatively straightforward approach for determining the Cr speciation [27]. As shown in Fig. 8, for the Cr(VI)-treated cells, the typical strong pre-edge absorbance of Cr(VI) was absent and a small peak was observed at 6009 eV, which was consistent with results observed for model Cr(OH)3 compounds. These results confirmed that the majority of the Cr(VI) was reduced by P. saccharolyticum LY10, and the chromium immobilized by the cells was in the Cr(III) state.

Fig. 7. TEM images of P. saccharolyticum LY10 cells not treated with Cr(VI) (A, B), incubated with 110 mg L−1 Cr(VI) (C, D) and with 220 mg L−1 Cr(VI) (E, F) for 96 h. Corresponding EDS spectra of untreated bacteria (a) and cells treated with 220 mg L−1 Cr(VI) (b, c) for 96 h.

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Fig. 8. The Cr K-edge XANES spectra of P. saccharolyticum LY10 treated with 220 mg L−1 Cr(VI) for 96 h and the reference compounds, K2 Cr2 O7 and Cr(OH)3 .

4. Discussion Bioreduction of toxic Cr(VI) to stable Cr(III) by microorganisms has been regarded as a promising approach for the remediation of chromium contamination [2,30–32]. Since the first microbial Cr(VI) reducer was discovered by Romanenko and Korenkov [33], the search for Cr(VI)-reducing bacteria has been enthusiastically pursued, and various species have been isolated [34,35]. However, in this study, it is the first time that P. saccharolyticum has been clearly identified to be a Cr(VI) reducing bacteria. P. saccharolyticum LY10 demonstrated a typical characteristic of alkali-tolerance. It could efficiently reduce Cr(VI) under neutralalkaline conditions (pH 7.0–10.7). As Cr(VI) is known to desorb from soil more rapidly at elevated pH levels [36], and Cr(III) is immobile under alkaline pH conditions [35], the Cr(VI) reduction under alkaline conditions would help to reduce the mobility and availability of the Cr ions in the environment [9]. The wide pH adaptability and efficient Cr(VI) reducing ability under neutral-alkaline condition suggested that P. saccharolyticum LY10 could play an important role in the bioremediation of alkaline Cr-polluted sites. In addition, strain LY10 was able to adapt to high NaCl concentrations (1–20 g L−1 ). According to the definition of Margesin et al. [37], microorganisms requiring salt for growth are referred as halophiles, and those able to grow in the absence as well as in the presence of salt, are designated as halotolerant. In the present study, strain LY10 could survive high salinity (60 g L−1 NaCl), and the presence of NaCl appeared to be a requirement for its effective Cr(VI) reducing activity, indicating that strain LY10 had a halotolerant nature [10,11,37]. The high salt tolerance of strain LY10 would enhance its actual performance in bioremediation applications under salty conditions. Kinetic studies demonstrated that the initial stage (t < 72 h) of Cr(VI) bioreduction by strain LY10 could be well described by the first-order kinetic model (Fig 2B, Fig 4B, Fig 5 and Fig 6). The effects of different factors on Cr(VI) bioreduction were analyzed by comparing the first-order rate constants observed under varied conditions (Table S2). Results indicated that Cr(VI) reduction rates varied with experiment conditions, including initial pH, salinity, biomass and Cr(VI) concentration. In addition, a close relationship between microbial growth and Cr(VI) reduction rate was observed. For pH values ranging from 7.0 to 10.7 and NaCl concentrations between 1 g L−1 and 20 g L−1 , strain LY10 grew well and rapid Cr(VI) reduction rate were observed. However, when the cell growth was inhibited at pH 5.5 and the NaCl concentration of 0 g L−1 , the Cr(VI) reduction rate and reduction amount declined significantly. It is well known that the pH has a direct impact on the structure of

nucleic acids [38], and sodium is an essential element for the ionic pumps in halophiles [11,37]. Thus, the presence of disfavored pH and absence of NaCl retarded the cell growth, and further hindered the enzymatic activity for Cr(VI) reduction [11]. The direct positive correlation between cell growth and Cr(VI) reduction indicated that the reducing process was cellular-metabolism dependent [22]. Kinetic analysis indicated that the optimum conditions for bioreduction by strain LY10 were pH 8.3, 20 g L−1 of NaCl, 55 mg L−1 of initial Cr(VI), and 1.47 × 109 cells mL−1 of initial cell concentration (Table S2). Complete Cr(VI) reduction and the maximum reduction rate of (4.45 ± 0.11) × 10−2 h−1 was achieved under optimum conditions. The reduction rate of Cr(VI) removal by P. saccharolyticum LY10 is high compared to (2.52 ± 0.33) × 10−2 h−1 reported for Sphaerotilus natans [29] and 3.50 × 10−3 h−1 reported for B. subtilis [27]. However, different first-order rate constants (k) were reported for Chlorella miniata because of the modifications of kinetic models [39]. A fair comparison of these reduction rate values is cumbersome, because kinetic models for Cr(VI) reduction by different bacteria are not identical, and reduction rates are calculated in varied methods [14,15,29]. Moreover, there are differences in experiment conditions (e.g. pH, Cr(VI) concentration, cell density, and electron-donors/electron-acceptors) [40]. Nonetheless, the specific reduction rates provided here are valuable results. Such rate information will help in optimizing the operation conditions for Cr(VI) bioremediation by P. saccharolyticum LY10. To verify that P. saccharolyticum LY10 has a strong Cr(VI) reductive capacity rather than adsorption ability, the mechanism of Cr(VI) reduction was further studied. Apart from the deleterious effects of toxic Cr(VI) on bacterial cells, which were also reported for A. haemolyticus [41], an obvious precipitation of chromium in and around the P. saccharolyticum LY10 cells was observed by TEM-EDS analysis. Although chromium accumulation on the cell surface has been reported in many Cr(VI) reducing bacteria [18,22,41], quite a few have been reported to have chromium precipitation both within and surrounding cells [17]. For P. saccharolyticum LY10, it could accumulate chromium in the cells as well as on their surfaces, with the cell wall acting as the main binding site. Since most Cr(VI) compounds are highly soluble, and the majority of Cr(III) compounds are relatively insoluble, it is tempting to speculate that the chromium in and around the cells is in a reduced form [most likely Cr(III)] resulting from the Cr(VI) reducing process [17,42]. To further confirm the bioreduction activity of P. saccharolyticum LY10, speciation of cell-associated chromium was assessed with XANES analysis. The bioreduction of Cr(VI) involves electron transfer processes, which can result in a direct metal speciation change [35,43]. In contrast, biosorption of metals is known to be controlled by certain forces, such as the electrostatic force [44] and van der Waals [45]. Therefore, this process has no effect on the speciation of metals [19]. In our study, the absence of the characteristic sharp pre-edge absorbance for Cr(VI) and the appearance of a small peak at 6009 eV confirmed that the speciation of chromium had changed from the initial Cr(VI) to Cr(III) [2,46]. The results clearly indicated that bioreduction process was involved in the Cr(VI) removal activity of P. saccharolyticum LY10. Furthermore, recent studies on biosorption have demonstrated that the adsorption of Cr(VI) decreased with the increased pH, especially for pH levels above 4.0 [44]. However, in this study, the amount of Cr(VI) removed dramatically increased when pH increased from 5.5 to 8.3. On the contrary, great inhibition of Cr(VI) reduction was observed when enzyme activity was suppressed under acidic conditions (pH 5.5). The preference for alkaline environment rather than acidic condition further substantiated that the Cr(VI) removal by P. saccharolyticum LY10 was mainly achieved by an enzyme-mediated bioreduction process, rather than by a surface-related bioadsorption activity [47].

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5. Conclusion In the present study, a potent novel alkaliphilic and halotolerant Cr(VI)-reducing strain, P. saccharolyticum LY10, was isolated from Cr-contaminated soil. This versatile Cr(VI) reducer could effectively function under harsh alkaline and salty conditions. Mechanism studies revealed that Cr(VI) was reduced to Cr(III) by strain LY10, and the chromium immobilized both within and surrounding the cells was in the Cr(III) state. Results from this research will provide a novel biological resource for Cr(VI) bioremediation. The kinetic information will also be useful for the optimal Cr(VI) bioremediation by P. saccharolyticum LY10. Acknowledgements This study was financially supported by the National High-Tech Research and Development Program of China (2009AA063101) and the China Postdoctoral Science Foundation (2011M500103). We are grateful to the Centre of Analysis and Measurement of Zhejiang University for TEM-EDS analysis. We are especially grateful to the BL14W1 beamline of Shanghai Synchrotron Radiation Facility for assistance in Cr K-edge XANES. Appendix A. Supplementary data Supplementary data associated with this article can be found, in the online version, at http://dx.doi.org/10.1016/j.jhazmat.2013. 04.020. References [1] B. Wielinga, M.M. Mizuba, C.M. Hansel, S. Fendorf, Iron promoted reduction of chromate by dissimilatory iron-deducing bacteria, Environ. Sci. Technol. 35 (2001) 522–527. [2] Y. Cheng, F. Yan, F. Huang, W. Chu, D. Pan, Z. Chen, J. Zheng, M. Yu, Z. Lin, Z. Wu, Bioremediation of Cr(VI) and immobilization as Cr(III) by Ochrobactrum anthropi, Environ. Sci. Technol. 44 (2010) 6357–6363. [3] J. Kotas, Z. Stasicka, Chromium occurrence in the environment and methods of its speciation, Environ. Pollut. 107 (2000) 263–283. [4] V. Gomez, M.P. Callao, Chromium determination and speciation since 2000, Trends Anal. Chem. 25 (2006) 1006–1015. [5] C. Desai, K. Jain, D. Madamwar, Evaluation of In vitro Cr(VI) reduction potential in cytosolic extracts of three indigenous Bacillus sp. isolated from Cr(VI) polluted industrial landfill, Bioresour. Technol. 99 (2008) 6059–6069. [6] D.F. Ackerley, Y. Barak, S.V. Lynch, J. Curtin, A. Matin, Effect of chromate stress on Escherichia coli K-12, J. Bacteriol. 188 (2006) 3371–3381. [7] A.N. Mabbett, D. Sanyahumbi, P. Yong, L.E. Macaskie, Biorecovered precious metals from industrial wastes: single-step conversion of a mixed metal liquid waste to a bioinorganic catalyst with environmental application, Environ. Sci. Technol. 40 (2006) 1015–1021. [8] R. Boopathy, Factors limiting bioremediation technologies, Bioresour. Technol. 74 (2000) 63–67. [9] M.S.M. Mangaiyarkarasi, S. Vincent, S. Janarthanan, T.S. Rao, B.V.R. Tata, Bioreduction of Cr(VI) by alkaliphilic Bacillus subtilis and interaction of the membrane groups, Saudi J. Biol. Sci. 18 (2011) 157–167. [10] A.S.S. Ibrahim, M.A. El-Tayeb, Y.B. Elbadawi, A.A. Al-Salamah, Bioreduction of Cr (VI) by potent novel chromate resistant alkaliphilic Bacillus sp. strain KSUCr5 isolated from hypersaline Soda lakes, Afr. J. Biotechnol. 10 (2011) 7207–7218. [11] M.A. Amoozegar, A. Ghasemi, M.R. Razavi, S. Naddaf, Evaluation of hexavalent chromium reduction by chromate-resistant moderately halophile, Nesterenkonia sp. strain MF2, Process Biochem. 42 (2007) 1475–1479. [12] L. Xu, M. Luo, C. Jiang, X. Wei, P. Kong, X. Liang, J. Zhao, L. Yang, H. Liu, In vitro reduction of hexavalent chromium by cytoplasmic fractions of Pannonibacter phragmitetus LSSE-09 under aerobic and anaerobic conditions, Appl. Microbiol. Biotechnol. 166 (2012) 933–941. [13] K.M. Kemner, S.D. Kelly, B. Lai, J. Maser, E.J. O‘Loughlin, D. Sholto-Douglas, Z.H. Cai, M.A. Schneegurt, C.F. Kulpa, K.H. Nealson, Elemental and redox analysis of single bacterial cells by X-ray microbeam analysis, Science 306 (2004) 686–687. [14] C. Liu, Y.A. Gorby, J.M. Zachara, J.K. Fredrickson, C.F. Brown, Reduction kinetics of Fe(III), Co(III), U(VI), Cr(VI), and Tc(VII) in cultures of dissimilatory metalreducing bacteria, Biotechnol. Bioeng. 80 (2002) 637–649. [15] H. Guha, K. Jayachandran, F. Maurrasse, Kinetics of chromium (VI) reduction by a type strain Shewanella alga under different growth conditions, Environ. Pollut. 115 (2001) 209–218.

31

[16] C. Liu, J.M. Zachara, Uncertainties of Monod kinetic parameters nonlinearly estimated from batch experiments, Environ. Sci. Technol. 35 (2001) 133–141. [17] S.S. Middleton, R. Bencheikh-Latmani, M.R. Mackey, M.H. Ellisman, B.M. Tebo, C.S. Criddle, Cometabolism of Cr(VI) by Shewanella oneidensis MR-1 produces cell-associated reduced chromium and inhibits growth, Biotechnol. Bioeng. 83 (2003) 627–637. [18] L. Chai, S. Huang, Z. Yang, B. Peng, Y. Huang, Y. Chen, Hexavalent chromium reduction by Pannonibacter phragmitetus BB isolated from soil under chromium-containing slag heap, J. Environ. Sci. Health Part A: Toxic/Hazard Substances Environ. Eng. 44 (2009) 615–622. [19] Z.A. Zakaria, N. Aruleswaran, S. Kaur, W.A. Ahmad, Biosorption and bioreduction of Cr(VI) by locally isolated Cr-resistant bacteria, Water Sci. Technol. 56 (2007) 117–123. [20] Z. Chen, Z. Huang, Y. Cheng, D. Pan, X. Pan, M. Yu, Z. Pan, Z. Lin, X. Guan, Z. Wu, Cr(VI) uptake mechanism of Bacillus cereus, Chemosphere 87 (2012) 211–216. [21] F.A.O. Camargo, B.C. Okeke, F.M. Bento, W.T. Frankenberger, In vitro reduction of hexavalent chromium by a cell-free extract of Bacillus sp. ES 29 stimulated by Cu2+ , Appl. Microbiol. Biotechnol. 62 (2003) 569–573. [22] J.S. McLean, T.J. Beveridge, D. Phipps, Isolation and characterization of a chromium-reducing bacterium from a chromated copper arsenatecontaminated site, Environ. Microbiol. 2 (2000) 611–619. [23] J.G. Holt, N.R. Krieg, P.H.A. Sneath, J.T. Staley, S.T. Williams, Bergey’s Manual of Determinative Bacteriology, 9th edition, Williams & Wilkins, Baltimore, 1994. [24] J.B. Jones, A.R. Chase, G.K. Harris, Evaluation of the Biolog GN Microplate system for identification of some plant-pathogenic bacteria, Plant Dis. 77 (1993) 553–558. [25] P.F. Urone, Stability of colorimetric reagent for chromium, s-diphenylcarbazide, in various solvents, Anal. Chem. 27 (1955) 1354–1355. [26] X. Chen, S. Hu, C. Shen, C. Dou, J. Shi, Y. Chen, Interaction of Pseudomonas putida CZ1 with clays and ability of the composite to immobilize copper and zinc from solution, Bioresour. Technol. 100 (2009) 330–337. [27] J.B. Fein, D.A. Fowle, J. Cahill, K. Kemner, M. Boyanov, B. Bunker, Nonmetabolic reduction of Cr(VI) by bacterial surfaces under nutrient-absent conditions, Geomicrobiol. J. 19 (2002) 369–382. [28] G. Chen, X. Chen, Y. Yang, A. Hay, X. Yu, Y. Chen, Sorption and distribution of copper in unsaturated Pseudomonas putida CZ1 biofilms as determined by X-ray fluorescence microscopy, Appl. Environ. Microb. 77 (2011) 4719–4727. [29] A.H. Caravelli, L. Giannuzzi, N.E. Zaritzky, Reduction of hexavalent chromium by Sphaerotilus natans a filamentous micro-organism present in activated sludges, J. Hazard. Mater. 156 (2008) 214–222. ˜ [30] E.M.S. Brito, H.A. Pinón-Castillo, R. Guyoneaud, C.A. Caretta, J.F. GutiérrezCorona, R. Duran, G.E. Reyna-López, G.V. Nevárez-Moorillón, A. Fahy, M. ˜ Bacterial biodiversity from anthropogenic extreme environGoni-Urriza, ments: a hyper-alkaline and hyper-saline industrial residue contaminated by chromium and iron, Appl. Microbiol. Biotechnol. 97 (2013) 369–378. [31] D.I. Stewart, I.T. Burke, D.V. Hughes-Berry, R.A. Whittleston, Microbially mediated chromate reduction in soil contaminated by highly alkaline leachate from chromium containing waste, Ecol. Eng. 36 (2010) 211–221. [32] Q. Ye, Y. Roh, S.L. Carroll, B. Blair, J.Z. Zhou, C.L. Zhang, M.W. Fields, Alkaline anaerobic respiration: isolation and characterization of a novel alkaliphilic and metal-reducing bacterium, Appl. Environ. Microb. 70 (2004) 5595–5602. [33] V.I. Romanenko, V.N. Korenkov, Pure Culture of bacteria utilizing chromates and bichromates as hydrogen acceptors in growth under anaerobic conditions, Mikrobiologiia 46 (1977) 414–417. [34] N.K. Kilic, A. Stensballe, D.E. Otzen, G. Donmez, Proteomic changes in response to chromium(VI) toxicity in Pseudomonas aeruginosa, Bioresour. Technol. 101 (2010) 2134–2140. [35] N.M. Dogan, C. Kantar, S. Gulcan, C.J. Dodge, B.C. Yimaz, M.A. Mazmanci, Chromium(VI) bioremoval by pseudomonas bacteria: role of microbial exudates for natural attenuation and biotreatment of Cr(VI) contamination, Environ. Sci. Technol. 45 (2011) 2278–2285. [36] R. Wani, K.M. Kodam, K.R. Gawai, P.K. Dhakephalkar, Chromate reduction by Burkholderia cepacia MCMB-821, isolated from the pristine habitat of alkaline crater lake, Appl. Microbiol. Biotechnol. 75 (2007) 627–632. [37] R. Margesin, F. Schinner, Potential of halotolerant and halophilic microorganisms for biotechnology, Extremophiles 5 (2001) 73–83. [38] J.L. Slonczewski, M. Fujisawa, M. Dopson, T.A. Krulwich, Cytoplasmic pH measurement and homeostasis in bacteria and archaea, Adv. Microb. Physiol. 55 (1-79) (2009) 317. [39] X. Han, Y.S. Wong, M.H. Wong, N.F. Tam, Biosorption and bioreduction of Cr(VI) by a microalgal isolate, Chlorella miniata, J. Hazard. Mater. 146 (2007) 65–72. [40] B.C. Okeke, J. Laymon, S. Crenshaw, C. Oji, Environmental and kinetic parameters for Cr(VI) bioreduction by a bacterial monoculture purified from Cr(VI)-resistant consortium, Biol. Trace Elem. Res. 123 (2008) 229–241. [41] Z.A. Zakaria, Z. Zakaria, S. Surif, W.A. Ahmad, Hexavalent chromium reduction by Acinetobacter haemolyticus isolated from heavy-metal contaminated wastewater, J. Hazard. Mater. 146 (2007) 30–38. [42] Q.H. Pei, S. Shahir, A.S.S. Raj, Z.A. Zakaria, W.A. Ahmad, Chromium(VI) resistance and removal by Acinetobacter haemolyticus, World J. Microbiol. Biotechnol. 25 (2009) 1085–1093. [43] K.H. Cheung, J.D. Gu, Mechanism of hexavalent chromium detoxification by microorganisms and bioremediation application potential: a review, Int. Biodeterior. Biodegrad. 59 (2007) 8–15.

Chemical Geology 341 (2013) 75–83

Contents lists available at SciVerse ScienceDirect

Chemical Geology journal homepage: www.elsevier.com/locate/chemgeo

Complexation of neptunium(V) with Bacillus subtilis endospore surfaces and their exudates Drew Gorman-Lewis a,⁎, Mark P. Jensen b, Zoë R. Harrold a, Mikaela R. Hertel a a b

University of Washington, Department of Earth and Space Sciences, 070 Johnson Hall, Seattle, WA 98195, United States Chemical Sciences and Engineering Division, Argonne National Laboratory, Argonne, IL 60439, United States

a r t i c l e

i n f o

Article history: Received 10 September 2012 Received in revised form 2 January 2013 Accepted 6 January 2013 Available online 20 January 2013 Editor: Carla M. Koretsky Keywords: Neptunium Endospores Bacillus subtilis Surface complexation Dipicolinic acid

a b s t r a c t The neptunyl ion is very toxic and has the potential to be highly mobile in the environment. In an effort to understand how its interactions with biological surfaces may affect its movement in the environment, we investigated neptunyl interactions with Bacillus subtilis endospores and their exudates. The exudates were dominated by dipicolinic acid. Spectrophotometric investigations of the chemical form of neptunyl in exudate solutions are consistent with the formation of 1:1 neptunyl–dipicolinate complexes. Using neptunyl–endospore adsorption data and spectrophotometric measurements of neptunyl–dipicolinate complexes, we determined thermodynamic stability constants for both species. Neptunyl adsorption onto the endospore surface decreased with an increasing pH, which corresponds to increasing aqueous complexation of neptunyl by dipicolinate. Adsorption was also highly ionic strength dependent with adsorption increasing as ionic strength decreased. With stability constants determined in this work, we compared controls on neptunyl partitioning in a simulated system with B. subtilis endospores, vegetative cells, and generic natural organic matter. Neptunyl complexation by B. subtilis endospore exudates exerted the greatest biological control in the simulated systems. © 2013 Elsevier B.V. All rights reserved.

1. Introduction Neptunium is a manmade element produced as a byproduct of nuclear fission in nuclear fuel. Np can be found as dissolved Np 4+, NpO2+, or NpO22 + ions under aqueous conditions relevant to natural systems, though neptunyl(V), NpO2+, is predicted to be the most common species in natural environments (Fahey, 1986). As a monovalent cation in aqueous solution, Np(V) typically forms less stable complexes than ions with higher charge (Keller, 1971). Due to its monovalency, Np(V) also is less likely to adsorb onto environmental surfaces than Np(IV) and is predicted to be more mobile in the environment (Lemire et al., 2001). However, bacteria can influence the mobility of Np in the environment (Law et al., 2010; Anderson et al., 2011); thus, understanding the fate and transport of Np in the subsurface necessitates investigations of its interactions with biological interfaces and their exudates. Np-related geomicrobiological research has primarily focused on interactions with microbes resulting in the reduction of Np(V) to Np(IV) or surficial interactions with vegetative microbial cells. Many workers reported the occurrence of biologic reduction of Np(V) by metabolizing pure cultures and indigenous microbial communities (Banaszak et al., 1998, 1999; Lloyd et al., 2000; Soderholm et al., 2000; Rittmann et al., 2003; Icopini et al., 2007; Law et al., 2010). ⁎ Corresponding author. E-mail address: [email protected] (D. Gorman-Lewis). 0009-2541/$ – see front matter © 2013 Elsevier B.V. All rights reserved. http://dx.doi.org/10.1016/j.chemgeo.2013.01.004

One instance of non-metabolizing bacterial reduction of Np(V) to Np(IV) has also been reported under low pH-high ionic strength conditions (Gorman-Lewis et al., 2005b). In addition to metabolic transformations of Np(V), direct Np(V)-vegetative bacterial cell surface interactions have been investigated by batch adsorption, X-ray absorption spectroscopy, and surface complexation modeling (Sasaki et al., 2001; Songkasiri et al., 2002; Gorman-Lewis et al., 2005b). Previous Np(V)-microbial research focused on direct Np(V)–cell interactions with little attention to Np(V) reactions occurring with microbial exudates. It is common for microbes to release exudates with substantial complexation capabilities (Tourney et al., 2008; Deo et al., 2010): extracellular polysaccharides (EPS) are one such example. Microbial exudates are widely known to interact with mineral surfaces and dissolved species thus influencing the partitioning of metals (Omoike and Chorover, 2004; Guibaud et al., 2005; Comte et al., 2006; Omoike and Chorover, 2006). While it has been observed many times for other metal ions, the only study to investigate Np(V) interactions with microbial exudates, by Deo et al. (2010), found that Np(V) had a similar adsorption affinity for Shewanella alga EPS as for S. alga whole cells and cell walls. Consequently, in that system S. alga exudates exert controls on Np(V) partitioning if the concentration of the exudate (EPS) is comparable or greater than the effective concentration of the surface EPS/cell walls. The work of Deo et al. highlights the importance of considering not only the direct interactions of Np(V) with microbial cells but also interactions with microbial exudates.

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a

b 60

60 Adsorption Reversibility Reversibility Model

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Fig. 1. (a) (○) Represents Np adsorption at ionic strength of 0.005 M. (∇) Represents adsorption with initial Np–endospore solution equilibrated at pH 6.7 and adjusted down in pH. (□) Represents adsorption with initial Np–endospore solution equilibrated at pH 4.1 and then adjusted up in pH. The curve represents the surface complexation model described in Table 2. (b) (○) Represents Np adsorption at ionic strength of 0.1 M. (∇) Represents adsorption with initial Np–endospore solution equilibrated at pH 6.7 and adjusted down in pH. (□) Represents adsorption with initial Np–endospore solution equilibrated at pH 2.5 and adjusted up in pH. The curve represents the surface complexation model described in Table 2.

Fig. 1b depicts adsorption reversibility at 0.1 M ionic strength, which exhibits similar behavior to low ionic strength systems. Initial adsorption at pH 6.7 was 6% and subsequently lowering pH to 5.7, 4.5, and 3.0 altered adsorption to 3, 7, and 10%, respectively. Experiments with an initial pH of 2.5 and 15% adsorption experience a decrease in the amount of Np adsorbed down to 9, 4, and 3% by raising the pH to 3.3, 4.4, and 5.0, respectively. Adsorption as a function of NpO2+ concentration with a 10 g/L endospore suspension is depicted in Fig. 2. The NpO2+ concentration to endospore ratio was varied from 1.5 to 14. As NpO2+ concentration increases, adsorption increases until a plateau appears at a NpO2+ concentration of 43 μM (Fig. 2). Ionic strength dependence of monovalent cation adsorption onto B. subtilis vegetative cells is a known phenomenon. Alessi et al. (2010) measured Rb + and Li+ adsorption onto B. subtilis vegetative cells and found both ions to be weakly adsorbing and highly ionic strength dependent, similar to NpO2+ adsorption. However, a major difference between monovalent cations Rb+ and Li+ and NpO2+ adsorption is the shape of the adsorption edge as a function pH. Rb + and Li+ adsorption increased as pH increased up to ca. 7 while we observed the opposite behavior for NpO2+ adsorption.

80 Data Model

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Total Np (µM) Fig. 2. Concentration isotherm at ca. pH 6 (exact pH values in Table S1) with a 10 g/L endospore suspension.

Np adsorption onto endospores is very different from adsorption onto vegetative B. subtilis cells. Cation adsorption onto vegetative cells typically increases with increasing pH due to proton active functional groups on the bacterial surface deprotonating, which creates a more negatively charged surface. NpO2+ exhibits this behavior in the presence of vegetative cells at low ionic strength (I = 0.001 M) and high ionic strength (I = 0.1 M) above pH 4.5 (Gorman-Lewis et al., 2005b). Np in contact with endospores exhibited the opposite behavior; adsorption decreased with increased pH at both low and high ionic strength. As shown below, the observed Np–endospore adsorption edge is due to competition between the endospore-surface complex and NpO2(dip)−aqueous complex. Competition between complexation of cations by bacterial surfaces and aqueous organic ligands has also been observed by Song et al. (2009) who found Cd adsorption onto Comamonas spp. inhibited by pthalic acid. Another substantial difference between B. subtilis endospores and vegetative cells in aqueous solutions is the ability of vegetative cells to promote non-metabolic reduction of NpO2+ in low pH and high ionic strength solutions as reported by Gorman-Lewis et al. (2005b). B. subtilis vegetative cells also appear to be capable of non-metabolic reduction of Cr(VI). Fein et al. (2002) found vegetative cells reduced Cr(VI) to Cr(III) at the cell wall and the reaction proceeded much faster under acidic conditions. In both previous studies the authors noted irreversible adsorption and an increase in adsorption with time under the conditions that promoted reduction. Additionally, Fein et al. (2002) performed X-ray adsorption near edge spectroscopy to confirm the oxidation state of Cr on the bacterial surface. While the present work lacks spectroscopic confirmation of the oxidation state of Np on the endospore surface, the adsorption behavior (reversibility and rapid achievement of steady-state adsorption) exhibits none of the indicators of reduction found in the previous work described above. Furthermore, the optical spectroscopy of the supernate shows no evidence of Np(IV) under the conditions investigated. The apparent inability of B. subtilis endospores to promote reduction of Np(V) suggests some fundamental difference between endospore and vegetative cell surface reactivity toward NpO2+. 3.2. NpO2+-exudate interactions Endospores released dipicolinic acid in Np-bearing solutions forming complexes that we investigated with optical spectroscopy.

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Burkholder, W., Grossman, A., 2000. Regulation of the Initiation of Endospore Formation in Bacillus subtilis. American Society for Microbiology, Washington, DC, pp. 151–166. Chinni, S., Anderson, C.R., Ulrich, K.U., Giammar, D.E., Tebo, B.M., 2008. Indirect UO2 oxidation by Mn(II)-oxidizing spores of Bacillus sp. strain SG-1 and the effect of U and Mn concentrations. Environmental Science & Technology 42, 8709–8714. Comte, S., Guibaud, G., Baudu, M., 2006. Relations between extraction protocols for activated sludge extracellular polymeric substances (EPS) and complexation properties of Pb and Cd with EPS: Part II. Consequences of EPS extraction methods on Pb 2+ and Cd2+ complexation. Enzyme and Microbial Technology 38, 246–252. Deo, R.P., Songkasiri, W., Rittmann, B.E., Reed, D.T., 2010. Surface complexation of neptunium(V) onto whole cells and cell components of Shewanella alga: modeling and experimental study. Environmental Science & Technology 44, 4930–4935. Driks, A., 1999. Bacillus subtilis spore coat. Microbiology and Molecular Biology Reviews 63, 1–20. Driks, A., 2002. Maximum shields: the assembly and function of the bacterial spore coat. Trends in Microbiology 10, 251–254. Erlendsson, L.S., Moller, M., Hederstedt, L., 2004. Bacillus subtilis StoA is a thiol-disulfide oxidoreductase important for spore cortex synthesis. Journal of Bacteriology 186, 6230–6238. Fahey, J., 1986. Neptunium, 2nd ed. Chapman Hall, New York. Fein, J.B., Fowle, D.A., Cahill, J., Kemner, K., Boyanov, M., Bunker, B., 2002. Nonmetabolic reduction of Cr(VI) by bacterial surfaces under nutrient-absent conditions. Geomicrobiology Journal 19, 369–382. Fein, J.B., Boily, J.F., Yee, N., Gorman-Lewis, D., Turner, B.F., 2005. Potentiometric titrations of Bacillus subtilis cells to low pH and a comparison of modeling approaches. Geochimica et Cosmochimica Acta 69, 1123–1132. Fisher, M.M., Graham, J.M., Graham, L.E., 1998. Bacterial abundance and activity across sites within two northern Wisconsin sphagnum bogs. Microbial Ecology 36, 259–269. Ginn, B.R., Fein, J.B., 2008. The effect of species diversity on metal adsorption onto bacteria. Geochimica et Cosmochimica Acta 72, 3939–3948. Ginn, B., Fein, J.B., 2009. Temperature dependence of Cd and Pb binding onto bacterial cells. Chemical Geology 259, 99–106. Ginn, B.R., Szymanowski, J.E.S., Fein, J.B., 2010. Calibration of a linear free energy estimation approach for estimating stability constants for metal-bacterial surface complexes. Geomicrobiology Journal 27, 321–328. Gorman-Lewis, D., 2009. Calorimetric measurements of proton adsorption onto Pseudomonas putida. Journal of Colloid and Interface Science 337, 390–395. Gorman-Lewis, D., Elias, P.E., Fein, J.B., 2005a. Adsorption of aqueous uranyl complexes onto Bacillus subtilis cells. Environmental Science & Technology 39, 4906–4912. Gorman-Lewis, D., Fein, J.B., Soderholm, L., Jensen, M.P., Chiang, M.H., 2005b. Experimental study of neptunyl adsorption onto Bacillus subtilis. Geochimica et Cosmochimica Acta 69, 4837–4844. Gorman-Lewis, D., Fein, J.B., Jensen, M.P., 2006. Enthalpies and entropies of proton and cadmium adsorption onto Bacillus subtilis bacterial cells from calorimetric measurements. Geochimica et Cosmochimica Acta 70, 4862–4873. Gould, G.W., Hurst, A., 1969. The Bacterial Spore. Academic Press, London, New York. Guibaud, G., Comte, S., Bordas, F., Baudu, M., 2005. Metal removal from single and multimetallic equimolar systems by extracellular polymers extracted from activated sludges as evaluated by SMDE polarography. Process Biochemistry 40, 661–668. Guine, V., Spadini, L., Sarret, G., Muris, M., Delolme, C., Gaudet, J.P., Martins, J.M.F., 2006. Zinc sorption to three gram-negative bacteria: combined titration, modeling, and EXAFS study. Environmental Science & Technology 40, 1806–1813. Harrold, Z., Gorman-Lewis, D., in press. Thermodynamic characteriaztion of Bacillus subtilis spore-proton adsorption. Geochimica et Cosmochiica Acta. Harrold, Z.R., Hertel, M.R., Gorman-Lewis, D., 2011. Optimizing Bacillus subtilis spore isolation and quantifying spore harvest purity. Journal of Microbiological Methods 87, 325–329. Henriques, A.O., Moran, C.P., 2000. Structure and assembly of the bacterial endospore coat. Methods 20, 95–110. Henriques, A.O., Moran, C.P., 2007. Structure, assembly, and function of the spore surface layers. Annual Review of Microbiology 61, 555–588 (Henriques, Adriano O. Moran, Charles P., Jr.). Herbelin, A., Westall, J., 1994. FITEQL 3.1, a computer program for determination of chemical equilibrium constants from experimental data. Technical Report 94-01. Dept. of Chem. Oregon St. Univ. Hong, H.A., To, E., Fakhry, S., Baccigalupi, L., Ricca, E., Cutting, S.M., 2009. Defining the natural habitat of Bacillus spore-formers. Research in Microbiology 160, 375–379. Icopini, G.A., Boukhalfa, H., Neu, M.P., 2007. Biological reduction of Np(V) and Np(V) citrate by metal-reducing bacteria. Environmental Science & Technology 41, 2764–2769. Inoue, Y., Tochiyama, O., 1983. Study of the complexes of Np(V) with organic-ligands by solvent-extraction with TTA and 1,10-phenanthroline. Polyhedron 2, 627–630. Jensen, M.P., Nash, K.L., 2001. Thermodynamics of dioxoneptunium(V) complexation by dicarboxylic acids. Radiochimica Acta 89, 557–564. Junier, P., Frutschi, M., Wigginton, N.S., Schofield, E.J., Bargar, J.R., Bernier-Latmani, R., 2009. Metal reduction by spores of Desulfotomaculum reducens. Environmental Microbiology 11, 3007–3017. Keller, C., 1971. Neptunium. Kernchemie in Einzeldarstellungen, vol. 3. V C H Publishers, Weinheim. Kisand, V., Tuvikene, L., Noges, T., 2001. Role of phosphorus and nitrogen for bacteria and phytopankton development in a large shallow lake. Hydrobiologia 457, 187–197.

Lalonde, S.V., Smith, D.S., Owttrim, G.W., Konhauser, K.O., 2008a. Acid–base properties of cyanobacterial surfaces I: influences of growth phase and nitrogen metabolism on cell surface reactivity. Geochimica et Cosmochimica Acta 72, 1257–1268. Lalonde, S.V., Smith, D.S., Owttrim, G.W., Konhauser, K.O., 2008b. Acid–base properties of cyanobacterial surfaces. II: silica as a chemical stressor influencing cell surface reactivity. Geochimica et Cosmochimica Acta 72, 1269–1280. Law, G.T.W., Geissler, A., Lloyd, J.R., Livens, F.R., Boothman, C., Begg, J.D.C., Denecke, M.A., Rothe, J., Dardenne, K., Burke, I.T., Charnock, J.M., Morris, K., 2010. Geomicrobiological redox cycling of the transuranic element neptunium. Environmental Science & Technology 44, 8924–8929. Lee, Y., Tebo, B.M., 1994. Cobalt (II) oxidation by the marine manganese (II)-oxidizing Bacilllus sp. strain SG-1. Applied and Environmental Microbiology 60, 2949–2957. Leggett, D., 1985. SQUAD: Stability Quotients from Absorbance Data. Plenum Press, New York. Lemire, R.J., Fuger, J., Nitsche, H., Rand, M.H., Potter, P., Rydberg, J., Spahiu, K., Sullivan, J.C., Ullman, W.J., Vitorage, P., Wanner, H., 2001. Chemical Thermodynamics of Neptunium and Plutonium. Elsevier, New York. Lloyd, J.R., Yong, P., Macaskie, L.E., 2000. Biological reduction and removal of Np(V) by two microorganisms. Environmental Science & Technology 34, 1297–1301. Martell, A.E., Smith, R.M., Motekaitis, R.J., 1998. Critically selected stability constants of metal complexes database version 5.0. NIST Standard Reference Data, Gaithersburg, Maryland, USA: NIST Standard Reference Database, vol. 46. Nicholson, W.L., 2002. Roles of Bacillus endospores in the environments. Cellular and Molecular Life Sciences 59, 410–416. Nicholson, W.L., Munakata, N., Horneck, G., Melosh, H.J., Setlow, P., 2000. Resistance of Bacillus endospores to extreme terrestrial and extraterrestrial environments. Microbiology and Molecular Biology Reviews 64, 548–572. Nicholson, W.L., Fajardo-Cavazos, P., Rebeil, R., Slieman, T.A., Riesenman, P.J., Law, J.F., Xue, Y.M., 2002. Bacterial endospores and their significance in stress resistance. Antonie Van Leeuwenhoek International Journal of General and Molecular Microbiology 81, 27–32. Omoike, A., Chorover, J., 2004. Spectroscopic study of extracellular polymeric substances from Bacillus subtilis: aqueous chemistry and adsorption effects. Biomacromolecules 5, 1219–1230. Omoike, A., Chorover, J., 2006. Adsorption to goethite of extracellular polymeric substances from Bacillus subtilis. Geochimica et Cosmochimica Acta 70, 827–838. Paidhungat, M., Setlow, P., 2000. Role of Ger proteins in nutrient and nonnutrient triggering of spore germination in Bacillus subtilis. Journal of Bacteriology 182, 2513–2519. Phoenix, V.R., Martinez, R.E., Konhauser, K.O., Ferris, F.G., 2002. Characterization and implications of the cell surface reactivity of Calothrix sp. strain KC97. Applied and Environmental Microbiology 68, 4827–4834. Phoenix, V.R., Korenevsky, A.A., Ferris, F.G., Gorby, Y.A., Beveridge, T.J., 2007. Influence of lipopolysaccharide on the surface proton-binding behavior of shewanella spp Current Microbiology 55, 152–157. Rao, L., Choppin, G.R., 1995. Thermodynamic study of the complexation of neptunium(V) with humic acids. Radiochimica Acta 69, 87–95. Rittmann, B.E., Banaszak, J.E., Reed, D.T., 2003. Reduction of Np(V) and precipitation of Np(IV) by an anaerobic microbial consortium. Biodegradation 13, 329–342. Rizkalla, E.N., Nectoux, F., Dabosseignon, S., Pagès, M., 1990a. Complexation of neptunium(V) by halo- and hydroxycarbonate ligands. Radiochimica Acta 51, 113–117. Rizkalla, E.N., Nectoux, F., Dabosseignon, S., Pagès, M., 1990b. Complexation of neptunium(V) by polyaminocarboxylate ligands. Radiochimica Acta 51, 151–155. Sasaki, T., Zheng, J., Asano, H., Kudo, A., 2001. Interaction of Pu, Np and Pa with anaerobic microorganisms at geologic repositories. Radioactivity in the Environment 1, 221–232. Setlow, P., 2003. Spore germination. Current Opinion in Microbiology 6, 550–556. Setlow, B., Loshon, C., Genest, P., Cowan, A., Setlow, C., Setlow, P., 2002. Mechanisms of killing spores of Bacillus subtilis by acid, alkali and ethanol. Journal of Applied Microbiology 92, 362–375. Siala, A., Hill, I.R., Gray, T.R.G., 1974. Populations of spore-forming bacteria in acid forest soil, with special reference to Bacillus subtilis. Journal of General Microbiology 81, 183–190. Skanthakumar, S., Gorman-Lewis, D., Locock, A., Chiang, M.H., Jensen, M.P., Burns, P.C., Fein, J., Jonah, C.D., Attenkofer, K., Soderholm, L., 2004. Changing Np redox speciation in the synchrotron beam. Materials Research Society Symposium Proceedings 802, 151–156. Slieman, T.A., Nicholson, W.L., 2001. Role of dipicolinic acid in survival of Bacillus subtilis spores exposed to artificial and solar UV radiation. Applied and Environmental Microbiology 67, 1274–1279. Soderholm, L., Williams, C., Antonio, M., Tischler, M., Markos, M., 2000. The influence of Deulfovibrio desulfuricans on neptunium chemistry. Materials Research Society Symposium Proceedings 590, 27–32. Song, Y., Swedlund, P.J., Singhal, N., Swift, S., 2009. Cadmium(II) speciation in complex aquatic systems: a study with ferrihydrite, bacteria, and an organic ligand. Environmental Science & Technology 43, 7430–7436. Songkasiri, W., Reed, D., Rittman, B., 2002. Bio-sorption of neptunium(V) by Pseudomonas fluorescens. Radiochimica Acta 90, 785–789. Thurman, E., 1985. Organic Geochemistry of Natural Waters, 1 ed Martinus Nijhoff/Dr W Junk Publishers, Dordrecht. Tian, G., Rao, L., Teat, S.J., 2009. Thermodynamics, optical properties, and coordination modes of Np(V) with dipicolinic acids. Inorganic Chemistry 48, 10158–10164. Tochiyama, O., Inoue, Y., Narita, S., 1992. Complex-formation of Np(V) with various carboxylates. Radiochimica Acta 58−9, 129–136.

Geomicrobiology Journal (2013) 30, 422–429 C Taylor & Francis Group, LLC Copyright  ISSN: 0149-0451 print / 1521-0529 online DOI: 10.1080/01490451.2012.705228

Chromium Adsorption by Three Yeast Strains Isolated from Sediments in Morocco WIFAK BAHAFID, NEZHA TAHRI JOUTEY, HANANE SAYEL, MOHAMED IRAQUI-HOUSSAINI, and NAI¨MA EL GHACHTOULI∗ Microbial Biotechnology Laboratory, Faculty of Sciences and Technology, Sidi Mohammed Ben Abdellah University, Fez, Morocco

Downloaded by [University of Notre Dame] at 12:51 24 August 2014

Received February 2012, Accepted June 2012

Biosorption is an effective method to remove heavy metals from wastewater. In this work, Biosorption of Cr(VI) has been investigated by live and dead cells of three yeasts species: Cyberlindnera fabianii, Wickerhamomyces anomalus and Candida tropicalis. Sorption experiments were conducted in aqueous solutions at various pH conditions. Cr(VI) adsorption was highly pH dependent and the results indicated that the most effective pH range was found to be between 2 and 4 for the three species. Adsorption isotherms were modeled with the Langmuir and Freundlich equations and isotherm constants were calculated. The adsorption capacity calculated from Langmuir isotherm was 18.9 mg, 28.14 mg and 29.1 mg Cr(VI) g−1 Cr(VI) g−1 for C. fabianii, W. anomalus and C. tropicalis, respectively. The results suggest that the three yeasts could be used as effective adsorbents for the removal of Cr(VI) ions from contaminated sites. Keywords: biosorption, chromium, isotherms, pH, yeast biomass

Introduction Heavy metal ions are extremely toxic and harmful even at low concentrations, which can seriously affect plants and animals and have been involved in causing a large number of afflictions (Bulut and Tez 2007; Chua et al. 1999; Martins et al. 2006). Chromium heavy metal, is widely used in many important industrial applications, such as steel production, electroplating, leather tanning, textile industries, wood preservation, anodizing of aluminum, water-cooling and chromate preparation (Garg et al. 2007), but is also one of the most toxic heavy metals (Silva et al. 2009), with high water solubility and mobility. It is generally accepted that Cr(III) species are not highly toxic and due to their limited solubility, they can be easily removed by chemical and microbial approaches. It has also been suggested that this ion is involved in the tertiary structure of proteins and the conformation of cell RNA and DNA (Gulan et al. 2001; Zayed and Terry 2003). In contrast, Cr(VI) is always toxic and exhibits mutagenic and carcinogenic effects on biological systems (Codd et al. 2001; Costa 2003). Therefore,

The authors thankfully acknowledge the financial and scientific support of Microbial Biotechnology Laboratory, Faculty of Sciences and Technology and of Regional Center of Interface (CURI), SMBA University, Fez, Morocco. ∗ Address correspondence to Na¨ıma El Ghachtouli, Microbial Biotechnology Laboratory, SMBA University, Faculty of Sciences and Technology, Route Immouzer, P. O. Box 2202, Fez, Morocco; Email: [email protected]

the elimination of this metal from water and wastewaters is important to protect public health. The use of microbial biomass for the removal of toxic heavy metal ions from wastewaters has emerged as an alternative to the existing methods which include chemical precipitation, ion exchange, membrane separation, reverse osmosis, evaporation and electrochemical treatment as a result of the search of lowcost, innovative methods (Kapoor and Viraraghavan 1998; Rengeraj et al. 2001). Biosorption is energy independent binding of metal to the cell wall of organism such as algae, fungi and bacteria, for removal of metal ions (Gupta et al. 2001; Nourbakhsh et al. 2002; Sag and Kutsal 2000). Particularly, fungal biomass can be cheaply and easily procured in rather substantial quantities, as a byproduct from established industrial fermentation processes. Furthermore, since such abundant dead fungal biomass is of little use, it has been identified as a potential source of biomaterial for the removal of chromium from wastewaters. Live or dead fungal cells can be used as an adsorbent material for the removal of toxic metal ions from aqueous solutions (Sanghi et al. 2009), but non-living biomass appears to present specific advantages in comparison with the use of living microorganisms. Killed cells may be stored or used for extended periods at room temperature. They are not subject to metal toxicity, nutrient supply is not necessary and the biosorbed metal ions can be easily desorbed and biomass can be reused. It has also been reported that cell walls, consisting mainly of polysaccharides, proteins and lipids, offer many functional groups that can bind metal ions. In addition to these functional binding groups, polysaccharides often have ion exchange properties (Sag and Kutsal 1996; Veglio et al. 1997; Zouboulis et al. 1999). As a result, use of dead fungal

Chromium Adsorption by Three Yeast Strains Isolated

425

Table 1. Concentration of chromium adsorbed by live and dead yeasts biomass Yeasts

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C. fabianii Live Dead W. anomalus Live Dead C. tropicalis Live Dead

Uptake (mg/g)

% Inhibition

0.48 ± 0.07 0.37 ± 0.13

— 22%

0.47 ± 0.09 0.3 ± 0.10

— 37%

0.5 ± 0.06 0.41 ± 0.09

— 18%

mechanism. Batch biosorption experiment with 10 mg/L of chromium at pH 3 showed that after 24 h, 0.2 g live yeasts adsorb about 0.5 mg of Cr(VI) per g of cells for all the yeasts, while dead biomass adsorb 0.37 ± 0.13 mg/g, 0.3 ± 0.1 mg/g and 0.41 ± 0.09 mg/g of Cr(VI) for C. fabianii HE650139, W. anomalus HE648168 and C. tropicalis HE650140, respectively (Table 1). It was observed that the dead biomass adsorbs less chromium than the corresponding live biomass under identical conditions. Similar results were found by Das and Guha (2009), using Termitomyces clypeatus. The reduced uptake of chromium by dead biomass may be due to either loss of some binding sites resulting from heat inactivation of cells or restraint of intracellular chromium accumulation and enzymatic reduction as in the case of viable cells (Burnett et al. 2006; Fein et al. 2002; Tobin et al. 1994). Effect of pH The pH plays a vital role in biosorption of Cr(VI) due to the nature of chemical interactions of each metal with the functional groups present on the microbial cell surface (Wang and Can 2006). The results of the effect of pH on the biosorption of Cr(VI) ions onto dead and living cells of C. fabianii HE650139, W. anomalus HE648168 and C. tropicalis HE650140 were reported as the percentage of Cr(VI) removal (Figures 2a, 2b and 2c). As it can be seen in Figure 2, the removal of Cr(VI) was strongly pH dependent. The removal efficiency of Cr(VI) increased slightly between pH 2 and 4. Above pH 4, there was gradual decrease in removal efficiency up to pH 9. Maximum adsorption capacities by both living and dead yeast were found at pH 4.0 for C. fabianii HE650139 and W. anomalus HE648168 (Figures 2a and 2b) and 3.0 for C. tropicalis HE650140 (Figure 2c), with a percentage of removal of 100%, 70% and 97%, respectively, by dead cells, and of 100% by all living microorganisms. Our results are in agreement with the findings of CardenasGonzalez and Acosta-Rodriguez (2010), who demonstrated that the maximum uptake was observed at pH 4.0 (96% at 7 days) using respectively Paecilomyces sp. and Helmintosporium sp. Acosta et al. (2004) also found that the biosorption of chromium on C. neoformans and Helmintosporium sp. increased as the initial pH of medium decreased.

Fig. 2. Effect of pH on percent removal of Cr(VI) by C. fabianii HE650139 (a), W. anomalus HE648168 (b) and C. tropicalis HE650140 (c).

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as indicated by high maximum biosorption capacity qmax (29.1 mg/g, 28.14 mg/g and 18.9 mg/g) by C. tropicalis HE650140, W. anomalus HE648168 and C. fabianii HE650139, respectively. The variation of the adsorption intensity (RL ) with the initial concentration of the solution (C 0 mg/L) was determined (Table 3).From the result it appears that the RL value approaches zero with increase in the C 0 value, which confirmed that the three yeasts are a suitable biosorbent for adsorption of chromium from wastewater under the conditions used in this study. The three yeasts were compared with other adsorbents based on their maximum adsorption capacity for Cr(VI) and shown in Table 4. It can be observed that the three yeasts compares well with the other adsorbents listed in Table 4.

Conclusion The results of the present study demonstrated the ability of the three yeasts to remove Cr(VI) from aqueous solutions. It showed that the sorption capability of the three biosorbents was directly affected by the pH condition in the solution of the sorption experiments. For the maximum adsorption, the optimal pH of solution were optimized as pH 4 for W. anomalus HE648168 and C. fabianii HE650139, although for C. tropicalis HE650140, the uptake of Cr(VI) was maximum at pH 3. Linear Langmuir and Freundlich isotherm models were used to represent the experimental data. Adsorption data fitted well with the Langmuir and Freundlich models. However, Freundlich isotherm displayed a better fitting model than Langmuir isotherm for W. anomalus HE648168 and C. tropicalis HE650140 because of the higher correlation coefficient, thus, indicating the applicability of monolayer coverage of the Cr(VI) ion on the surface of adsorbent. Based on the results of this research, the three yeasts can be considered as an effective, available, naturals and excellent adsorbents for removing chromium.

References Acosta RI, Rodriguez X, Gutiorrez C, Moctezuma MG. 2004. Biosorption of chromium (VI) from aqueous solutions onto fungal biomass. Bioorg Chem Appl 2:1–2. Babarinde NAA, Oyesiku OO, Babalola JO, Olatunji JO. 2008. Isothermal and thermodynamic studies of the biosorption of Zinc (II) ions by Calymperes erosum. J Appl Sci Res 4:716–721. Bahafid W, Sayel H, Tahri Joutey N, Ghachtouli N EL. 2011. Removal mechanism of hexavalent chromium by a novel strain of Pichia anomala isolated from industrial effluents of Fez (Morocco). J Environ Sci Eng 5:980–991. Bulut Y, Baysal Z. 2006. Removal of Pb(II) from wastewater using wheat bran. J Environ Manage 78:107–113. Bulut Y, Gozubenli N, Aydın H. 2007. Equilibrium and kinetics studies for adsorption of direct blue 71 from aqueous solution by wheat shells. J Hazard Mater 144:300–306. Bulut Y, Tez Z. 2007. Adsorption studies on ground shells of hazelnut and almond. J Hazard Mater 149:35–41. Burnett PGG, Daughney CJ, Peak D. 2006. Cd adsorption onto Anoxybacillus flavithermus: surface complexation modeling and spectroscopic investigations. Geochim Cosmochim Acta 70:5253–5269.

Bahafid et al. Cardenas-Gonzalez JF, Acosta-Rodrıguez I. 2010. Hexavalent chromium removal by a Paecilomyces sp. fungal strain isolated from environment. Bioinorg Chem Appl 2010:1–6. Chojnacka K. 2010. Biosorption and bioaccumulation–The prospects for practical applications. Environ Inter 36:299–307. Chua WLO, Lam HKH, Bi SP. 1999. A comperative investigation on the biosorption of lead by filamentous fungal biomass. Chemosphere 39:2723–2736. Cieslak-Golonka M. 1991. Chem Inform Abstract: Spectroscopy of chromium(VI) species. Coord Chem Rev 109:223–249. Codd R, Dillon CT, Levina A, Lay PA. 2001. Studies on the genotoxicity of chromium: from the test tube to the cell. Coord Chem Rev (216–217):537–582. Costa M. 2003. Potential hazards of hexavalent chromate in our drinking water. Toxicol Appl Pharmacol 188:1–5. Cotton FA, Wilkinson G. 1980. Advanced Inorganic Chemistry. 4th ed. New York: John Wiley & Sons. Das SK, Guha AK. 2007. Biosorption of chromium by Termitomyces clypeatus. Coll Surf B Biointer 60:46–54. Das SK, Guha AK. 2009. Biosorption of hexavalent chromium by Termitomyces clypeatus biomass: Kinetics and transmission electron microscopic study. J Hazard Mater 167:685–691. Febrianto J, Kosasiha AN, Sunarso J, Ju Y, Indraswati N, Ismadji S. 2009. Equilibrium and kinetic studies in adsorption of heavy metals using biosorbent: A summary of recent studies. J Hazard Mater 162:616–645. Fein JB, Fowle DA, Cahill J, Kemner K, Boyanov M, Bunker B. 2002. Nonmetabolic reduction of Cr(VI) by bacterial surfaces under nutrient-absent conditions. J Geomicrobiol 19:369–382. Freundlich H. 1907. Ueber die adsorption in Loesungen. Z Phys Chem 57:385–470. Garg UK, Kaur MP, Garg VK, Sud D. 2007. Removal of hexavalent chromium from aqueous solution by agricultural waste biomass. J Hazard Mater 140:60–68. Gulan ZV, Stehlik TV, Grba S, Lutilsky LD. 2001. Chromium uptake by Saccharomyces cerevisiae and isolation of glucose tolerance factor from yeast biomass. J Biosci 26:217–223. Gupta VK, Shrivastava AK, Jain N. 2001. Biosorption of chromium(VI) from aqueous solutions by green algae Spirogyra species. Water Res 35:4079–4085. Hall K, Eagleton L, Acrivos A, Vermeulen T. 1966. Pore and solid diffusion kinetics in fixed bed adsorption under constant pattern conditions. Ind Eng Chem Fundam 5:212–223. Kapoor A, Viraraghavan T. 1998. Biosorption of heavy-metal on Aspergillus niger: Effect of pretreatment. Biores Technol 63:109–113. Khambhaty Y, Mody K, Basha S, Jha B. 2009. Biosorption of inorganic mercury onto dead biomass of marine Aspergillus niger: Kinetic, equilibrium, and thermodynamic studies. Environ Engin Sci 26:531–539. Kiran I, Akar T, Tunali S. 2005. Biosorption of Pb (II) and Cu (II) from aqueous solutions by pretreatment biomass on Neurospora crassa. Proc Biochem 40:3550–3558. Ksungur YG, Ren S, Ven UG. 2003. Biosorption of copper ions by caustic treated waste Baker’s Yeast biomass. Turk J Biol 27:23–29. Kumar R, Bishnoi NR, Garima K. 2008. Biosorption of chromium(VI) from aqueous solution and electroplating wastewater using fungal biomass. Chem Eng J 135:202–208. Langmuir I. 1918. The adsorption of gases in plain surfaces of glass, mica, and platinum. J Am Chem Soc 40:1361–1403. Levankumar L, Muthukumaran V, Gobinath MB. 2009. Batch adsorption and kinetics of chromium (VI) removal from aqueous solutions by Ocimum americanum L. seed pods. J Hazard Mater 161:709–713. Li J, Lin Q, Zhang X, Yan Y. 2009. Kinetic parameters and mechanisms of the batch biosorption of Cr(VI) and Cr(III) onto Leersia hexandra Swartz biomass. J Coll Interf Sci 333:71–77. Lokeshwari N, Joshi K. 2009. Biosorption of heavy metal (Chromium) using biomass. Global J Environ Res 3:29–35.

BIOMINERALIZATION AND BIOSORPTION INVOLVING BACTERIA: METAL PHOSPHATE PRECIPITATION AND MERCURY ADSORPTION EXPERIMENTS

A Dissertation

Submitted to the Graduate School of the University of Notre Dame in Partial Fulfillment of the Requirements for the Degree of

Doctor of Philosophy by Sarrah M. Dunham-Cheatham

Jeremy B. Fein, Director

Graduate Program in Civil and Environmental Engineering and Earth Sciences Notre Dame, Indiana August, 2012

magnesium and calcium adsorption to the bacterial cell wall. The adsorbed metal ions then attract carbonate anions, which result from the metabolism of organic nutrients, beginning the precipitation of calcium and magnesium carbonate phases on the bacterial cell wall. Virtually all research investigating biomineralization has involved metabolizing bacteria. However, bacteria exist under oligotrophic conditions in a wide range of natural systems (Billen et al., 1990; Noe et al., 2001). A number of studies (e.g. Ferris et al., 1987; Lowenstam & Weiner, 1989; Châtellier et al., 2001; Ben Chekroun et al., 2004; Beazley et al., 2007; Dupraz et al., 2009) have proposed that the functional groups on the cell walls of bacteria can act as nucleation sites for the non-metabolic precipitation of minerals, leading to a third type of biomineralization which I refer to as passive biomineralization. Despite these claims in the literature, the evidence in support of passive biomineralization is equivocal. Studies have shown associations between bacterial cells and mineral precipitates (e.g. Konhauser et al., 1993), but a spatial association itself does not prove that the cell wall caused the mineral precipitation; the association could be a result of electrostatic interactions between previously precipitated minerals and the cells. Despite the growing number of claims, no study to date has unequivocally demonstrated that the process of passive binding of metal cations to cell wall ligands affects mineral precipitation or that cell wall nucleation of precipitates can occur. Chapter 2 presents research that unequivocally demonstrates the ability of cell walls to passively nucleate the precipitation of minerals within the cell wall matrix under some saturation state conditions and for some elements. Metal transport in groundwater systems can also be affected by the adsorption of aqueous metal cations onto charged surfaces (e.g., bacterial cell walls) and by the formation of aqueous complexes. The adsorption of a wide range of metals onto bacterial cells has been studied (e.g. Beveridge and Murray, 1976, 1980; Beveridge, 1989; Mullen et al., 1989; Fein et al., 3

1997, 2002; Borrok et al., 2004, 2007; Wu et al., 2006). The cell wall of a bacterium contains proton-active functional groups, such as carboxyl, phosphoryl, hydroxyl, amino, and sulfhydryl groups (Beveridge and Murray, 1976; Degens and Ittekkot, 1982; Guiné et al., 2006; Madigan et al., 2009; Mishra et al., 2009, 2010). When deprotonated, these functional groups have the ability to adsorb cations (e.g. metals, aqueous complexes) from solution (Beveridge and Murray, 1976; Ledin et al., 1996; Fortin and Beveridge, 1997; Warren and Ferris, 1998; Ohnuki et al., 2005; Borrok et al., 2007). It has been shown that adsorption of metals to bacterial surfaces is rapid (Fowle and Fein, 2000; Yee et al., 2000), dependent on solution pH (Fein, 2006), and reversible (Fowle and Fein, 2000). In addition to affecting metal mobility, metal adsorption likely represents the first step in bioavailability of metals to bacteria. According to the Biotic Ligand Model, the bioavailability of toxic metals, such as Hg, is a result of the adsorption of the metal to a biological surface of the living organism (Di Toro et al., 2001; Santore et al., 2001; Paquin et al., 2002; Niyogi and Wood, 2004; van Leeuwen et al., 2005). Thus, it is important to construct quantitative models of Hg adsorption onto bacteria that are capable of accounting for Hg partitioning under a range of conditions of geologic and environmental interest. Mercury is of particular interest because it might exhibit different aqueous complexation behavior and/or form different types of bonds than other previously studied metals. For instance, because it is a B-type metal, Hg has a high affinity to bond with sulfur ligands (Reddy and Aiken, 2000; Ravichandran et al., 2004). Because bacterial cell walls contain sulfhydryl functional groups (Mishra et al., 2009, 2010) and natural organic matter contains sulfur compounds (Haitzer et al., 2003; Hertkorn et al., 2008), the affinity of Hg for sulfur compounds may have a significant effect on the behavior of Hg adsorption behavior in the presence of bacteria and natural organic matter.

4

experimental conditions, individual stability constants for Hg-bacterial surface complexes cannot be determined in the Cl-free system. In the presence of chloride, all of the bacterial species exhibit minimal Hg adsorption below pH 4, increasing adsorption between pH 4 and 8, and slightly decreasing extents of adsorption with increasing pH above 8. The low extent of adsorption at low pH suggests that HgCl20, which dominates aqueous Hg speciation below pH 5.5, adsorbs only weakly. The increase in Hg adsorption above pH 4 is likely due to adsorption of HgCl(OH)0, and is limited by site availability and transformation to Hg(OH) 20 as pH increases. I use the adsorption data to determine stability constants of the HgCl(OH)- and Hg(OH)2-bacterial cell envelope complexes, and the values enable estimations to be made for Hg adsorption behavior in bacteria-bearing geologic systems.

3.2 Introduction Bacteria are present in soils and groundwater systems (Madigan et al., 2009), and adsorption onto bacterial cell envelope functional groups can affect the speciation, distribution and transport of heavy metals (Beveridge and Murray, 1976; Fortin et al., 1997; Ledin et al., 1999; Small et al., 1999; Daughney et al., 2002). Although the adsorption behaviors of a wide range of bacteria have been studied for a wide range of metals (e.g., Beveridge and Murray, 1976, 1980; Beveridge, 1989; Mullen et al., 1989; Fein et al., 1997, 2002; Borrok et al, 2004, 2007; Wu et al., 2006), Hg has received less attention. Recent studies have found that protonactive sulfhydryl functional groups exist on the surface of bacterial cell envelopes (Guine et al., 2006; Mishra et al., 2009; 2010). Many studies have demonstrated that Hg has a high binding affinity for sulfur compounds (Fuhr and Rabenstein, 1973; Blum and Bartha, 1980; Compeau and Bartha, 1987; Winfrey and Rudd, 1990; Benoit et al., 1999), and thus the adsorption of Hg to bacteria may be dominated by this type of binding. Due to the high affinity for this bond to 63

Deo R. P., Songkasiri W., Rittmann B. E., and Reed D. T. (2010) Surface complexation of neptunium(V) onto whole cells and cell components of Shewanella alga: Modeling and experimental study. Environmental Science & Technology. 44. 4930-4935. Di Toro D. M., Allen H. E., Bergman H. L., Meyer J. S., Paquin P. R., Santore R. C. (2001) Biotic ligand model of the acute toxicity of metals: Technical basis. Environmental Toxicology and Chemistry. 20. 2383-2396. Dong W. M., Bian Y. R., Liang L. Y., and Gu B. H. (2011) Binding Constants of Mercury and Dissolved Organic Matter Determined by a Modified Ion Exchange Technique. Environmental Science & Technology. 45. 3576-3583. Douglas S. and Beveridge T. J. (1998) Mineral formation by bacteria in natural microbial communities. FEMS Microbiology Ecology. 26. 79-88. Drexel R. T., Haitzer M., Ryan J. N., Aiken G. R., and Nagy K. L. (2002) Mercury(II) Sorption to Two Florida Everglades Peats: Evidence for Strong and Weak Binding and Competition by Dissolved Organic Matter Released from the Peat. Environmental Science & Technology. 36. 4058-4064. Dunham-Cheatham S., Farrell B., Mishra B., Myneni S., and Fein J. B. (2012) The effect of chloride on the adsorption of Hg onto three bacterial species. In preparation. Dupraz C., Reid R. P., Braissant O., Decho A. W., Norman R. S., and Visscher P. T. (2009) Processes of carbonate precipitation in modern microbil mats. Earth-Science Reviews. 96. 141-162. Ephraim J. H. (1992) Heterogeneity as a concept in the interpretation of metal ion binding by humic substances. The binding of zinc by an aquatic fulvic acid. Analytica Chimica Acta. 267. 39-45. Farley K. J., Dzombak D. A., and Fmm M. (1985) A surface precipitation model for the sorption of cations on metal-oxides. Journal of Colloid and Interface Science. 106. 226-242. Fein J. B., Daughney C. J., Yee N., and Davis T. A. (1997) A chemical equilibrium model for metal adsorption onto bacterial surfaces. Geochimica et Cosmochimica Acta. 61. 3319-3328. Fein J. B., Boily J.-F., Güçlü K., and Kaulbach E. (1999) Experimental study of humic acid adsorption onto bacteria and Al-oxide mineral surfaces. Chemical Geology. 162. 33-45. Fein J. B., Martin A. M., and Wightman P. G. (2001) Metal adsorption onto bacterial surfaces: Development of a predictive approach. Geochimica et Cosmochimica Acta. 65. 42674273. Fein J. B., Fowle D. A., Cahill J., Kemner K., Boyanov M. and Bunker B. (2002) Nonmetabolic reduction of Cr(VI) by bacterial surfaces under nutrient-absent conditions. Geomicrobiology Journal. 19. 369-382. 113

Geomicrobiology Journal, 29:173–185, 2012 Copyright © Taylor & Francis Group, LLC ISSN: 0149-0451 print / 1521-0529 online DOI: 10.1080/01490451.2010.539662

Aerobic Reduction of Chromium(VI) by Pseudomonas corrugata 28: Influence of Metabolism and Fate of Reduced Chromium Iso Christl,1 Martin Imseng,1 Enrico Tatti,2 Jakob Frommer,1 Carlo Viti,2 Luciana Giovannetti,2 and Ruben Kretzschmar1 1

Institute of Biogeochemistry and Pollutant Dynamics, ETH Zurich, Zurich, Switzerland Dipartimento di Biotecnologie Agrarie, Sez. Microbiologia, Universit`a degli Studi di Firenze, Florence, Italy

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Pseudomonas corrugata 28 represents a microorganism that can potentially be applied for in situ bioremediation of Cr(VI) contaminated sites. This strain combines a high resistance toward toxic Cr(VI) with the ability to reduce Cr(VI) to Cr(III) under oxic conditions. In this study, the aerobic reduction of Cr(VI) by Pseudomonas corrugata 28 was examined under different carbon and sulfur supply conditions to assess the influence of microbial carbon and sulfur metabolism on Cr(VI) reduction. The fate of reduced chromium was elucidated by investigating the speciation of chromium in solution as well as the interaction of chromium with bacterial surfaces. Reduction of Cr(VI) was found to be a metabolic process resulting mainly in the formation of dissolved organic Cr(III)-complexes. Small amounts of reduced chromium were weakly associated with bacterial surfaces. The formation of inorganic Cr(III)-precipitates was not indicated. Keywords

hexavalent chromium, microbial reduction, aerobic conditions, organic metal complexes, scanning transmission X-ray microscopy (STXM), Pseudomonas corrugata 28

INTRODUCTION The inorganic speciation of chromium is strongly determined by its redox properties. Under strongly oxidizing conditions,

Received 26 July 2010; accepted 1 November 2010. The X-ray microscopy studies were conducted at the STXM beamlines X07D (PolLux), SLS, Paul Scherrer Institute, Villigen, Switzerland and X-1A, NSLS, Brookhaven National Laboratory, Upton/NY, USA. The development of the X-ray microscopes at SLS and NSLS was financially supported by BMBF, Germany under project 05 KS4 WE1/6 as well as the Office of Biological and Environmental Research, U.S. DoE under contract DE-FG02-89ER60858 and the NSF under grant DBI-9605045, respectively. We are deeply grateful to George Tzvetkov and Sue Wirick for their help at the beamlines. Address correspondence to Iso Christl, Institute of Biogeochemistry and Pollutant Dynamics, ETH Zurich, CHN, 8092 Z¨urich, Switzerland. E-mail: [email protected]

hexavalent chromium, Cr(VI), predominates (Katz and Salem 1994). The anions HCrO4 − and CrO4 2− are the prevailing aqueous species of Cr(VI) at pH values below and above 6.5, respectively (Martell et al. 2004). In suboxic and anoxic systems, Cr(III) is the dominant form of chromium. The cation Cr3+ hydrolyzes strongly and tends to form sparingly soluble oxides and hydroxides as e.g., Cr(OH)3 which limits Cr3+ activities to very low values (≤10−9) at neutral pH (Baes and Mesmer 1976; Rai et al. 1987). In addition, Cr3+ strongly sorbs to mineral phases such as e.g., iron (hydr-)oxides (Fischer et al. 2007) and can substitute for Fe(III) in (hydr-)oxides (Frommer et al. 2010; Frommer et al. 2009). Therefore, dissolved Cr(III) concentrations are typically very low in aquatic systems and soils. As for the environmental concern of chromium, large differences between Cr(III) and Cr(VI) exist. Cr(III) represents an essential micronutrient. Cr(VI), however, is highly toxic and known to be carcinogenic and mutagenic (Katz and Salem 1994). As Cr(VI), being present as an anion, is generally much less immobilized by sorption to or incorporation in solid phases than is Cr(III), hexavalent chromium poses a high risk when released into the environment. For the remediation of Cr(VI) contaminated sites, technical solutions such as, e.g., the installation of reactive barriers containing zerovalent iron in contaminated groundwater zones have been shown to be effective (Flury et al. 2009). Reduction of Cr(VI) by aerobic bacteria has been proposed as an alternative remediation strategy for oxic environments (Cheng et al. 2010). The capability of reducing Cr(VI) to Cr(III) has been documented for a large variety of microbial species (Ackerley et al. 2004; Al Hasin et al. 2010; Bopp and Ehrlich 1988; Daulton et al. 2007; Fein et al. 2002; Fendorf et al. 2000; Lovley 1993; Lovley and Phillips 1994; Opperman et al. 2008; Park et al. 2000; Silver 1997; Suzuki et al. 1992; Viti et al. 2006; Viti et al. 2003). For an effective remediation of a highly contaminated site, however, bacteria applied to reduce Cr(VI) must exhibit a high resistance toward the toxicity of Cr(VI). As for many

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other metals, bacterial resistance to Cr(VI) is plasmid-based in most cases (Silver 1997). The ability to reduce Cr(VI) and to resist high Cr(VI) concentrations were found to be independent properties of bacteria (Bopp and Ehrlich 1988; Silver 1997). Bacteria may take up Cr(VI) primarily via the sulfate binding protein (SBP) of their membranes because of the similarity of both chromate and sulfate anions (Jacobson et al. 1991). In addition to anion efflux-mediated resistance, the discrimination between chromate and sulfate anions leading to a selective uptake of sulfate as well as an active regulation of the sulfate uptake mechanism in the presence of chromate were discussed as microbial responses to chromate exposure being linked to chromate resistance (Nies 2003; Viti et al. 2009). Inside the cell, Cr(VI) may readily react with the DNA causing adverse effects and with reduced compounds being associated with the membrane and present in the cytoplasm. Microbial Cr(VI) reduction counteracting adverse effects has been reported to follow various mechanisms. Reduction of Cr(VI) by a compound dissolved in the cytosol was shown for Bacillus sphaericus AND303 (Pal and Paul 2004). Pseudomonas fluorescens LB300 was found to reduce Cr(VI) enzymatically in the cell (Bopp and Ehrlich 1988). The bacterial species Bacillus subtilis, Sporosarcina ureae, and Shewanella putrefaciens were shown to reduce Cr(VI) under aerobic conditions at the outer surface of the cells in a nonmetabolic process (Fein et al. 2002). However, chemical oxidation of bacterial surfaces by Cr(VI) leads most likely to an irreversible damage of the integrity and functionality of the cell membranes. Alternative microbial Cr(VI) reduction mechanisms may include electron transfer to Cr(VI) in the periplasm and the excretion of Cr(VI) reducing agents. The reduction of Cr(VI) may not only represent a detoxification mechanism. Cr(VI) can also be used as a terminal electron acceptor as shown for Shewanella oneidensis under anaerobic conditions (Daulton et al. 2007). In addition, microbial Cr(VI) reduction is supposedly linked to further microbial processes. For instance, reduction of Cr(VI) in the cytoplasm and the periplasm may condition a mechanism to export the sparingly soluble Cr(III) to avoid a detrimental accumulation of Cr(III) in the cell (Bencheikh-Latmani et al. 2007). Among bacterial strains able to reduce Cr(VI), the strain Pseudomonas corrugata 28 was found to combine both a high Cr(VI) resistance and a high Cr(VI) reduction capability (Viti et al. 2006) making strain 28 a valuable candidate for remediation purposes. The minimum inhibitory concentration (MIC), which indicates the lowest concentration of a substance that inhibits bacterial growth, was reported to be as high as 40 mM Cr(VI) for P. corrugata 28. Growth of P. corrugata 28 and reduction of Cr(VI) can be decoupled as indicated by different growth and reduction rates when varying the carbon/energy source (Viti et al. 2007). Exposure of P. corrugata 28 to Cr(VI) was found to result in a strong and fast expression of the genes oscA (organosulfur compounds) and sbp (encoding sulfate ABC transporter periplasmic sulfate binding protein) forming a tran-

scriptional unit in P. corrugata 28 (Viti et al. 2009). Gene expression and Phenotype MicroArray analysis indicated that sulfur uptake was modulated when P. corrugata 28 was exposed to Cr(VI). However, it is still unknown to which extent changes in sulfate uptake of P. corrugata 28 affect its capability of reducing Cr(VI). In case of metabolic Cr(VI) reduction, a decrease of sulfate uptake may slow or stop metabolic processes, which in turn might also decrease and slow down Cr(VI) reduction. The ability to utilize sulfur sources other than sulfate may be a key feature to maintain metabolic processes in the presence of high Cr(VI) concentrations. P. corrugata 28 is able to utilize a large variety of sulfur sources including organosulfur compounds (Viti et al. 2009). However, the effect of utilizing organosulfur compounds for sulfur nutrition on Cr(VI) reduction has not been investigated. In addition, the fate of chromium after reduction by P. corrugata 28 has not been studied. Information on the speciation of chromium after reduction is crucial to assess the environmental fate of the microbially reduced chromium. The objectives of this study were (i) to investigate how variations of carbon and sulfur supply affect growth of P. corrugata 28 and Cr(VI) reduction and (ii) to study the fate of chromium after aerobic microbial reduction. In the first part, we aimed at elucidating whether Cr(VI) reduction by P. corrugata 28 was controlled by metabolic or nonmetabolic processes. A Tris minimal medium (TMM) containing varying concentrations of gluconate (carbon source) was used for the experiments. Furthermore, sulfate and ethanesulfonate were used as the sulfur sources to find out if the presence of Cr(VI) can induce sulfur starvation conditions. In the second part of this study, the formation of Cr(III) solid phases and Cr(III) adsorption to bacterial surfaces were investigated. Complementary X-ray absorption microscopy analysis of bacterial cells was conducted at the C K-edge and the Cr L2,3 -edge to further elucidate the fate of microbially reduced chromium. For chromium speciation and mass balance, dissolved Cr(VI) and total dissolved chromium, which also includes Cr(III), were measured.

MATERIALS AND METHODS Materials All chemicals mentioned in the following were analytical grade unless stated otherwise. Solutions and standards were prepared with high purity water (>18 M
cm, Milli-Q, Millipore). All glass and plastic labware used was thoroughly washed with ∼3% HCl and rinsed with high purity water prior to use. Bacterial Strain and Cultivation The bacterial strain Pseudomonas corrugata 28 was used in this study. P. corrugata 28 was isolated from a soil which was artificially polluted with up to 1000 mg Cr(VI) kg−1 soil (Viti et al. 2006). The Gram-negative wild-type strain 28 was shown

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would be transferred to chromium in case of complete Cr(VI) reduction. The reduction of Cr(VI) was strongly dependent on both carbon and sulfur supply (Figure 3). Small amounts of Cr(VI) corresponding to less than 5% of total Cr(VI) added were reduced in the absence of gluconate and sulfur supply within 24 h implying that P. corrugata 28 may possibly slowly reduce Cr(VI) in a nonmetabolic way. Abiotic reduction of Cr(VI) was quantitatively similar to reduction in experiments containing bacteria but no gluconate or sulfur source. Consequently, the small decrease of Cr(VI) in the absence of carbon and sulfur supply most likely includes slow abiotic Cr(VI) reduction occurring in the experiments in addition to microbial reduction. Aerobic nonmetabolic Cr(VI) reduction at the outer surface of bacterial cells was shown for Bacillus subtilis, Sporosarcina ureae, and Shewanella putrefaciens (Fein et al. 2002). Nonmetabolic reduction was faster for Sporosarcina ureae and Shewanella putrefaciens as compared to Bacillus subtilis and was found to strongly decrease with increasing pH as shown for Bacillus subtilis (Fein et al. 2002). At pH 7, only a small fraction of Cr(VI) was reduced in a nonmetabolic way despite high cell concentrations (∼7% of 0.1 mM Cr(VI) within 4 h with 12 g bacteria L−1 and ∼1.5% of 0.06 mM Cr(VI) within 24 h with 1.2 g bacteria L−1). In our nutrient-absent experiments, the concentration of P. corrugata 28 biomass was only ∼40 mg L−1, but cells were exposed to 0.2 mM Cr(VI) at pH 7. Assuming nonmetabolic Cr(VI) reduction kinetics of P. corrugata 28 surfaces at pH 7 to be of similar magnitude as reported by Fein et al. (2002), the extent of nonmetabolic Cr(VI) reduction is negligible in our growth and reduction experiments. Comparison of bacterial growth and Cr(VI) reduction results shows that Cr(VI) reduction was clearly growth-phase dependent (Figure 4). This result is consistent with previous investigations (Viti et al. 2007) and demonstrates that Cr(VI) reduction of P. corrugata 28 at pH 7 in TMM was clearly dominated by metabolic reduction. Furthermore, the marginal abiotic reduction of Cr(VI) in spent TMM containing excreted compounds (Figure 3d) and the similar excretion of carbon compounds by P. corrugata 28 in the presence and absence of Cr(VI) (Figure 5b) suggest that excretion of organic compounds in order to reduce Cr(VI) in the extracellular space may play –if any– a minor role for microbial reduction of Cr(VI) by P. corrugata 28 in our experiments. Comparison of microbial Cr(VI) reduction using different sulfur sources reveals that Cr(VI) reduction by P. corrugata 28 was similar, irrespectively whether sulfur was supplied as sulfate or ethanesulfonate (Figures 3 and 4). It was shown previously that exposure of P. corrugata 28 to Cr(VI) caused a strong and fast expression of the genes oscA (organosulfur compounds) and sbp forming a transcriptional unit (Viti et al. 2009). The expression of sbp may lead to a modulation of sulfate uptake since sbp encodes the periplasmic sulfate binding protein (Sekowska et al. 2000; Viti et al. 2009). Due to the physicochem-

ical similarity of sulfate anions and chromate anions, chromate is transported into the cell via the sulfate transport system which may not be able to discriminate both ions (Jacobson et al. 1991). Therefore, it is expected that the concomitant presence of sulfate and Cr(VI) would lead to sulfur starvation conditions and increased uptake and reduction of Cr(VI). Our experimental results, however, showed that P. corrugata 28 reduced a similar amount of Cr(VI) in case of sulfate and ethanesulfonate supply when cells were exposed to 0.2 mM Cr(VI). A plausible explanation for our experimental results is that Cr(VI) reduction by P. corrugata 28 is primarily linked to the expression of oscA involved in organosulfur cell homeostasis. Reduced organosulfur compounds can act as reductants for Cr(VI). Chromate is a strong oxidant having a high standard reduction potential E0 of +0.55 V at pH 7, whereas reduced organosulfur compounds exhibit low standard reduction potentials, e.g., E0 = –0.39 V for cysteine at pH 7 (Schwarzenbach et al. 2003). Immediate abiotic reduction of Cr(VI) by cysteine in standard TMM was also verified experimentally (not shown). We conclude from our growth and reduction experiments that Cr(VI) reduction of P. corrugata 28 is regulated by an effective intracellular reduction mechanism rather than by a modulation of Cr(VI) transport coupled with extracellular or surface-mediated reduction. Reduction of Cr(VI) to Cr(III) strongly affects the aqueous speciation of chromium as chromium is transformed from an anionic species (CrO4 2−) into a cationic species (Cr3+). Trivalent chromium strongly hydrolyzes and in contrast to Cr(VI), Cr(III) forms sparingly soluble (hydr-)oxides such as e.g., Cr(OH)3 (Baes and Mesmer 1976; Martell et al. 2004; Rai et al. 1987). Therefore, the reduction of Cr(VI) to Cr(III) may potentially lead to a formation of Cr(III)-phases either in the cytoplasm or in the extracellular space. The formation of biogenic minerals in the vicinity of bacterial surfaces and even inside cells has been reported for various metal-microbe systems (Bazylinski and Moskowitz 1997; Fortin et al. 1997; Hunter et al. 2008). Based on the hydrolysis of Cr(III), low concentrations of total dissolved Cr(III) are expected in noncomplexing media at pH 7. Soluble Cr(III) was found to amount to 0.7 ± 0.5 10−6 M in 50 mM sodium nitrate solutions after 24 days (Figure 6) indicating that nitrate was complexing Cr(III) only weakly. Similarly, soluble Cr(III) concentrations are limited to 10−6–10−7 M at pH 7 in other noncomplexing media such as perchlorate and chloride solutions (Martell et al. 2004; Rai et al. 1987). For TMM, the solubility of Cr(III) cannot be predicted reliably as stability constants for Cr(III)-complexation are not available for all potential ligands initially present in the medium. In addition, the composition of the medium changed during the bacterial growth as indicated by the decrease of gluconate (Figure 5a). Furthermore, the chemical structure and the concentration of excreted organic compounds, which also complex Cr(III), are unknown. The results of our solubility experiments carried out at pH 7 with TMM showed that the

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AEROBIC REDUCTION OF CHROMIUM(VI) Bencheikh-Latmani R, Obraztova A, Mackey MR, Ellisman MH, Tebo BM. 2007. Toxicity of Cr(III) to Shewanella sp. strain MR-4 during Cr(VI) reduction. Environ Sci Technol 41:214–220. Benzerara K, Menguy N, L´opez-Garc´ıa P, Yoon T-H, Kazimierczak J, Tyliszczak T, Guyot F, Brown GE, Jr. 2006. Nanoscale detection of organic signatures in carbonate microbialites. Proc Nation Acad Sci 103:9440–9445. Biebl H, Pfennig N. 1981. Isolation of members of the family Rhodospirillaceae. In: Starr MP, Stolp J, Truper HG, Balows A, Schlegel HG, editors. The Prokaryotes A Handbook on Habitats, Isolation and Identification of Bacteria. Berlin: Springer. P 267–273. Boese J, Osanna A, Jacobsen C, Kirz J. 1997. Carbon edge XANES spectroscopy of amino acids and peptides. J Electron Spectrosc Related Phenom 85:9– 15. Bopp LH, Ehrlich HL. 1988. Chromate resistance and reduction in Pseudomonas fluorescens strain LB300. Arch Microbiol 150:426–431. Cheng Y, Yan F, Huang F, Chu W, Pan D, Chen Z, Zheng J, Yu M, Lin Z, Wu Z. 2010. Bioremediation of Cr(VI) and immobilization as Cr(III) by Ochrobactrum anthropi. Environ Sci Technol 44:6357–6363. Claessens J, van Lith Y, Laverman AM, Van Cappellen P. 2006. Acid–base activity of live bacteria: implications for quantifying cell wall charge. Geochim Cosmochim Acta 70:267–276. Daulton TL, Little BJ, Jones-Meehan J, Blom DA, Allard LF. 2007. Microbial reduction of chromium from the hexavalent to divalent state. Geochim Cosmochim Acta 71:556–565. Dynes JJ, Tyliszczak T, Araki T, Lawrence JR, Swerhone GW, Leppard GG, Hitchcock AP. 2006. Speciation and quantitative mapping of metal species in microbial biofilms using scanning transmission X-ray microscopy. Environ Sci Technol 40:1556–1565. Fein JB, Fowle DA, Cahill J, Kemner K, Boyanov M, Bunker B. 2002. Nonmetabolic reduction of Cr(VI) by baterial surfaces under nutrient-absent conditions. Geomicrobiol J 19:369–382. Fendorf S, Wielinga B, Hansel CM. 2000. Chromium transformations in natural environments: The role of biological and abiological processes in chromium(VI) reduction. Intern Geol Rev 42:691–701. Fischer L, Br¨ummer GW, Barrow NJ. 2007. Observation and modelling of the reactions of 10 metals with goethite: Adsorption and diffusion processes. Eur J Soil Sci 58:1304–1315. Flury B, Frommer J, Eggenberger U, M¨ader U, Nachtegaal M, Kretzschmar R. 2009. Assessment of long-term performance and chromate reduction mechanisms in a field scale permeable reactive barrier. Environ Sci Technol 43:6786–6792. Fortin D, Ferris FG, Beveridge TJ. 1997. Surface-mediated mineral development by bacteria. In: Banfield JF, Nealson KH, editors. Geomicrobiology: Interactions between Microbes and Minerals. Washington, DC: Mineralogical Society of America. P 161–180. Frommer J, Nachtegaal M, Czekaj I, Kretzschmar R. 2010. The Cr Xray absorption near K-edge structure of poorly crystalline Fe(III)-Cr(III)oxyhydroxides. Am Mineral 95:1202–1213. Frommer J, Nachtegaal M, Czekaj I, Weng T-C, Kretzschmar R. 2009. X-ray absorption and emission spectroscopy of CrIII (hydr)oxides: Analysis of the K-pre-edge region. J Phys Chem A 113:12171–12178. Grolimund D, Trainor TP, Fitts JP, Kendelewicz T, Liu P, Chambers SA, Brown GE, Jr. 1999. Identification of Cr species at the aqueous solution-hematite interface after Cr(VI)-Cr(III) reduction using GI-XAFS and Cr L-edge NEXAFS. J Synchrotr Rad 6:612–614. Hunter RC, Hitchcock AP, Dynes JJ, Obst M, Beveridge TJ. 2008. Mapping the speciation of iron in Pseudomonas aeruginosa biofilms using scanning transmission X-ray microscopy. Environ Sci Technol 42:8766–8772. Jacobson BL, He JJ, Vermersch PS, Lemon DD, Quiocho FA. 1991. Engineered interdomain disulfide in the periplasmic receptor for sulfate transport reduced flexibility. J Biol Chem 266:5220–5225.

185

Jacobsen C, Wirick S, Flynn G, Zimba C. 2000. Soft x-ray spectroscopy from image sequences with sub-100nm spatial resolution. J Microsc 197:173–184. Katz SA, Salem H. 1994. The Biological and Environmental Chemistry of Chromium. New York, NY: VCH Publishers. Koch AL. 2007. Growth Measurement. In: Reddy CA, editor. Methods for General and Molecular Microbiology. Washington, DC: ASM Press. P 172–199. Lerotic M, Jacobsen C, Sch¨afer T, Vogt S. 2004. Cluster analysis of soft X-ray spectromicroscopy data. Ultramicrosc 100:35–57. Lovley DR. 1993. Dissimilatory metal reduction. Annu Rev Microbiol 47:263–290. Lovley DR, Phillips EJP. 1994. Reduction of chromate by Desulfovibrio vulgaris and its c3 cytochrome. Appl Environ Microbiol 60:726–728. Martell AE, Smith RM, Motekaitis RJ. 2004. NIST Critically Selected Stability Constants of Metal Complexes. Gaithersburg, MD: National Institute of Standards and Technology. Mergeay M, Nies D, Schlegel HG, Gerits J, Charles P, Van Gijsegem F. 1985. Alcaligenes eutrophus CH34 is a facultative chemolithotroph with plasmidbound resistance to heavy metals. J Bacteriol 162:328–334. Nies DH. 2003. Efflux-mediated heavy metal resistance in prokaryotes. FEMS Microbiol Rev 27:313–339. Opperman DJ, Piater LA, van Heerden E. 2008. A novel chromate reductase from Thermus scotoductus SA-01 related to old yellow enzyme. J Bacteriol 190:3076–3082. Pal A, Paul AK. 2004. Aerobic chromate reduction by chromate-resistant bacteria isolated from serpentine soil. Microbiol Res 159:347–354. Park CH, Keyhan M, Wielinga B, Fendorf S, Matin A. 2000. Purification to homogeneity and characterization of a novel Pseudomonas putida chromate reductase. Appl Environ Microbiol 66:1788–1795. Pilkington ES, Smith PR. 1967. The spectrophotometric determination of chromium in ilmenite. Anal Chim Acta 39:321–328. Rai D, Sass BM, Moore DA. 1987. Chromium(III) hydrolysis constants and solubility of chromium(III) hydroxide. Inorg Chem 26:345–349. Schwarzenbach RP, Gschwend PM, Imboden DM. 2003. Environmental Organic Chemistry. New York: Wiley-Interscience. Sekowska A, Kung H-F, Danchin A. 2000. Sulfur metabolism in Escherichia coli and related bacteria: Facts and fiction. J Mol Microbiol Biotechnol 2:145–177. Silver S. 1997. The bacterial view of the periodic table: Specific functions for all elements. In: Banfield JF, Nealson KH, editors. Geomicrobiology: Interactions between Microbes and Minerals. Washington, D.C.: Mineralogical society of America. P 345–360. Suzuki T, Miyata N, Horitsu H, Kawai K, Takamizawa K, Tai Y, Okazaki M. 1992. NAD(P)H-dependent chromium(VI) reductase of Pseudomonas ambigua G-1: A Cr(V) intermediate is formed during the reduction of Cr(VI) to Cr(III). J Bacteriol 174:5340–5345. Toner B, Fakre S, Villalobos M, Warwick T, Sposito G. 2005. Spatially resolved characterization of biogenic manganese oxide production within a bacterial biofilm. Appl Environ Microbiol 71:1300–1310. Viti C, Decorosi F, Mini A, Tatti E, Giovannetti L. 2009. Involvement of the oscA gene in the sulphur starvation response and Cr(VI) resistance in Pseudomonas corrugata 28. Microbiol 155:95–105. Viti C, Decorosi F, Tatti E, Giovannetti L. 2007. Characterization of chromateresistant and -reducing bacteria by traditional means and by a highthroughput phenomic technique for bioremedation purposes. Biotechnol Prog 23:553–559. Viti C, Mini A, Ranalli G, Lustrato G, Giovannetti L. 2006. Response of microbial communities to different doses of chromate in soil microcosms. Appl Soil Ecol 34:125–139. Viti C, Pace A, Giovannetti L. 2003. Characterization of Cr(VI)-resistant bacteria isolated from contaminated soil by tannery activity. Curr Microbiol 46:1–5.

RESEARCH ARTICLES

Uptake and removal of toxic Cr(VI) by Pseudomonas aeruginosa: physico-chemical and biological evaluation Suparna Chatterjee, Indranil Ghosh and Kalyan K. Mukherjea* Department of Chemistry, Jadavpur University, Kolkata 700 032, India

The present study evaluates the biosorption of Cr(VI) by Pseudomonas aeruginosa from synthetic solution and tannery effluents. The absorption was studied under different initial Cr(VI) concentrations at different pH values and in the presence of other metals. The Cr(VI) concentration in the effluent, sludge and soil of tannery industries was measured. A maximum absorption was found at 30 mg/l of Cr(VI) at pH 8, which decreased in the presence of cadmium. Cyclic voltammogram confirmed the reduction of Cr(VI). FTIR analysis showed that the carboxyl and amino groups on the bacterial surface bind chromium. SEM and EDX revealed that Cr(VI) is reduced to Cr(III). Keywords: Bioremediation, metal adsorption, Pseudomonas aeruginosa, tannery effluents. CHROMIUM compounds are being widely used in leather tanning, steel production, as metal corrosion inhibitor, alloy formation, in paints as pigments and various other applications1. Chromium thus is a contaminant in the soil, sediment, surface and groundwater in the trivalent and hexavalent forms2. Hence chromium-associated pollution is a cause of great concern. Chromium is an essential trace element for living organisms3. Of all the oxidation states of chromium, Cr(III) and Cr(VI) occur most commonly. Cr(VI) induces toxicity as it causes mutation4 and cancer5 in animals and mutation in bacteria. It is toxic even at a concentration of 20 μg/l (ref. 6). Heavy metal removal by chemical precipitation7, ion exchange, reverse osmosis and solvent extraction has disadvantages due to high cost and energy of complex processes8. On the other hand, bioremediation has advantages like the possibility of metal recovery, easy waste disposal of the incineration process and low cost. Metal uptake by microorganisms is an environment-friendly alternative9 of heavy metal remediation. Both living and dead microbial mass are capable of taking up metal ions from aqueous solution10. Microorganisms take up metal ions either actively (bioaccumulation) and/or passively (biosorption)11. Biosorption is the ability of biological materials to accumulate heavy

*For correspondence. (e-mail: [email protected]) CURRENT SCIENCE, VOL. 101, NO. 5, 10 SEPTEMBER 2011

metals from waste water through metabolically mediated or physico-chemical pathways of uptake12. It is based on mechanisms such as complexation, ion exchange, adsorption, chelation and micro precipitation10. Removal of heavy metals using different biosorbents has been found to be highly selective depending on the typical binding profile of the biosorbents13. Successful application of biosorption depends on the parameters like initial metal concentration and contact time14. In this present study a Gram-negative, ubiquitous, aerobic rod, Pseudomonas aeruginosa, has been used to assess the removal of chromium with a view to provide environment-friendly methods of removal of toxic chromium Cr(VI) from industrial effluents. The cell wall of this bacterium is composed of peptidoglycan, teichoic acids along with carboxyl, phosphoryl, hydroxyl and amino functional groups at the surface15. Fourier transform infrared (FTIR) analysis was carried out to determine the involvement of the type of functional groups in metal adsorption. Fein et al.16 reported that some components of the cell wall serve as electron donors for the reduction reaction while metal binding16.

Materials and methods All the chemicals were either AR or GR grade. All-glass triple-distilled water (TDW) was used throughout the study.

Biosorption studies The parameters responsible for removal such as time of contact, initial metal concentration, pH of culture media and other interfering metal ions like cadmium and iron17 which are normally present in tannery effluents have been studied. This study has been performed both by supplementing synthetic solution of Cr(VI) and also treating Cr(VI) from the effluents of tannery industries in and around Kolkata, India. P. aeruginosa was cultured and maintained in Luria Bertani (LB) agar plates and slants in our laboratory. The organism was grown and cultured aerobically with agitation at 37°C in the presence of Cr(VI) in LB broth 645

RESEARCH ARTICLES 13. Knauer, K., Behra, R. and Sigg, L., Adsorption and uptake of copper by the green alga Scenedesmus subspicatus (Chlorophyta). J. Phycol., 1997, 33, 596–601. 14. Ahalya, N., Kanamadi, R. D. and Ramachandra, T. V., Biosorption of chromium (VI) from aqueous solutions by the husk of Bengal gram (Cicer arientinum). Electron. J. Biotechnol., 2005, 8, 258–264. 15. Beveridge, T. J., Role of cellular design in bacterial metal accumulation and mineralization. Annu. Rev. Microbiol., 1989, 43, 147–171. 16. Fein, J. B., Fowle, D. A., Cahill, J., Kemner, K., Boyanov, M. and Bunker, B., Nonmetabolic reduction of chromium (VI) by bacterial surface under nutrient absent conditions. Geomicrobiol. J., 2002, 19, 369–382. 17. Deepali, G. K. K., Metals concentration in textile and tannery effluents, associated soils and groundwater. N.Y. Sci. J., 2010, 3, 82–89. 18. Clesceri, L. S., Greenberg, A. E. and Eaton, A. D., Standard Methods for the Examination of Water and Wastewater, American Public Health Association, Washington DC, 1998, 20th edn. 19. Tunali, S., Kiran, I. and Akar, T., Chromium (VI) biosorption characteristics of Neurospora crassa fungal biomass. Miner. Eng., 2005, 18, 681–689. 20. Talapatra, S. N. and Banerjee, S. K., Detection of micronucleus and abnormal nucleus in erythrocytes from the gill and kidney of Labeo bata cultivated in sewage-fed fish farms. Food Chem. Toxicol., 2007, 45, 210–215. 21. Hamil, H. W., Williams, R. R. and Mackay, C., Principles of Physical Chemistry, Prentice Hall, New Jersey, 1966, 2nd edn. 22. Smutok, O., Broda, D., Smutok, H., Dmytruk, K. and Gonchar, M., Chromate-reducing activity of Hansenula polymorpha recombinant cells over-producing flavocytochrome b2. Chemosphere, 2011, 83, 449–454. 23. Deng, S. and Ting, Y. P., Characterization of PEI-modified biomass and biosorption of Cu(II), Pb(II). Water Res., 2005, 39, 2167–2177. 24. Bai, R. S. and Abraham, T. E., Studies on enhancement of Cr(VI) biosorption by chemically modified biomass of Rhizopus nigricans. Water. Res., 2002, 36, 1224–1236. 25. Kapoor, A. and Viraraghvan, T., Heavy metal biosorption sites in Aspergillus niger. Bioresour. Technol., 1997, 61, 221–227. 26. Yee, N., Benning, L. G., Phoenix, V. R. and Ferris, F. G., Characterization of metal–Cyanobacteria sorption reactions: a combined macroscopic and infrared spectroscopic investigation. Environ. Sci. Technol., 2004, 38, 775–782.

652

27. Kang, S. Y., Lee, J. U. and Kim, K. W., Biosorption of Cr(III) and Cr(VI) on to the cell surface of Pseudomonas aeruginosa. Biochem. Eng. J., 2007, 36, 54–58. 28. Park, D., Yun, Y. S. and Park, J. M., Studies on hexavalent chromium biosorption by chemically treated biomass of Ecklonia sp. Chemosphere, 2005, 60, 1356–1364. 29. Gonzalez, C. F., Ackerley, D. F., Park, C. H. and Matin, A., A soluble flavoprotein contributes to chromate reduction and tolerance by Pseudomonas putida. Acta Biotechnol., 2003, 23, 233– 239. 30. Chai, L. Y., Huang, S. H., Yang, Z. H., Peng, B., Huang, Y. and Chen, Y. H., Hexavalent chromium reduction by Pannonibacter phragmitetus BB isolated from soil under chromium-containing slag heap. J. Environ. Sci. Health A, 2009, 44, 615–622. 31. Zakaria, Z. A., Zakaria, Z., Surif, S. and Ahmad, W. A., Hexavalent chromium reduction by Acinetobacter haemolyticus isolated from heavy metal contaminated waste water. J. Hazard. Mater., 2007, 146, 30–38. 32. McLean, J. and Beveridge, T. J., Chromate reduction by a pseudomonad isolated from a site contaminated with chromate copper arsenate. Appl. Environ. Microbiol., 2001, 67, 1076–1084. 33. Chattopadhyay, B., Chatterjee, A. and Mukhopadhyay, S. K., Bioaccumulation of metals in the East Calcutta wetland ecosystem. Aquat. Ecosyst. Health Manage., 2002, 5, 191–203. 34. Bidwell, A. M. and Dowdy, R. H., Cadmium and zinc availability to corn following termination of sewage sludge application. J. Environ. Qual., 1987, 16, 438–442. 35. DeFilippis, L. F. and Pallaghy, C. K., Heavy metals: sources and biological effects. In Advances in Limnology Series: Algae and Water Pollution (eds. Rai, L. C. and Soeder, C. J.), E. Scheizbartsche Press, Stuttgart, 1994, pp. 31–77. 36. Gowd, S. S. and Govil, P. K., Distribution of heavy metals in surface water of Ranipet industrial area in Tamil Nadu, India. Environ. Monit. Assess., 2008, 136, 197–207.

ACKNOWLEDGEMENTS. We thank the Labonya Prova Bose Trust, Kolkata for providing a fellowship to S.C. I.G. thanks Netaji Subhas Engineering College, Kolkata, for permission to continue his research (Ph D) at Jadavpur University, Kolkata. Partial financial assistance from the Centre of Advanced Studies (CAS), PURSE, DST, Department of Chemistry, Jadavpur University is acknowledged.

Received 28 March 2011; revised accepted 10 August 2011

CURRENT SCIENCE, VOL. 101, NO. 5, 10 SEPTEMBER 2011

Journal of Hazardous Materials 186 (2011) 756–764

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Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat

Adsorption profile of lead on Aspergillus versicolor: A mechanistic probing Himadri Bairagi a , Md. Motiar R. Khan b , Lalitagauri Ray a , Arun K. Guha b,∗ a b

Dept. of Food Technology & Biochemical Engineering, Jadavpur University, Kolkata 700032, India Dept. of Biological Chemistry, Indian Association for the Cultivation of Science, Kolkata 700032, India

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Article history: Received 16 July 2010 Received in revised form 11 November 2010 Accepted 16 November 2010 Available online 24 November 2010 Keywords: Aspergillus versicolor Lead Adsorption Binding mechanism Chemical modification

a b s t r a c t The adsorption of lead on Aspergillus versicolor biomass (AVB) has been investigated in aqueous solution with special reference to binding mechanism in order to explore the possibilities of the biomass to address environmental pollution. AVB, being the most potent of all the fungal biomasses tested, has been successfully employed for reducing the lead content of the effluents of battery industries to permissible limit (1.0 mg L−1 ) before discharging into waterbodies. The results establish that 1.0 g of the biomass adsorbs 45.0 mg of lead and the adsorption process is found to depend on the pH of the solution with an optimum at pH 5.0. The rate of adsorption of lead by AVB is very fast initially attaining equilibrium within 3 h following pseudo second order rate model. The adsorption process can better be described by Redlich–Peterson isotherm model compared to other ones tested. Scanning electron micrograph demonstrates conspicuous changes in the surface morphology of the biomass as a result of lead adsorption. Zeta potential values, chemical modification of the functional groups and Fourier transform infrared spectroscopy reveal that binding of lead on AVB occurs through complexation as well as electrostatic interaction. © 2010 Elsevier B.V. All rights reserved.

1. Introduction Lead is a highly toxic metal and its exposure even at low concentration leads to adverse effects on human health particularly in the nervous and reproductive system as well as in kidneys, liver and brain [1,2]. The metal enters into environment as a result of various industrial activities such as electroplating, battery manufacturing, pigment and dye industries, lead smelting and using leaded petroleum engine fuels [3,4]. Because of toxic nature, the concentration of lead in the industrial effluents must be brought down to permissible limit (1.0 mg L−1 ) before discharging into public sewers as per instruction of Environmental Regulatory Authority, India. The available methodologies that are employed for the treatment of lead containing wastewater include precipitation with lime, ultrafiltration, reverse osmosis or ion exchange process [5,6]. However, these methods suffer from limitations like poor cost effectivity, incomplete precipitation, etc. In addition they generate huge amount of toxic metal bearing sludge difficult to dispose of. Thus research on the development of effective and inexpensive technology to treat toxic metal bearing wastewater is very important. ∗ Corresponding author at: Dept. of Biological Chemistry, Indian Association for the Cultivation of Science, 2A & B, Raja S.C. Mullick Road, Jadavpur, Kolkata, West Bengal 700032, India. Tel.: +91 33 2473 4971/5904x502; fax: +91 33 2473 2805. E-mail addresses: [email protected], [email protected], [email protected], [email protected] (A.K. Guha). 0304-3894/$ – see front matter © 2010 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2010.11.064

In recent years considerable attention has been focused on adsorption technology to remove metal ions from wastewater from the standpoint of eco-friendly, effective and economic considerations. Of the different adsorbents like saw dust, rice hulls, palm kernel husk, coconut husk, banana and orange peels, modified lignin, de-oiled allspice husk and different agricultural by-products [7,8]; activated carbon, is the best. However, its use is much restricted due to high cost in many countries including India. This leads to search for efficient adsorbents preferably biological materials covering microbial biomass to remove metal ions from wastewater known as biosorption or bioaccumulation [9]. This method has certain advantages over conventional ones, e.g., non generation of toxic sludge, effective in reducing the concentration of metal ions below the permissible limit and the possibility of recovery of the biomaterials by regeneration. Thus it provides an economic means for the treatment of wastewater [7]. The uptake of heavy metals by microbial biomass is a two step process involving initial cell surface binding followed by intracellular accumulation which takes place only in living cells [10]. Adsorption of metal ions on the biomass occurs through electrostatic, physical and chemical interactions with the functional groups present on the cell wall [11–13]. Volesky [14] has recently reviewed the state of art in the field of biosorption of heavy metals. However, only a few reports are available on fungal systems [15,16]. Fungal biomass has certain advantages over bacterial biomass in respect of processing and handling of the biomass. The fungal genera of Rhizopus and Penicillium [17,18] and the use of raw and pretreated Aspergillus niger

H. Bairagi et al. / Journal of Hazardous Materials 186 (2011) 756–764 Table 3 Removal of Pb (II) from industrial effluent by AVB. Concentration of Pb (II) in industrial effluent before treatment (mg L−1 )

Concentration of Pb (II) in industrial effluent after treatment (mg L−1 )

3.20 5.10

0.45 0.72

3.6. Reusability of AVB Desorption of lead from the lead adsorbed biomass to the extent of 85% can be achieved with 0.1 M HCl and the biomass can again be used for adsorption of lead. The adsorption–desorption cycle can be conducted for five times after which loss in activity was noted probably due to loss of integrity of the cell (data not shown). Thus we believe that adsorption involving AVB is a highly effective and efficient methodology for removing lead from polluted waterbodies. 3.7. Interaction of lead with AVB in industrial effluent The efficiency of AVB to remove lead was also executed using the effluent of battery industry as a feed solution. The concentration of lead in the effluent varied from 3.2 mg L−1 to 5.1 mg L−1 . The adsorption experiment was carried out after adjusting the pH at 5.0. It was observed that the percentage removal of lead from industrial effluent was 86% (Table 3). This value is very similar to that obtained from the monometallic system, suggesting that the presence of anions and cations present in the industrial effluent has no inhibitory effect on lead adsorption under the studied experimental condition. 4. Conclusion We demonstrate a viable option for the removal of lead from contaminated water with AVB. The maximum adsorption capacity of AVB has been found to be 45 mg Pb (II) per gram of the dry weight of the biomass. The Redlich–Peterson isotherm model describes the adsorption process satisfactorily suggesting that the adsorption mechanism is a hybrid one and does not follow ideal monolayer adsorption and the possibility of multilayer adsorption. Scatchard plot analysis reveals multiple and non equivalent binding sites on the AVB cell surface. The adsorption process is very fast initially and more than 80% is completed within 60 min. FTIR study and chemical modifications of biomass cell surface suggest the major involvement of carboxyl functional groups in the adsorption process. Thus it may be summarized that AVB can remove lead from its aqueous solution successfully. Acknowledgements One of the authors (Mr. H. Bairagi) is thankful to the University Grants Commission, New Delhi, India for awarding the Rajiv Gandhi National Fellowship. The authors also gratefully acknowledge Mr. D. Halder (Dept. of Food Technology and Biochemical Engineering, Jadavpur University, Kolkata), Mr. S. Majhi and Mr. S. Naskar (Dept. of Material Science, IACS, Kolkata), Dr. R. Chakravarty and Mr. A. Ghosh (Dept. of Biological Chemistry, IACS, Kolkata), Dr. S.K. Das (Dublin City University, Ireland) and Dr. A.R. Das (Polymer Science Unit, IACS, Kolkata) for their kind help throughout the work. References [1] M. Nadeem, A. Mahmood, S.A. Shahid, S.S. Shah, A.M. Khalid, G. Mckay, Sorption of lead from aqueous solution by chemically modified carbon adsorbents, J. Hazard. Mater. B 138 (2006) 604–613.

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[2] V.K. Gupta, A. Rastogi, Biosorption of lead from aqueous solutions by green algae Spirogyra species: kinetics and equilibrium studies, J. Hazard. Mater. 152 (2008) 407–414. [3] C. Raji, T.S. Anirudhan, Chromium (VI) adsorption by sawdust carbon: kinetics and equilibrium, Indian J. Chem. Technol. 4 (1997) 228–236. [4] A. Ornek, M. Ozacar, I.A. Sengil, Adsorption of lead onto formaldehyde or sulphuric acid treated acorn waste: equilibrium and kinetic studies, Biochem. Eng. J. 37 (2007) 192–200. [5] K. Jayaram, I.Y.L.N. Murthy, H. Lalhruaitluanga, M.N.V. Prasad, Biosorption of lead from aqueous solution by seed powder of Strychnos potatorum L., Colloids Surf. B: Biointerfaces 71 (2009) 248–254. [6] W.Y. Baik, J.H. Bae, K.M. Cho, W. Hartmeier, Biosorption of heavy metals using whole mold mycelia and parts thereof, Bioresour. Technol. 81 (2002) 167–170. [7] A. Demirbas, Heavy metal adsorption onto agro-based waste materials: a review, J. Hazard. Mater. 157 (2008) 220–229. [8] J. Cruz-Olivares, C. Pérez-Alonso, C. Barrera-Díaz, G. López, P. Balderasernández, Inside the removal of lead(II) from aqueous solutions by de-oiled allspice husk in batch and continuous processes, J. Hazard. Mater. 181 (2010) 1095–1101. [9] F. Veglio, F. Beolchini, Removal of metals by biosorption: a review, Hydrometallurgy 44 (1997) 301–316. [10] J.M.C. Tobin, C. White, G.M. Gadd, Metal accumulation by fungi: applications in environmental biotechnology, J. Ind. Microbiol. Biotechnol. 13 (1994) 126–130. [11] P.G.G. Burnett, C.J. Daughney, D. Peak, Cd adsorption onto Anoxybacillus flavithermus: surface complexation modeling and spectroscopic investigations, Geochim. Cosmochim. Acta 70 (2006) 5253–5269. [12] J.B. Fein, D.A. Fowle, J. Cahill, K. Kemner, M. Boyanov, B. Bunker, Nonmetabolic reduction of Cr(VI) by bacterial surfaces under nutrient-absent conditions, Geomicrobiol. J. 19 (2002) 369–382. [13] A.Y. Dursun, A comparative study on determination of the equilibrium, kinetic and thermodynamic parameters of biosorption of copper (II) and lead (II) ions onto pretreated Aspergillus niger, Biochem. Eng. J. 28 (2006) 187–195. [14] B. Volesky, Biosorption and me, Water Res. 41 (2007) 4017–4029. [15] J.N.L. Latha, K. Rashmi, M.P. Maruthi, Cell-wall-bound metal ions are not taken up in Neurospora crassa, Can. J. Microbiol. 51 (2005) 1021–1026. [16] J. Wang, C. Chen, Biosorption of heavy metals by Saccharomyces cerevisiae: a review, Biotechnol. Adv. 24 (2006) 427–451. [17] Y.S. Saˇg, T. Kutsal, Determination of the biosorption heats of heavy metal ions on Zoogloea ramigera and Rhizopus arrhizus, Biochem. Eng. J. 6 (2000) 145–151. [18] T. Fan, Y. Liu, B. Feng, G. Zeng, C. Yang, M. Zhou, H. Zhou, Z. Tan, X. Wang, Biosorption of cadmium(II), zinc(II) and lead(II) by Penicillium simplicissimum: isotherms, kinetics and thermodynamics, J. Hazard. Mater. 160 (2008) 655–661. [19] W. Jianlong, Z. Xinmin, D. Decai, Z. Ding, Biosorption of lead (II) from aqueous solution by fungal biomass of Aspergillus niger, J. Biotechnol. 87 (2001) 273–277. [20] M. Amini, H. Younesi, N. Bahramifar, A.A.Z. Lorestani, F. Ghorbani, A. Daneshi, M. Sharifzadeh, Application of response surface methodology for optimization of lead biosorption in an aqueous solution by Aspergillus niger, J. Hazard. Mater. 154 (2008) 694–702. [21] A. Kapoor, T. Viraraghavan, D.R. Cullimore, Removal of heavy metals using the fungus Aspergillus niger, Bioresour. Technol. 70 (1999) 95–104. [22] S.K. Das, A.K. Guha, Biosorption of chromium by Termitomyces clypeatus, Colloids Surf. B: Biointerfaces 60 (2007) 46–54. [23] S.K. Das, A.K. Guha, Biosorption of hexavalent chromium by Termitomyces clypeatus biomass: kinetics and transmission electron microscopic study, J. Hazard. Mater. 167 (2009) 685–691. [24] S. Majumdar, S.K. Das, T. Saha, G.C. Panda, T. Bandyopadhyou, A.K. Guha, Adsorption behavior of copper ions on Mucor rouxii biomass through microscopic and FTIR analysis, Colloids Surf. B: Biointerfaces 63 (2008) 138–145. [25] S.K. Das, A.R. Das, A.K. Guha, A study on the adsorption mechanism of mercury on Aspergillus versicolor biomass, Environ. Sci. Technol. 41 (2007) 8281–8287. [26] G.M. Loudon, Organic Chemistry, First ed., Reading, MA, USA, Addison-Wesley, 1984. [27] A. Markowska, J. Olejnik, J. Michalski, Selektive Alkylierung von mehrbasigen Sa¨uren des 4-bindigen Phosphors mit Trialkylphosphit, Chem. Ber. 108 (1975) 2589–2592. [28] L.F. Fieser, M. Fieser, Reagents for Organic Synthesis, vol. 1, Wiley, New York, 1967. [29] J. Gardea-Torresdey, M.K. Becker-Hapak, J.M. Hosea, D.W. Darnall, Effect of chemical modification of algal carboxyl groups on metal ion binding, Environ. Sci. Technol. 24 (1990) 1372–1378. [30] M. Sastri, A. Ahmad, M.I. Khan, R. Kumar, Biosynthesis of metal nanoparticles using fungi and actinomycete, Curr. Sci. 85 (2003) 162–170. [31] W. Lo, H. Chua, K.-H. Lam, S.-P. Bi, A comparative investigation on the biosorption of lead by filamentous fungal biomass, Chemosphere 39 (1999) 2723–2736. [32] J.D. Merifield, W.G. Davids, J.D. MacRae, A. Amirbahman, Uptake of mercury by thiol-grafted chitosan gel beads, Water Res. 38 (2004) 3132–3138. [33] P. Miretzky, M.C.B.M.F. Jardim, J.C. Rocha, Factors affecting Hg(II) adsorption in soils from the Rio Negro Basin (Amazon), Quim. Novo 28 (2005) 438–443. [34] S. Kraemer, J.G. Hering, Biogeochemical controls on the mobility and bioavailability of metals in soils and groundwater, Aquat. Sci. 66 (2004) 1–2. [35] G. Issabayeva, M.K. Aroua, N.M.N. Sulaiman, Removal of lead from aqueous solutions on palm shell activated carbon, Bioresour. Technol. 97 (2006) 2350–2355. [36] B. Xiao, K.M. Thomas, Adsorption of aqueous metal ions on oxygen and nitrogen functionalized nanoporous activated carbons, Langmuir 21 (2005) 3892–3902.

Chemical Engineering Journal 162 (2010) 122–126

Contents lists available at ScienceDirect

Chemical Engineering Journal journal homepage: www.elsevier.com/locate/cej

Long-term chromate reduction by immobilized fungus in continuous column Rashmi Sanghi ∗ , Ashish Srivastava 302 Southern Laboratories, Facility for Ecological and Analytical Testing, Indian Institute of Technology Kanpur, Kanpur, UP 208016, India

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Article history: Received 5 February 2010 Received in revised form 1 May 2010 Accepted 10 May 2010 Keywords: Thioester Chromium reduction Fungal biomass Column Immobilization

a b s t r a c t The immobilized fungus Coriolus versicolor was examined in a continuous fixed bed column for long-term Cr(VI) reduction at its physiological pH. The effects of operating parameters like flow rate, glucose concentration in the influent feed, COD, initial Cr(VI) concentration on the Cr(VI) reduction were investigated. Increase in the inlet Cr(VI) concentration and flow rate through the column led to a higher breakthrough of the Cr(VI) ions in the effluent. Cr(VI) reduction rate increased with increase in initial Cr(VI) concentration of up to 60 mg/L and thereafter showed a gradual decline. A Fourier transform infrared spectra were employed to elucidate the possible biosorption mechanism as well. The readiness of the thiol group of the fungal protein to interact with the Cr(VI) ion in addition to its strong reducing ability makes it a particularly important entity in the metabolism of Cr(VI). The possible role of thiol in the Cr(VI) reduction via the formation of Cr(VI) thioester is discussed. The study clearly exhibits the usage of live fungus for the long-term continuous removal of Cr(VI) as well as recovery of the metal ions from wastewater. © 2010 Elsevier B.V. All rights reserved.

1. Introduction Chromium has been widely recognized as a toxic mutagen [1] and a carcinogen yet is an important metal, which is used in a variety of industrial applications. Chromium is a metal that can exist in oxidation states from −2 to +6, The trivalent oxidation state is the most stable form of chromium In biological systems, chromium is naturally found in its trivalent state at very variable levels, whereas the hexavalent form is generally a derivative of man’s activities. Cr(VI) tends to associate with oxygen generating the powerful oxidants chromate (CrO4 2− ) and dichromate (Cr2 O7 2− ). The biological effects of chromium are highly dependent on the oxidation state. Derivatives of Cr(III) are water insoluble compared to Cr(VI) derivative compounds that are highly soluble [2]. Requirements of large quantity of chemicals or energy can be a limitation for the application of physicochemical methods for removing Cr(VI). Removal of heavy metal ions using biosorption could be a promising technology and has received more and more attention in recent years [3,4,5]. Microorganisms, which are capable of transforming metals from one oxidation state to another, facilitate detoxification and/or the removal of chromium, and have thus received recognition [6]. In the concept of biosorption, several chemical processes may be involved, such as bioaccumulation, bioadsorption, precipitation by H2 S production, ion exchange, and covalent binding with the biosorptive sites, including carboxyl,

∗ Corresponding author. Fax: +91 512 2597866. E-mail address: [email protected] (R. Sanghi). 1385-8947/$ – see front matter © 2010 Elsevier B.V. All rights reserved. doi:10.1016/j.cej.2010.05.011

hydroxyl, sulphydryl, amino and phosphate groups of the microorganisms [7]. The surfaces of fungal cells appear to act as ion exchange resins [8]. From the quantitative point of view, the surface sorption usually can contribute the larger proportion to total metal uptake, and thus binding to cell walls appears to be the most significant mechanism of sorption. Since it is energy independent, it occurs in both living and dead microbial biomass, including fungal mycelium [9]. The aim of the present work is to assess the long-term performance of thiol containing live fungus Coriolus versicolor for the continuous reduction of Cr(VI) in upflow fixed bed columns in a growth-supportive medium at physiological pH. The effects of some operating parameters, such as inlet Cr(VI) concentration, media composition, and flow rate were examined and optimized for the long-term Cr(VI)-reduction performance in the column. In this study we have used white-rot fungus C. versicolor as a bioreductant since it has a high growth rate and can grow under a variety of environmental conditions including low pH, high pollutant concentration. The possible mechanism for the reduction of Cr(VI) to Cr(III) by the fungus is also discussed. 2. Materials and methods 2.1. Reagents A stock solution of Cr(VI) was prepared containing 18.6736 g K2 Cr2 O7 per litre of deionized water. Sterilized stock Cr(VI) solution was added to sterile medium to a desired concentration of Cr(VI) with minimal dilution of the medium. All chemicals used were of AR grade.

R. Sanghi, A. Srivastava / Chemical Engineering Journal 162 (2010) 122–126

2.2. Microorganism and media A white-rot fungal strain, Coriolus versicolor, was obtained from Institute of Microbial Technology, Chandigarh, India. The strain was maintained at 4 ◦ C on malt agar slants. The liquid growth medium used for inoculating the fungus, consisted of 10 g/L glucose and 5 g/L malt extract (S.D Fine Chemicals, Mumbai). The growth medium used was always autoclaved (WidWo Cat. AVD 500 horizontal autoclave) at 15 psi for 30 min and cooled to room temperature before use. The inlet pH of feed containing Cr(VI) was 7.2. No adjustments in pH were made. In case of continuous flow column study, the media prepared every third day, had the following ingredients per litre of tap water: glucose 2 g, malt extract 1 g, peptone 0.5 g, KH2 PO4 2 g, MgSO4 1.023 g, CaCl2 0.1325 g and MnCl2 0.099 g. 2.3. Immobilization of fungus in column The columns were made of borosilicate glass with 2.2 cm ID, 31 cm height, and 22.5 cm bed length. The glass column was packed with ceramic beads on which the fungus was immobilized. To avoid channel effects the Cr(VI) solutions (10–80 mg/L) were continuously pumped upward through the column by a peristaltic pump (Mclin). The flow rates of the solutions were varied from 20 to 50 mL/h. The hydraulic residence time for the column reactor was 54.8 min. Initially the columns were conditioned by recycling funguscontaining growth media for a week followed by the feeding of fresh growth media until the COD reduction reached a steady state. Thereafter the influent growth medium was supplied with known concentration of metal solution. The conditions for maximum Cr(VI) reduction like influent Cr(VI) concentration, flow rate and glucose concentration in growth medium, were investigated and then subsequent analyses were performed under these optimized condition. Column effluent samples were collected at regular time intervals and COD, pH, Cr(total), Cr(VI), Cr(III), concentration were monitored. The COD measurements were made by closed reflux method according to the standard APHA procedure [12]. All experiments were performed in duplicates at room temperature.

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Table 1 Reduction of Cr(VI) ion at different flow rates (initial metal ion concentration 30 mg/L). Flow rate (mL/h)

mg Cr(VI) reduced/L h

20 25 30 35 40 45 50

11.51 11.13 11.55 11.72 8.88 6.67 5.41

± ± ± ± ± ± ±

0.4 0.5 0.5 0.7 0.5 0.4 0.2

2.6. Glucose analysis The determination of reducing sugar was done using 3,5dinitrosalicylic acid method [11]. The effect of glucose concentration on % Cr(VI) reduction was studied by varying the concentration from 0 to 5 g/L. Once the glucose concentration was optimized to 2 g/L, further experiments were carried out at this concentration only. The glucose consumed with time was also monitored and equated with the Cr(VI) reduction with time. 3. Results and discussion 3.1. Monitoring of column conditions The columns were run in duplicates. The results revealed the presence of both Cr(VI) and Cr(III) in the effluent solution after sorption of Cr(VI) on the biomass whereas on the fungal biomass, only Cr(III) was found to be present. A decrease in pH from 7.2 to 3.2 was observed during each Cr(VI) reduction cycle. In the first hour itself the pH reduced to 4.6 and thereafter the pH decrease was slow. Reduction in pH of medium is possibly due to the accumulation of organic acid metabolites [12]. With the decrease in solution pH, protonation of amine sites (NH2 ) of fungus increased favouring more electrostatic attraction of negative HCrO4 − ion yielding high removal of Cr(VI). Perusals of the literature had also reported the reduction of Cr(VI) to Cr(III) at acidic pH along with protons been consumed supported by Eq. (1) [13]:

2.4. Instrumental analysis

HCrO4 − + 7H+ + 3e− → Cr3+ + 4H2 O

Infrared (IR) spectra were recorded on a BRUCKER, VERTEX-70, and Infrared spectrophotometer making KBr pellets in reflectance mode. The pH of solutions was measured using a Digital pH-meter (MK VI Systronic). Spectrophotometric analysis was carried out on a Perkin Elmer, lambda-40 UV-VIS spectrophotometer with a 1 cm path length. For the SEM studies, samples of fungal biomass were coated under vacuum with a thin layer of gold and examined by scanning electron microscopy [FEI (QANTA 200)] at 10–17.5 kV with a tilt angle 45◦ .

Variation of flow rate though the column (shown in Table 1) revealed that maximum reduction occurs up to the range of 35 mL/h, above this flow rate chromate reduction decreases gradually due to lesser contact time between the metal ions and the biomass. The reduction rate of Cr(VI) was very fast initially; about 65% of the starting Cr(VI) (30 mg/L) was reduced within the first 2.8 h of the reaction. However, the residual concentration of Cr(VI) reached its minimum in 24 h. This rapid rate of Cr uptake by the immobilized fungus has a significant practical importance for applications in small reactor volumes, thus ensuring efficiency as well as economy.

E 0 = 1.33 V

(1)

2.5. Chromium analysis 3.2. Effect of glucose concentration on Cr(VI) reduction The analysis of Cr(VI) in solution was carried out by the diphenylcarbazide colorimetric method [10]. Diphenylcarbazide forms a red-violet complex selectively with Cr(VI). The color was fully developed after 15 min and the sample solutions were analysed using the UV-VIS spectrophotometer and the absorbance of the color was measured at 540 nm. The total chromium concentration was determined by oxidizing any trivalent chromium with potassium permanganate, followed by analysis as hexavalent chromium. Cr(III) was determined from the difference between total chromium and Cr(VI) concentrations. The instrument response was periodically checked with metal ion standard solutions.

Glucose concentration was varied from 0 to 5 g/L and it was observed that the reduction rate significantly increased from 0 to 2 g/L but became almost constant after 2 g/L. In the absence of glucose only sorption process takes place which results in poor reduction (Table 2). A glucose concentration of 2 g/L and above in media resulted in the best Cr(VI) reduction. These results suggested that glucose plays a vital role of carbon source in the reduction process. A comparison of quantitative uptake of glucose reveals that the rate and extent of metal uptake is significantly enhanced by the presence of glucose (Fig. 1). This enhanced metal removal capability may be related to an increase in the availability of energy and cel-

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4. Conclusions Use of C. versicolor for the very long-term Cr(VI) reduction ability in continuous upflow column represents a very potential successful strategy to bioremediate Cr(VI) toxic wastewaters in their natural habitat with extremes of weather conditions. Due to the surface immobilization of fungal hyphal biomass, the solute could easily pass through the highly porous matrix of ceramic beads for reaching the functional groups of the biomass. The main purpose of this paper is to establish the optimization parameters for the reduction of Cr(VI) with C. versicolor such that the process is sustainable for a very long term. The long-term potential of the column packed with fungal biomass for Cr(VI) detoxification was demonstrated very effectively. The fungus clearly is able to thrive, grow and successfully reduce toxic Cr(VI) to its less toxic form even under repeated exposure of Cr(VI) for more than a year. The results indicated that the packed columns with the immobilized living cells is more convenient in operation and economic in treatment compared with traditional methods as it could resolve the problem of blockage as well as recovery of the metal ions from wastewater. Acknowledgments The authors are thankful to International Foundation of Science (Sweden) for the financial support to carry out this work. References [1] J.E. Gruber, K.W. Jennette, Metabolism of the carcinogen chromate by rat liver microsomes, Biochem. Biophys. Res. Commun. 82 (1978) 700–770. [2] E.W. Cary, Chromium in air, soil and natural waters, in: S. Langard (Ed.), Biological and Environmental Aspects of Chromium, Elsevier, Amsterdam, 1982, pp. 49–64. [3] J. Wang, Biosorption of copper (II) by chemically modified biomass of Saccharomyces cerevisiae, Process Biochem. 3 (2002) 847–850.

[4] A.I. Zouboulis, M.X. Loukidou, K.A. Matis, Biosorption of toxic metals from aqueous solutions by bacteria strains isolated from metal-polluted soils, Process Biochem. 39 (2004) 909–916. [5] A.Y. Dursun, C. Uslu, Y. Cuci, Z. Aksu, Bioaccumulation of copper(II), lead(II) and chromium(VI) by growing Aspergillus niger, Process Biochem. 38 (2003) 1647–1651. [6] B. Volesky, Z.R. Holan, Biosorption of heavy metal, Biotechnol. Prog. 11 (1995) 235–250. [7] B. Krantz-Rulcker, B. Allard, J. Schunrer, Interactions between a soil fungus, Trichoderma harzianum, and IIb metals—adsorption to mycelium and production of complexing metabolites, Biometals 6 (1993) 223–230. [8] B. Krantz-Rulcker, E. Frandberg, J. Schunrer, Metal loading and enzymatic degradation of fungal cell walls and chitin, Biometals 8 (1995) 12–18. [9] N. Saglam, R. Say, A. Denizli, S. Patir, M.Y. Arica, Biosorption of inorganic mercury and alkylmercury species on to Phanerochaete chrysosporium mycelia, Process Biochem. 34 (1999) 725–730. [10] American Public Health Association, Standard Methods for the Analysis of Water and Wastewater, American Public Health Association, Washington, DC, 1998. [11] G.L. Miller, Use of dinitrosalicylic acid reagent for determination of reducing sugar, Anal. Chem. 31 (1959) 426–428. [12] J.T. Spadaro, M.H. Gold, V. Renganathan, Degradation of azo dyes by the lignindegrading fungus Phanerochaete chrysosporium, Appl. Environ. Microbiol. 58 (1992) 2397–2401. [13] J.B. Fein, D.A. Fowle, J. Cahill, K. Kemner, M. Boyanov, B. Bunker, Nonmetabolic reduction of Cr(VI) by bacterial surfaces under nutrient-absent conditions, Geomicrobiol. J. 19 (2002) 369–382. [14] R.S. Bai, T.E. Abraham, Studies on enhancement of Cr (VI) biosorption by chemically modified biomass of Rhizopus nigricans, Water Res. 36 (2002) 1224–1236. [15] D.W.J. Kwong, D.E. Pennington, Stoichiometry, kinetics and mechanism of chromium (VI) oxidation of l-cysteine at neutral pH, Inorg. Chem. 23 (1984) 2528–2532. [16] D.M.L. Goodgame, A.M. Joy, EPR study of Cr (V) and radical species produced in the reduction of Cr (VI) by ascorbate, Inorg. Chim. Acta 135 (1987) 115–118. [17] S. Cakir, E. Bicer, Interaction of cystein with Cr(VI) ion under UV irradiation, Bioelectrochemistry 679 (2005) 75–80. [18] S.E. Shumate, G.N. Standberg, The principles, applications and regulations of biotechnology in industry, agriculture and medicine, in: M. Moo-Young (Ed.), Comprehensive Biotechnology, Pergamon Press, 1983, p. 235. [19] P.H. Connett, K.E. Wetterhahn, Metabolism of the carcinogen chromate by cellular constituents, Struct. Bonding 54 (1983) 93–125. [20] W. Mazijrek, P.J. Nichols, B.O. West, Chromium (VI)-thioester formation in N,Ndimethylformamide, Polyhedron 10 (1991) 753–762.

APPLIED AND ENVIRONMENTAL MICROBIOLOGY, Apr. 2010, p. 2433–2438 0099-2240/10/$12.00 doi:10.1128/AEM.02792-09 Copyright © 2010, American Society for Microbiology. All Rights Reserved.

Vol. 76, No. 8

Immobilization of Cr(VI) and Its Reduction to Cr(III) Phosphate by Granular Biofilms Comprising a Mixture of Microbes䌤 Y. V. Nancharaiah,1,2 C. Dodge,1 V. P. Venugopalan,2 S. V. Narasimhan,2 and A. J. Francis1* Environmental Sciences Department, Brookhaven National Laboratory, Upton, New York 11973,1 and Water and Steam Chemistry Division, Bhabha Atomic Research Centre Facilities, Kalpakkam 603102, India2

We assessed the potential of mixed microbial consortia, in the form of granular biofilms, to reduce chromate and remove it from synthetic minimal medium. In batch experiments, acetate-fed granular biofilms incubated aerobically reduced 0.2 mM Cr(VI) from a minimal medium at 0.15 mM dayⴚ1 gⴚ1, with reduction of 0.17 mM dayⴚ1 gⴚ1 under anaerobic conditions. There was negligible removal of Cr(VI) (i) without granular biofilms, (ii) with lyophilized granular biofilms, and (iii) with granules in the absence of an electron donor. Analyses by X-ray absorption near edge spectroscopy (XANES) of the granular biofilms revealed the conversion of soluble Cr(VI) to Cr(III). Extended X-ray absorption fine-structure (EXAFS) analysis of the Cr-laden granular biofilms demonstrated similarity to Cr(III) phosphate, indicating that Cr(III) was immobilized with phosphate on the biomass subsequent to microbial reduction. The sustained reduction of Cr(VI) by granular biofilms was confirmed in fed-batch experiments. Our study demonstrates the promise of granular-biofilm-based systems in treating Cr(VI)-containing effluents and wastewater. aqueous environments. Bioflocs, the active biomass of activated sludge-process systems are transformed into dense granular biofilms in sequencing batch reactors (SBRs). As granular biofilms settle extremely well, the treated effluent is separated quickly from the granular biomass by sedimentation (9, 24). Previous work demonstrated that aerobic granular biofilms possess tremendous ability for biosorption, removing zinc, copper, nickel, cadmium, and uranium (19, 26, 31, 32, 40). However, no study has investigated the role of cellular metabolism of aerobically grown granular biofilms in metal removal experiments. Despite vast knowledge about biotransformation by pure cultures, very little is known about reduction and immobilization by mixed bacterial consortia (8, 12, 13, 16, 20, 31, 36). Our research explored, for the first time, the metabolically driven removal of Cr(VI) by microbial granules. The main aim of this study was to investigate Cr(VI) reduction and immobilization by mixed bacterial consortia, viz., aerobically grown granular biofilms. Such biofilm-based systems are promising for developing compact bioreactors for the rapid biodegradation of environmental contaminants (17, 24, 29). Accordingly, we investigated the microbial reduction of Cr(VI) by aerobically grown biofilms in batch and fed-batch experiments and analyzed the oxidation state and association of the chromium immobilized on the biofilms by X-ray absorption near edge spectroscopy (XANES) and extended X-ray absorption fine structure (EXAFS).

Chromium is a common industrial chemical used in tanning leather, plating chrome, and manufacturing steel. The two stable environmental forms are hexavalent chromium [Cr(VI)] and trivalent chromium [Cr(III)] (20). The former is highly soluble and toxic to microorganisms, plants, and animals, entailing mutagenic and carcinogenic effects (6, 22, 33), while the latter is considered to be less soluble and less toxic. Therefore, the reduction of Cr(VI) to Cr(III) constitutes a potential detoxification process that might be achieved chemically or biologically. Microbial reduction of Cr(VI) seemingly is ubiquitous; Cr(VI)-reducing bacteria have been isolated from both Cr(VI)-contaminated and -uncontaminated environments (6, 7, 23, 38, 39). Many archaeal/eubacterial genera, common to different environments, reduce a wide range of metals, including Cr(VI) (6, 16, 21). Some bacterial enzymes generate Cr(V) by mediating one-electron transfer to Cr(VI) (1, 4), while many other chromate reductases convert Cr(VI) to Cr(III) in a single step. Biological treatment of Cr(VI)-contaminated wastewater may be difficult because the metal’s toxicity potentially can kill the bacteria. Accordingly, to protect the cells, cell immobilization techniques were employed (31). Cells in a biofilm exhibit enhanced resistance and tolerance to toxic metals compared with free-living ones (15). Therefore, biofilm-based reduction of Cr(VI) and its subsequent immobilization might be a satisfactory method of bioremediation because (i) the biofilmbound cells can tolerate higher concentrations of Cr(VI) than planktonic cells, and (ii) they allow easy separation of the treated liquid from the biomass. Ferris et al. (11) described microbial biofilms as natural metal-immobilizing matrices in

MATERIALS AND METHODS Cultivation of aerobic granular sludge. Aerobic granular biofilms were grown in a 3-liter working-volume laboratory-scale sequencing batch reactor (SBR). SBR setup and operation details have been described previously (26, 27). The SBR was inoculated with seed sludge collected from the outlet of an aeration tank of an operating domestic wastewater treatment plant at Kalpakkam, India. The reactor was operated at room temperature (30 ⫾ 2°C) at a volumetric exchange ratio of 66% and a 6-h cycle, comprising 60 min of anaerobic static fill, 282 min of aeration, 3 min of settling, 10 min of effluent decantation, and 5 min of being idle. The SBR was fed with acetate-containing synthetic wastewater as discussed by Nancharaiah et al. (27). Granules, collected 2 months after the

* Corresponding author. Mailing address: Brookhaven National Laboratory, Environmental Sciences Department, Building 490A, Upton, NY 11973. Phone: (631) 344-4534. Fax: (631) 344-7303. E-mail: [email protected]. 䌤 Published ahead of print on 19 February 2010. 2433

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Received 18 November 2009/Accepted 8 February 2010

VOL. 76, 2010

GRANULAR BIOFILM Cr(VI) IMMOBILIZATION AND REDUCTION

2437

TABLE 1. EXAFS fit of Cr(III) with aerobic microbial granulesa Type of atom

N

R (Å)

␴2

⌬E0

F

Cr(III) phosphate Cr-O Cr-P

5.7 ⫾ 0.7 4.0 ⫾ 1.3

1.97 ⫾ 0.01 3.11 ⫾ 0.05

0.002 ⫾ 0.001 0.007 ⫾ 0.002

0.5 ⫾ 1.0 10.5 ⫾ 1.4

0.038

Bacterial granules Cr-O Cr-P

6.3 ⫾ 1.5 4.0 ⫾ 1.3

1.98 ⫾ 0.04 3.11 ⫾ 0.03

0.001 ⫾ 0.001 0.007 ⫾ 0.002

1.2 ⫾ 0.8 13.3 ⫾ 2.5

0.017

a

N, coordination number (number of atoms); R, interatomic distances; ␴2, disorder parameter; ⌬E0, energy shift; F, goodness-of-fit parameter.

ACKNOWLEDGMENTS This research was supported by the Department of Atomic Energy, Government of India, and in part by the Environmental Remediation Sciences Division, Office of Biological and Environmental Research,

Office of Science, U.S. Department of Energy under contract no. DE-AC02-98CH10886. Y.V.N. gratefully acknowledges the American Society for Microbiology for the Indo-US Visiting Research Professorship Award. We thank Avril D. Woodhead for editorial help. REFERENCES 1. Ackerley, D. F., C. F. Gonzalez, M. Keyhan, R. Blake II, and A. Matin. 2004. Mechanism of chromate reduction by the Escherichia coli protein, NfsA, and the role of different chromate reductases in minimizing oxidative stress during chromate reduction. Environ. Microbiol. 6:851–860. 2. Al Hasin, A., S. J. Gurman, L. M. Murphy, A. Perry, T. J. Smith, and P. H. E. Gardiner. 2010. Remediation of chromium(VI) by a methane-oxidising bacterium. Environ. Sci. Technol. 44:400–405. 3. APHA. 1995. Standard methods for the examination of water and wastewater, 19th ed. American Public Health Association, Washington, DC. 4. Barak, Y., D. F. Ackerley, C. J. Dodge, L. Banwari, C. C. Alex, A. J. Francis, and A. Matin. 2006. Analysis of novel soluble chromate and uranyl reductases and generation of an improved enzyme by directed evolution. Appl. Environ. Microbiol. 72:7074–7082. 5. Beun, J. J., A. Hendriks, M. C. M. van Loosdrecht, E. Morgenroth, P. A. Wilderer, and J. J. Heijnen. 1999. Aerobic granulation in a sequencing batch reactor. Water Res. 33:2283–2290. 6. Cervantes, C., J. Campos-García, S. Devars, F. Gutie´rrez-Corona, H. LozaTavera, J. C. Torres-Guzma ´n, and R. Moreno-Sa ´nchez. 2001. Interactions of chromium with microorganisms and plants. FEMS Microbiol. Rev. 25:335– 347. 7. Chardin, B., M.-T. Giudici-Orticoni, G. De Luca, B. Guigliarelli, and M. Bruschi. 2003. Hydrogenases in sulfate-reducing bacteria function as chromium reductase. Appl. Microbiol. Biotechnol. 63:315–321. 8. Chung, J., R. Nerenberg, and B. E. Rittmann. 2006. Bioreduction of soluble chromate using a hydrogen-based membrane biofilm reactor. Water Res. 40:1634–1642. 9. de Kreuk, M. K., J. J. Heijnen, and M. C. M. van Loosdrecht. 2005. Simultaneous COD, nitrogen, and phosphate by aerobic granular sludge. Biotechnol. Bioeng. 90:761–769. 10. Fein, J. B., D. A. Fowle, J. Cahill, K. Kemner, M. Boyanov, and B. Bunker. 2002. Nonmetabolic reduction of Cr(VI) by bacterial surfaces under nutrient-absent conditions. Geomicrobiol. J. 19:369–382. 11. Ferris, F. G., S. Schultze, T. C. Witten, W. S. Fyfe, and T. J. Beveridge. 1989. Metal interactions with microbial biofilms in acidic and neutral pH environments. Appl. Environ. Microbiol. 55:1249–1257. 12. Francis, A. J. 1998. Biotransformation of uranium and other actinides in radioactive wastes. J. Alloys Compd. 271–273:78–84. 13. Francis, A. J. 2007. Microbial mobilization and immobilization of plutonium. J. Alloys Compd. 444–445:500–505. 14. Ganguli, A., and A. K. Tripathi. 2002. Bioremediation of toxic chromium from electroplating effluent by chromate-reducing Pseudomonas aeruginosa A2Chr in two bioreactors. Appl. Microbiol. Biotechnol. 58:416–420. 15. Harrison, J. J., H. Ceri, and R. J. Turner. 2007. Multimetal resistance and tolerance in microbial biofilms. Nat. Rev. Microbiol. 5:928–938. 16. Horton, R. N., W. A. Apel, V. S. Thompson, and P. P. Sheridan. 25 January 2006, posting date. Low temperature reduction of hexavalent chromium by a microbial enrichment consortium and a novel strain of Arthrobacter aurescens. BMC Microbiol. 6:5. doi:10.1186/1471-2180-6-5. 17. Inizan, M., A. Freval, J. Cigana, and J. Meinhold. 2005. Aerobic granulation in a sequencing batch reactor (SBR) for industrial wastewater treatment. Water Sci. Technol. 52:335–343. 18. Kemner, K. M., S. D. Kelly, B. Lai, J. Maser, E. J. O’Loughlin, D. SholtoDouglas, Z. Cai, M. A. Schneegurt, Jr., C. F. Kulpa, and K. H. Nealson. 2004. Elemental and redox analysis of single bacterial cells by X-ray microbeam analysis. Science 306:686–687. 19. Liu, Y., S. F. Yang, S.-F. Tan, Y.-M. Lin, and J.-H. Tay. 2002. Aerobic granules: a novel zinc biosorbent. Lett. Appl. Microbiol. 35:548–551.

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mass. The microbial species composition of the granular sludge was not identified in the present study; nonetheless, the commencement of reduction of chromium immediately after exposure to Cr(VI) suggests that bacteria able to reduce chromium already were present in the granules (without prior enrichment); the lack of a delay demonstrates that the necessary enzymes are constitutively expressed. Seemingly, previous exposure to chromium and subsequent microbial enrichment are not prerequisites for successful bioreduction. This could be mainly due to the involvement of constitutive chromate reductases, thus corroborating the earlier observation of the rapid reduction of Cr(VI) by Pseudomonas putida unsaturated biofilms (32). Aerobic granular sludge cultivated in an SBR using acetate and lacking prior exposure to chromium efficiently reduced Cr(VI) from minimal media. Passive biosorption by the granular biomass was ruled out because Cr(VI) removal was negligible in the absence of a carbon source and by lyophilized granules. Analysis of chromium speciation by XANES further confirmed the bioreduction of Cr(VI) to Cr(III), thereby pointing to the involvement of cell metabolism. Nonmetabolic reduction of Cr(VI) to Cr(III) by bacterial surfaces under nonnutrient conditions has been reported by Fein et al. (10). In this study, no such reduction of Cr(VI) to Cr(III) was observed under nonnutrient conditions. EXAFS analyses revealed that the granular biofilm-bound Cr(III) occurs as Cr(III) phosphate. Earlier, Neal et al. (28) reported that only Cr(III) was bound to live Shewanella oneidensis cells. XANES and EXAFS analyses of a Cr(III)-laden biomass of nonliving seaweed, Ecklonia, were very similar to spectra from Cr(III) acetate (30). Kemner et al. (18) reported that the speciation of chromium associated with Pseudomonas fluorescens cells was consistent with association of Cr(III) with a phosphoryl functional group. A recent study showed reduction of Cr(VI) to Cr(III) by methane-oxidizing bacteria, a ubiquitous group of environmental bacteria (2). EXAFS analysis showed that Methylococcus capsulatus-associated chromium predominantly existed as Cr(III) and most likely associated with phosphate groups. EXAFS spectra of our Cr(III)-laden granular biomass revealed the presence of Cr(III)-phosphate after Cr(VI) reduction. Overall, our findings suggest the potential use of mixed microbial granules to bioremediate Cr(VI)-containing wastewater or industrial effluents.

Appl Biochem Biotechnol (2010) 160:2000–2013 DOI 10.1007/s12010-009-8716-7

Sonoassisted Microbial Reduction of Chromium Mathur Nadarajan Kathiravan & Ramalingam Karthick & Naggapan Muthu & Karuppan Muthukumar & Manickam Velan

Received: 21 February 2009 / Accepted: 12 July 2009 / Published online: 29 July 2009 # Humana Press 2009

Abstract This study presents sonoassisted microbial reduction of hexavalent chromium (Cr(VI)) using Bacillus sp. isolated from tannery effluent contaminated site. The experiments were carried out with free cells in the presence and absence of ultrasound. The optimum pH and temperature for the reduction of Cr(VI) by Bacillus sp. were found to be 7.0 and 37°C, respectively. The Cr(VI) reduction was significantly influenced by the electron donors and among the various electron donors studied, glucose offered maximum reduction. The ultrasound-irradiated reduction of Cr(VI) with Bacillus sp. showed efficient Cr(VI) reduction. The percent reduction was found to increase with an increase in biomass concentration and decrease with an increase in initial concentration. The changes in the functional groups of Bacillus sp., before and after chromium reduction were observed with FTIR spectra. Microbial growth was described with Monod and Andrews model and best fit was observed with Andrews model. Keywords Bacillus sp. . Sonolysis . Chromium reduction . Electron donors . Growth kinetics Nomenclature C0 initial chromium concentration (mg/l) KS half saturation constant (mg/l) KI inhibition constant (mg/l) [S] substrate concentration (mg/l) S speed (rpm) t time (min) T temperature (°C) μ specific growth rate (h−1) M. N. Kathiravan : R. Karthick : K. Muthukumar (*) : M. Velan Department of Chemical Engineering, A.C. College of Technology, Anna University, Chennai 600 025, India e-mail: [email protected] N. Muthu Department of Biotechnology, Holy Cross College, Trichy 620002, India

Appl Biochem Biotechnol (2010) 160:2000–2013

2001

Introduction Extensive use of chromium in industries such as leather tanning, metallurgical, electroplating etc., resulted in industrial wastes containing hexavalent chromium (Cr(VI)). Toxic Cr(VI) ions cause physical discomfort and sometimes life-threatening illness including irreversible damage to vital body system [1]. Compared to Cr(VI), Cr(III) is nontoxic and, due to its lower environmental mobility, exhibits limited environmental impact. For this reason, the reduction of Cr(VI) to Cr(III) remains as a primary method for the treatment of chromium containing wastes. The traditional chemical and electrochemical methods used for the reduction are expensive and generate large volume of sludge. Microbial techniques developed to treat chromium-contaminated wastewater were found to be economic and was first demonstrated by Romanenko and Korenkov [2], following that a wide diversity of chromium reducing bacteria (CRB) has been isolated. Prime Cr(VI) reducing microorganisms include Escherichia, Pseudomonas, Pantoea, Cellulomonas, Micrococcus, Staphylococcus, Achromobacter sp. strain Ch1, Ochrobactrum intermedium SDCr-5 , P. agglomerans SP1, Sporosarcina ureae, Shewanella putrefaciens, Leucobacter sp., and Exiguobacterium sp. [3–10]. Conventional methods for reduction of Cr(VI) from industrial wastewater include chemical reduction, ion exchange, electrocoagulation [11], electrochemical reduction [12], photo-reduction [13], bulk liquid membranes process [14], and reduction using iron particles [15]. To improve the chromium reduction efficiency, an improved method, which combines ultrasound and microbial reduction of Cr(VI) is presented. The integration of ultrasonic with biological reactions demonstrated that the sonication increases the mass transfer in aqueous solution [16, 17], which in turn enhances the bioavailability [18, 19]. Gli et al. [20] analyzed the effect of ultrasonic irradiation on the reduction of Hg(II) and achieved a maximum reduction of 94%. The scope of the present investigation was to study the microbial reduction of Cr(VI) with free cells in the presence and absence of ultrasound and to optimize the parameters which influence the Cr(VI) reduction.

Materials and Methods Microorganism The microorganism used in this study was isolated from tannery effluent contaminated site in Chennai, India. The soil sample (10% w/v) was inoculated with nutrient broth containing 100 mg/l of Cr(VI) and incubated at 37°C under controlled conditions. After incubation, 10 ml of the culture was serially diluted (10−3 to 10−8). Samples (0.1 ml) were withdrawn from 10−5 dilution (in which single colonies were observed) and were then transferred to nutrient agar plates containing Cr(VI). After 2 days of incubation at 37°C, the colonies were screened for their ability to survive in the chromium-amended agar plates. Potential isolates were inoculated with fresh nutrient broth and purified by streak plate technique. The isolate was identified by colony morphology, cell morphology, and biochemical tests and isolated bacterium, identified as Bacillus sp., reduced Cr(VI) effectively. Minimal Inhibitory Concentration The minimal inhibitory concentration (MIC) of Cr(VI) was determined by inoculating overnight-grown culture of bacterial isolate into freshly prepared agar plates containing different concentrations of Cr(VI) (50 to 600 mg/l) at pH 7.0 and 37°C

Appl Biochem Biotechnol (2010) 160:2000–2013

2013

3. Viamajala, S., Smith, W. A., Sani, R. K., Apel, W. A., Petersen, J. N., Neal, A. L. (2007). Bioresource Technology, 98, 612–622. 4. Badar, U., Ahmed, N., Beswick, A. J., Pattanapitpais, P., & Macaskie, L. E. (2000). Biotechnology Letters, 22, 829–836. 5. Saxena, D., Levin, R., & Firer, M. A. (2000). Water Science Technology, 42, 93–98. 6. Zhu, L., Li, W., & Dong, X. (2003). International Journal of Systems and Evolution Microbiology, 53, 1619–1623. 7. Sultan, S., & Hasnain, S. (2007). Bioresource Technology, 98, 340–344. 8. Chris, A. F., Anna, Y. O., & Bradley, M. T. (2000). Applied and Environmental Microbiology, 66(2), 543–548. 9. Fein, J. B., Fowle, D. A., Cahill, J., Kemner, K., Boyanov, M., & Bunker, B. (2002). Geomicrobiology Journal, 19, 369–382. 10. Sarangi, A., & Krishnan, C. (2007). Bioresource Technology, . doi:10.1016/j.biortech. 2007.08.059. 11. Heidmann, I., & Calmano, W. (2008). Separation and Purification Technology, 61, 15–21. 12. Lakshmipathiraj, P., Bhasar Raju, G., Basariya, R., Parvathy, S., & Prabhakar, S. (2008). Separation and Purification Technology, 60, 96–102. 13. Kleber, R. J., & Helr, G. R. (1992). Environmental Science Technology, 26, 307–312. 14. Ahmet, O., Safi, S. A., & Sirit, A. (2006). Journal of Membrane Science, 283, 448–455. 15. Shao-Feng, N., Yong, L., Xin-hua, X., & Zhang-hua, L. (2005). Journal of Zhejiang University - Science B, 6(10), 1022–1027. 16. Thomas, J. M., Yordy, J. R., Amador, J. A., & Alexander, M. (1986). Applied Environmental Microbiology, 52, 290–296. 17. Harms, H., & Zehnder, A. J. B. (1995). Applied Environmental Microbiology, 61, 27–33. 18. Reynolds, C., & Wills, E. D. (1974). International Journal Radiation Biology, 25(2), 113–120. 19. Reid. I. M., & Sikov. M.R., Interaction of Ultrasound and Biological Tissues, Proceedings of Workshop, Washington, 1972 20. Gil, S., Lavilla, I., & Bendicho, C. (2008). Ultrasonics Sonochemistry, 15(3), 212–216. 21. Pattanapipitpaisal, P., Brown, N. L., & Macaskie, L. E. (2001). Applied Microbiology and Biotechnology, 57, 257–261. 22. APHA, 2005. Standard methods for the examination of water and wastewater. In: Eaton, A.D., Clesceri, L.S., Rice, E.W., Greenberg, A.E., & Franson, M.A.H. (Eds.), 21st ed. American Public Health Association, American Water Works Association, Water Environment Federation, USA (centennial edition) 23. Shuler, M. L., & Kargi, F. (2005). Bioprocess engineering—basic concepts (2nd ed., pp. 175–180). New Delhi: Prentice-Hall. 24. Megharaj, M., Avudainayagam, S., & Naidu, R. (2003). Current Microbiology, 47, 51–54. 25. Slobodkina, G. B., Bonch-Osmolovskaya, E. A., & Slobodkin, A. I. (2007). Microbiology, 76(5), 530–534. 26. Urvasi, T., & Datta, M. (2005). World Journal of Microbiology Biotechnology, 21, 891–899. 27. Wang, P. C., Mori, T., Toda, K., & Ohtake, H. (1990). Journal of Bacteriology, 172, 1670–1672. 28. Wang, Y. T., & Shen, H. (1995). Journal of Industrial Microbiology, 14, 159–163. 29. Camargo, F. A. O., Bento, F. M., Okeke, B. C., & Frankenberger, W. T. (2003). Journal of Environmental Quality, 32, 1228–1233. 30. Bae, W. C., Lee, H. K., Choe, Y. C., Jahng, D. J., Lee, S. H., Kim, S. J., et al. (2005). Journal of Microbiology, 43, 21–27. 31. Liu, Y., Xu, W., Zeng, G., Li, X., & Gao, H. (2006). Process Biochemistry, 41, 1981–1986. 32. Lakshman, R. S., & More, S. (2002). Minerals Engineering, 15, 831–837. 33. Deleo, P. C., & Ehrlich, H. L. (1994). Applied Microbiology Biotechnology, 40, 756–759. 34. Philip, L., Iyengar, L., & Venkobachar, C. (1998). Journal of Environmental Engineering ASCE, 124, 1165–1170. 35. Mclean, J. S., Beveridge, T. J., & Phipps, D. (2000). Environmental Microbiology, 2, 611–619. 36. Fulladose, E., Dejardin, V., Murat, J. C., Gourdon, R., & Villaescusa, I. (2006). Chemosphere, 65, 644–650.

Journal of Hazardous Materials 167 (2009) 685–691

Contents lists available at ScienceDirect

Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat

Biosorption of hexavalent chromium by Termitomyces clypeatus biomass: Kinetics and transmission electron microscopic study Sujoy K. Das, Arun K. Guha ∗ Department of Biological Chemistry, Indian Association for the Cultivation of Science, Jadavpur, Kolkata 700 032, India

a r t i c l e

i n f o

Article history: Received 1 August 2008 Received in revised form 8 January 2009 Accepted 8 January 2009 Available online 19 January 2009 Keywords: Termitomyces clypeatus Biosorption Cr(VI) Intracellular accumulation Transmission electron microscopy (TEM)

a b s t r a c t Biosorption of Cr+6 by Termitomyces clypeatus has been investigated involving kinetics, transmission electron microscopy (TEM) and Fourier transform infrared spectroscopic (FTIR) studies. Kinetics experiments reveal that the uptake of chromium by live cell involves initial rapid surface binding followed by relatively slow intracellular accumulation. Of the different chromate analogues tested, only sulfate ion reduces the uptake of chromium to the extent of ∼30% indicating chromate ions accumulation into the cytoplasm using sulfate transport system. Metabolic inhibitors, e.g. N,N -dicyclohexylcarbodiimide, 2,4-ditrophenol and sodium azide inhibit chromate accumulation by ∼30% in live cell. This indicates that accumulation of chromium into the cytoplasm occurs through the active transport system. TEM-EDXA analysis reveals that the chromium localizes in the cell wall and also in the cytoplasm. Reduction of chromate ions takes place by chromate reductase activity of cell-free extracts of T. clypeatus. FTIR study indicates that chromate ions accumulate into the cytoplasm and then reduced to less toxic Cr+3 compounds. © 2009 Elsevier B.V. All rights reserved.

1. Introduction Chromium, a toxic heavy metal, dissipates into the environment as a result of various industrial activities [1,2]. In view of toxicity and related environmental hazards [3], it is essential that the concentration of chromium in the effluent must be brought down to permissible limit [4] before discharging into water bodies. Among different available technologies [5,6] the removal of metal ions from wastewater by adsorption on biological materials specially microbial biomass known as biosorption/bioaccumulation [7–10] has recently gained much importance. This method does not generate toxic sludge, capable of reducing the concentration of metal ions below the permissible limit and the possibility of regeneration of the materials and thus provide an effective and economic means for the remediation of heavy metal polluted wastewater [11–14]. The uptake of heavy metals by microbial biomass is essentially a biphasic process consisting of metabolism independent initial cell surface binding that can occur either in living or inactivated organisms, followed by energy dependent intracellular accumulation which takes place only in the living cells [15]. The cell wall materials are involved in the initial surface binding of metal ions though electrostatic, physical and/or chemical interaction [16,17]. In living cells besides surface adsorption, metal ions may enter

into the cytoplasm through specific carrier system. The transport process in prokaryotic organisms has been studied in some details [18–22]. The state of art in the field of biosorption of heavy metals has recently been reviewed by Volesky [23]. However, only a few reports are available on fungal systems [24,25]. Fungal biomass has certain advantage over bacterial biomass in this natural ‘ecofriendly green technological process’ in respect of processing and handling of the biomass. Further, in comparison to bacteria, fungi are known to secret much higher amount of exopolymers, thereby significantly increasing the productivity of biosorption/bioremediation process [26]. In this manuscript we describe the biosorption/or bioaccumulation mechanism of chromium on Termitomyces clypeatus biomass (TCB) from kinetics study in presence of different co-ions and metabolic inhibitors with support from Fourier transform infrared spectroscopy and transmission electron microscopic investigations. 2. Materials and methods 2.1. Chemicals Dehydrated microbiological media and ingredients were procured from Himedia, India. All other reagents were of analytical grade and obtained from Merck, Germany and Sigma, USA. 2.2. Metal solution and analysis

∗ Corresponding author. Tel.: +91 33 2473 4971X502; fax: +91 33 2473 2805. E-mail addresses: [email protected], [email protected], [email protected] (A.K. Guha). 0304-3894/$ – see front matter © 2009 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2009.01.037

A stock solution of chromium (100 mg/l) was prepared by dissolving potassium dichromate (K2 Cr2 O7 ) in double distilled water

S.K. Das, A.K. Guha / Journal of Hazardous Materials 167 (2009) 685–691 [13] C. Lacina, G. Germain, A.N. Spiros, Utilization of fungi for biotreatment of raw wastewaters, Afr. J. Biotechnol. 2 (2003) 620–630. [14] S.P. Mishra, G. Roy Chowdhury, Biosorption of copper by Penicillium Sp., Miner. Process. Extr. Metall. Rev. 14 (1995) 111–126. [15] J.M.C. Tobin, C. White, G.M. Gadd, Metal accumulation by fungi: applications in environmental biotechnology, J. Ind. Microbiol. Biotechnol. 13 (1994) 126–130. [16] P.G.G. Burnett, C.J. Daughney, D. Peak, Cd adsorption onto Anoxybacillus flavithermus: surface complexation modeling and spectroscopic investigations, Geochim. Cosmochim. Acta 70 (2006) 5253–5269. [17] J.B. Fein, D.A. Fowle, J. Cahill, K. Kemner, M. Boyanov, B. Bunker, Nonmetabolic reduction of Cr(VI) by bacterial surfaces under nutrient-absent conditions, Geomicrobiol. J. 19 (2002) 369–382. [18] A.D. Costa, A. Carlos, F. Pereira, Bioaccumulation of copper, zinc, cadmium and lead by Bacillus sp., Bacillus cereus, Bacillus sphaericus and Bacillus subtilis, Braz. J. Microbiol. 32 (2001) 1–5. [19] H.A. Ghozlan, S.A. Sabry, R.A. Amer, Bioaccumulation of nickel, cobalt, and cadmium by free and immobilized cells of Pseudomonas spp., Fresenius Environ. Bull. 8 (1999) 428–435. [20] Y. Gomez, O. Coto, C. Hernandez, J. Marrero, L. Abin, Biosorption of nickel, cobalt and zinc by Serratia marcescens strain 7 and Enterobacter agglomerans strain 16, Process Metall. 11 (2001) 247–255. [21] H. Horitsu, S. Futo, K. Ozawa, K. Kawai, Comparison of characteristics of hexavalent chromium tolerant bacterium,Pseudomonas ambigua G-1, and its hexavalent chromium-sensitive mutant, Agric. Biol. Chem. 47 (1983) 2907–2908. [22] A.M. Marques, M.J.E. Tomas, F. Congregado, M.D. Simon-Pujol, Accumulation of chromium by Pseudomonas aeruginosa, Microbiol. Lett. 21 (1982) 143–147. [23] B. Volesky, Biosorption and me, Water Res. 41 (2007) 4017–4029. [24] J.N.L. Latha, K. Rashmi, M.P. Maruthi, Cell-wall-bound metal ions are not taken up in Neurospora crassa, Can. J. Microbiol. 51 (2005) 1021–1026. [25] J. Wang, C. Chen, Biosorption of heavy metals by Saccharomyces cerevisiae: a review, Biotechnol. Adv. 24 (2006) 427–451. [26] P. Mukherjee, A. Ahmed, D. Mandal, S. Senapati, S.R. Sainkar, M.I. Khan, R. Ramani, R. Parischa, P.V. Ajayakumar, M. Alam, M. Sastry, R. Kumar, Bioreduction of AuCl4 − Ions by the Fungus, Verticillium sp. and surface trapping of the gold nanoparticles formed, Angew. Chem. Int. Ed. 40 (2001) 3585–3588. [27] S.K. Das, A.K. Guha, Biosorption of chromium by Termitomyces clypeatus, Colloid Surf. B 60 (2007) 46–54. [28] R.K. Tandon, P.T. Crisp, J. Ellis, R.S. Baker, Effect of pH on chromium (VI) species in solution, Talanta 31 (1984) 227–228. [29] S.K. Das, A.R. Das, A.K. Guha, A study on the adsorption mechanism of mercury on Aspergillus versicolor biomass, Environ. Sci. Technol. 4 (2007) 8281–8287. [30] Y. Ishibashi, C. Cervantes, S. Silver, Chromium reduction in Pseudomonas putida, Appl. Environ. Microbiol. 56 (1990) 2268–2270. [31] M.F. Bryson, H.L. Drake, Energy-dependent transport of nickel by Clostridium pasteurianum, J. Bacteriol. 170 (1988) 234–238. [32] D. Blaudez, B. Botton, M. Chalot, Cadmium uptake and subcellular compartmentation in the ectomycorrhizal fungus Paxillus involutus, Microbiology 146 (2000) 1109–1117. [33] R.D. Tripathi, U.N. Rai, M. Gupta, M. Yunus, P. Chandra, Cadmium transport in submerged macrophyte Ceratophyllum demersum L. in presence of various metabolic inhibitors and calcium channel blockers, Chemosphere 31 (1995) 3783–3791. [34] H. Komeda, M. Kobayashi, S. Shimizu, A novel transporter involved in cobalt uptake, Proc. Natl. Acad. Sci. U.S.A. 94 (1997) 36–41.

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[35] S. Silver, Transport of cations and anions, in: B.P. Rosen (Ed.), Bacterial Transport, Marcel Dekker Inc., New York, 1978, p. 221. [36] R. Fuhrmann, A. Rothstein, The transport of Zn+2 , Co+2 and Ni+2 into yeast cells, Biochem. Biophys. Acta 163 (1968) 325–330. [37] P.R. Norris, D.P. Kelly, Accumulation of cadmium and cobalt by Saccharomyces cerevisiae, J. Gen. Microbiol 99 (1977) 317–324. [38] L.A. Okorokov, Main mechanisms of ion transport and regulation of ion concentrations in the yeast cytoplasm, in: I.S. Kulaev, E. Dawes, D.W. Tempest (Eds.), Environmental Regulation of Microbial Metabolism, Academic Press, London, 1985, p. 339. [39] E.L. Smith, R.L. Hill, I.R. Lehman, R. Lefkowiz, P. Handler, A. White, Principles of Biochemistry: General Aspects, 7th ed., McGraw-Hill Book Co., Singapore, 1985, p. 352. [40] H. Ohtake, C. Cervantes, S. Silver, Decreased chromate uptake in Pseudomonas fluorescens carrying a chromate resistance plasmid, J. Bacteriol. 169 (1987) 3853–3856. [41] N. Ohta, P.R. Galsworthy, A.B. Pardec, Genetics of Sulfate Transport by Salmonella typhimurium, J. Bacteriol. 105 (1971) 1053–1062. [42] Z. Tynecka, Z. Gos, J. Zajac, Reduced cadmium transport determined by a resistance plasmid in Staphylococcus aureus, J. Bacteriol. (147) (1981) 305–312. [43] S. Silver, T.K. Misra, Plasmid-mediated heavy metal resistances, Annu. Rev. Microbiol. 42 (1988) 717–743. [44] V. Farkas, Biosynthesis of cell walls of fungi, Microbiol. Mol. Biol. Rev. 43 (1979) 117–144. [45] F. Baldi, A.M. Vaughan, G.J. Olson, Chromium(VI)-resistant yeast isolated from a sewage treatment plant receiving tannery wastes, Appl. Environ. Microbiol 56 (1990) 913–918. [46] P.R. Puranik, K.M. Paknikar, Biosorption of lead, cadmium, and zinc by citrobacter Strain MCM B-181: characterization studies, Biotechnol. Prog. 15 (1999) 228–237. [47] J.-S. Chang, R. Law, C.-C. Chang, Biosorption of lead, copper and cadmium by biomass of Pseudomonas aeruginosa PU21, Water Res. 31 (1997) 1651–1658. [48] M.G.C. Baldry, A.C.R. Dean, Environmental change and copper uptake by Bacillus subtilis subsp. niger and by Pseudomonas fluorescens, Biotechnol. Lett. 3 (1981) 137–142. [49] R. Aravindam, B. Madhan, J.R. Rao, B.N. Nair, T. Ramasami, Bioaccumulation of chromium from tannery wastewater: an approach for chrome recovery and reuse, Environ. Sci. Technol. 38 (2004) 300–306. [50] J.N. Steel, J.L. Atwood, Supramolecular Chemistry, 1st ed., John Wiley & Sons, Chichester, 2000, p. 9. [51] L. Xia, R.L. McCreery, Chemistry of a chromate conversion coating on aluminum alloy AA2024-T3 probed by vibrational spectroscopy, J. Electrochem. Soc. 145 (1998) 3083–3089. [52] Y.G. Ko, U.S. Choi, T.K. Kim, D.J. Ahn, Y.J. Chun, FT-IR and isotherm study on anion adsorption onto novel chelating fibers, Macromol. Rapid Commun. 23 (2002) 535–539. [53] K. Nakamoto, Infrared Spectra of Inorganic and Co-ordination Compounds, 2nd ed., Wiley Interscience, New York, 1970. [54] A.N. Mabbett, J.R. Lloyd, L.E. Macaskie, Effect of complexing agents on reduction of Cr(VI) by Desulfovibrio vulgaris ATCC 29579, Biotechnol. Bioeng. 79 (2002) 389–397. [55] S. Viamajala, B.M. Peyton, J.N. Petersen, Modeling chromate reduction in Shewanella oneidensis MR-1: development of a novel dual-enzyme kinetic model, Biotechnol. Bioeng. 83 (2003) 790–797.

Environ. Sci. Technol. 2010, 44, 400–405

Remediation of Chromium(VI) by a Methane-Oxidizing Bacterium ABUBAKR AL HASIN,† STEPHEN J. GURMAN,‡ LORETTA M. MURPHY,§ ASHLEE PERRY,† T H O M A S J . S M I T H , * ,† A N D PHILIP H. E. GARDINER† Biomedical Research Centre, Sheffield Hallam University, Howard Street, Sheffield S1 1WB, United Kingdom, Department of Physics and Astronomy, University of Leicester, University Road, Leicester LE1 7RH, United Kingdom, and School of Chemistry, Bangor University, Bangor LL57 2UW, United Kingdom

Received June 11, 2009. Revised manuscript received November 13, 2009. Accepted November 20, 2009.

Methane-oxidizing bacteria are ubiquitous in the environment and are globally important in oxidizing the potent greenhouse gas methane. It is also well recognized that they have wide potential for bioremediation of organic and chlorinated organic pollutants, thanks to the wide substrate ranges of the methane monooxygenase enzymes that they produce. Here we have demonstrated that the well characterized model methanotroph Methylococcus capsulatus (Bath) is able to bioremediate chromium(VI) pollution over a wide range of concentrations (1.4-1000 mg L-1 of Cr6+), thus extending the bioremediation potential of this major group of microorganisms to include an important heavy-metal pollutant. The chromium(VI) reduction reaction was dependent on the availability of reducing equivalents from the growth substrate methane and was partially inhibited by the metabolic poison sodium azide. X-ray spectroscopy showed that the cell-associated chromium was predominantly in the +3 oxidation state and associated with cell- or medium-derived moieties that were most likely phosphate groups. The genome sequence of Mc. capsulatus (Bath) suggests at least five candidate genes for the chromium(VI) reductase activity in this organism.

Introduction Microbiologically catalyzed reduction reactions offer a possible solution to environmental pollution with chromate(VI), which is a highly oxidizing, soluble, mutagenic, and toxic form of the metal that is produced as an effluent from metal plating, tanning, paper making, and other industries (1, 2). Reduction of chromium(VI) to the +3 oxidation state produces a form of the metal that is less toxic, less bioavailable and more able to adsorb to negatively charged biopolymers and soil particles (1-3). Governments have sought to restrict its use and release; for instance, the European Union has restricted the use of chromium in electrical equipment (4) and exposure limits have been implemented to protect individuals from harm due to chromate (VI) contamination * Corresponding author phone +44 114 225 3042; fax +44 114 225 3066; e-mail [email protected]. † Sheffield Hallam University. ‡ University of Leicester. § Bangor University. 400

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(5). Nonetheless, large-scale industrial use of hexavalent chromium continues around the world and continues to be an environmental threat (6, 7). Chromate(VI) reductase activity, capable of reducing chromium from the +6 to the +3 oxidation state, has previously been characterized in aerobic bacteria and facultative anaerobes, such as Escherichia coli, Pseudomonas putida (8-11), Paracoccus denitrificans (12), and Bacillus subtilis (13), as well as in anaerobic sulfate-reducing bacteria (14, 15). Here, we have investigated the chromium(VI) reduction in methane-oxidizing bacteria, a ubiquitous group of environmental bacteria (16), in which to our knowledge this reaction has not previously been investigated. Methane-oxidizing bacteria (or methanotrophs) are defined by their ability to use methane as their sole carbon and energy source. The best characterized examples belong the R and γ subdivisions of the proteobacteria, although recent results have indicated that aerobic methanotrophs are extremely diverse in terms both of their phylogeny and the range of environments in which they live (16-21). The type I methanotrophs, belonging to the γ-proteobacteria, include the well-studied Methylococcus capsulatus (Bath); methanotrophs of the R-proteobacteria (type II methanotrophs) include Methylosinus trichosporium OB3b (16). Methanotrophs are well recognized as environmentally significant organisms. Mole-for-mole environmental methane (produced by anaerobic breakdown of organic matter) is 21 times more powerful as a greenhouse gas than carbon dioxide (22), and so oxidation of methane by methanotrophs is important in controlling global warming (23). In addition to their ability to oxidize their growth substrate methane, the methane monooxygenase enzymes produced by methanotrophic bacteria can also co-oxidize diverse hydrocarbons and halogenated organic compounds, including aromatics and the priority pollutant trichloroethylene, and so application of methanotrophs in bioremediation of such compounds has been widely investigated (24-26). Here we have investigated the interaction of chromate (VI) with well characterized representatives of the two major groups of methaneoxidizing bacteria and have identified a strain that is able to bioremediate this major heavy-metal pollutant.

Materials and Methods Bacterial Strains and Growth Conditions. The methanotrophs Ms. trichosporium OB3b and Mc. capsulatus (Bath) were obtained from the culture collection of H. Dalton and J. C. Murrell (University of Warwick, U.K.) and were grown and propagated aerobically in nitrate mineral salts (NMS) medium or on NMS agar (27) containing 1 mg L-1 of CuSO4 · 5H2O using methane (1:4 v/v in air) as the source of carbon and energy. All cultures and chromate-reduction experiments were incubated at the optimal growth temperature of the organism concerned, 30 and 45 °C, respectively, for Ms. trichosporium OB3b and Mc. capsulatus (Bath). Except where stated otherwise, chromate reduction experiments were performed in 50 mL liquid cultures in 250-mL conical Quickfit flasks sealed with a Subaseal (Fisher) to prevent loss of methane, while allowing addition of liquids and the taking of samples using hypodermic syringes. Cultures were allowed to grow to an OD600 of 0.3-0.8 before addition of potassium chromate (VI) or potassium dichromate (VI) to give the concentration of hexavalent chromium stated for each experiment. Fermentor cultivation of Mc. capsulatus (Bath) was performed using methane as the source of carbon and energy according to the published method (28), in a Bioflo 110 fermentor (New Brunswick Scientific, vessel capacity 4.5 10.1021/es901723c

 2010 American Chemical Society

Published on Web 12/03/2009

one significant homologue in Mc. capsulatus (accession no. YP_113831, E ) 6 × 10-30). The Old Yellow Enzyme-type chromate reductase of Thermus scotoductus (34) also has a highly significant homologue in Mc. capsulatus (accession no. YP_113154, E ) 2 × 10-28). The chromate reductase ChrR of P. putida ((9); accession no. Q93T20), a flavoprotein capable of chromate reduction, did not have any significant homologues in Mc. capsulatus. The known chromate efflux system ChrA, typified by the chromate efflux pump of Pseudomonas aeruginosa plasmid pUM505 ((35) accession no. P14285), which might have contributed to resistance of the cells to chromate(VI), also did not have any significant matches in the genome of Mc. capsulatus. The annotation of the genome sequence (36), however, indicates 14 putative proteins with possible roles in heavy-metal efflux systems in this organism and another 28 proposed to be involved in efflux generally (data not shown).

Discussion

FIGURE 4. EXAFS spectroscopy of the particulate fraction from a Mc. capsulatus culture treated with 1000 mg L-1 of hexavalent chromium for 96 h: (a) isolated EXAFS oscillations and (b) Fourier transform EXAFS. Experimental data are shown as open diamonds joined with dotted lines and fits as solid lines. precipitated from NMS medium when chromium(III) was added to a concentration of 1000 mg L-1, so we suggest that this sample contains a suspension of chromium phosphate. There was no sign of a diminution of the chromium fluorescence signal during the six hours of data taking, suggesting that no settling is occurring: the suspension must therefore be very finely divided. The environment of chromium in the cells is identical to that in the solution containing phosphate. Again the inclusion of a second shell significantly lowers the fit index. The results imply that chromium in the cells is in the Cr(III) state (as shown by the Cr-O distance) and probably in a suspension (as shown by the second shell contribution). This suspension is most probably insoluble Cr(PO4). To within our rather large uncertainties there is no change in the chromium environment with time of exposure of cells to chromium(VI) between 24 and 96 h nor with the temperature of EXAFS data acquisition (Supporting Information Table 1). This suggests that chromium in these cells is reduced to Cr(III) and precipitated out within 24 h. Genome of Mc. capsulatus Contains Candidate Chromate Reductase Genes. To identify genes that could encode proteins involved in reduction of chromate by Mc. capsulatus, the database of translated open reading frames from the complete genome sequence (i.e., all the known and potential proteins of the organism) was searched for homologues to the known classes of chromate reductases from other bacteria. BLAST searches indicated the presence of three homologues of the E. coli Fre chromate reductase (11; accession no. M74448) in the Mc. capsulatus (Bath) genome all of which are known or likely flavin/Fe2S2 oxidoreductases: (1) a protein annotated as a putative oxygenase (accession no. YP_114919, E ) 5 × 10-12), (2) the reductase component of sMMO (MmoC, E ) 4 × 10-10), and (3) a protein annotated as a putative Na+-translocating NADH-quinone reductase subunit (accession no. YP_114800, E ) 3 × 10-9). The E. coli nitroreductase NfsA, which also reduces chromate (8), has

Here we have shown that a methanotrophic bacterium, Mc. capsulatus (Bath), is able to detoxify chromate(VI) over a wide range of concentrations and that the product was chromium in the relatively nontoxic +3 oxidation state. The observation of chromium(III) in a phosphorus/oxygen coordination environment in the particulate fraction after exposure of cells to 1000 mg L-1 of hexavalent chromium is consistent with the formation of insoluble chromium(III) phosphate in the phosphate-containing growth medium. Indeed, the electron-dense particles seen in via transmission electron microscopy of cells from cultures exposed to 500 mg L-1 of chromium(VI) may be particles of such chromium(III) phosphate. Phosphorus/oxygen coordination environments would also be produced from association of (a proportion of) the chromium(III) with phosphate-containing cellular components such as nucleotide coenzymes (11) and DNA, where phosphorus/oxygen coordination of chromium(III) may be important in the mutagenic properties of chromate (VI) (37). The apparently uniform staining of the cells with electron dense material after exposure to the lower concentration of chromate (VI) of 100 mg L-1 suggests association of the chromium with some form of cellular material. The effect of the metabolic inhibitor sodium azide on the chromate (VI) reduction reaction, and also the fact that the reaction is not shown by autoclaved (dead) cells, supports the conclusion that the reaction is dependent on active cellular metabolism rather than merely a reaction between the chromate and cellular constituents as described by Fein et al. (38). The effect of azide is presumably an indirect one since inhibition of the reduction of dioxygen to water would not per se prevent channelling of reducing equivalents into reduction of chromate. The absolute dependence of chromate reductase activity at 1.4 mg L-1 of chromium(VI) on the presence of methane very strongly suggests that the growth substrate methane is required to supply electrons for reduction of chromate. While there is currently no clear evidence to show which cellular components are responsible for reduction of chromium(VI), sequence similarity searches using known chromate-reducing enzyme sequences indicate at least five possible candidates for the chromium-reducing enzyme in Mc. capsulatus. One of these is the well characterized reductase component of soluble methane monooxygenase, a component of one of the enzyme systems that is involved in oxidation of methane to methanol (26, 39). It is unlikely that this particular candidate is involved in reduction of chromate in these experiments because the cells were cultivated under conditions of high copper-to-biomass ratio, when the soluble methane monooxygenase is not expressed (40). Future availability of the genome sequence for Ms. VOL. 44, NO. 1, 2010 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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(29) McLean, J.; Beveridge, T. J. Chromate reduction by a pseudomonad isolated from a site contaminated with chromated copper arsenate. Appl. Environ. Microbiol. 2001, 67, 1076–1084. (30) Gurman, S. J.; Binsted, N.; Ross, I. A rapid, exact curved-wave theory for EXAFS calculations. J. Phys. (Paris) 1984, C17, 143–151. (31) Joyner, R. W.; Martin, K. J.; Meehan, P. Some applications of statistical tests in analysis of EXAFS and SEXAFS data. J. Phys. (Paris) 1987, C20, 4005–4012. (32) Bajt, S.; Clark, S. B.; Sutton, S. R.; Rivers, M. L.; Smith, J. V. Synchrotron X-ray microprobe determination of chromate content using X-ray-absorption near-edge structure. Anal. Chem. 1993, 65, 1800–1804. (33) M. V. Aldrich, Gardea-Torresdey, J. L.; Peralta-Videa, J. R.; Parsons, J. G. Uptake and reduction of Cr(VI) to Cr(III) by mesquite (Prosopis spp.): chromate-plant interaction in hydroponics and solid media studied using XAS. Environ. Sci. Technol. 2003, 37, 1859–1864. (34) Opperman, D. J.; Piater, L. A.; van Heerden, E. A novel chromate reductase from Thermus scotoductus SA-01 related to Old Yellow Enzyme. J. Bacteriol. 2008, 190, 3076–3082. (35) Cervantes, C.; Ohtake, H.; Chu, L.; Misra, T. K.; Silver, S. Cloning, nucleotide sequence, and expression of the chromate resistance determinant of Pseudomonas aeruginosa plasmid pUM505. J. Bacteriol. 1990, 172, 287–291. (36) Ward, N.; Larsen, O.; Sakwa, J.; Bruseth, L.; Khouri, H.; Durkin, A. S.; Dimitrov, D.; Jiang, L.; Scanlan, D.; Kang, K. H.; Lewis, M.; Nelson, K. E.; Methe´, B.; Wu, M.; Heidelberg, J. F.; Paulsen, I. T.; Fouts, D.; Ravel, J.; Tettelin, H.; Ren, Q.; Read, T.; DeBoy, R. T.; Seshadri, R.; Salzberg, S. L.; Jensen, H. B.; Birkeland, N. K.; Nelson, W. C.; Dodson, R.; Grindhaug, S. H.; Holt, I.; Eidhammer, I.; Jonasen, I.; Vanaken, S.; Utterback, T.; Feldblyum, T. V.; Fraser, C. M.; Lillehaug, J. R.; Eisen, J. A. Genomic insights into methanotrophy: The complete genome sequence of Methylococcus capsulatus (Bath). PLoS Biol. 2004, 2, 1616–1628.

(37) Zhitkovich, A.; Song, Y.; Quievryn, G.; Voitkun, V. Non-oxidative mechanisms are responsible for the induction of mutagenesis by reduction of Cr(VI) with cysteine: role of ternary DNA adducts in Cr(III)-dependent mutagenesis. Biochemistry 2001, 40, 549– 560. (38) Fein, J. B.; Fowle, D. A.; Cahill, J.; Kemner, K.; Boyanov, M.; Bunker, B. Nonmetabolic reduction of Cr(VI) by bacterial surfaces under nutrient-absent conditions. Geomicrobiol. J. 2002, 19, 369–382. (39) Lund, J.; Woodland, M. P.; Dalton, H. Electron transfer reactions in the soluble methane monooxygenase of Methylococcus capsulatus (Bath). Eur. J. Biochem. 1985, 147, 297–305. (40) Stanley, S. H.; Prior, S. D.; Leak, D. J.; Dalton, H. Copper stress underlies the fundamental change in intracellular location of methane monooxygenase in methane-oxidizing organismss Studies in batch and continuous cultures. Biotechnol. Lett. 1983, 5, 487–492. (41) Kim, H. J.; Graham, D. W.; DiSpirito, A. A.; Alterman, M. A.; Galeva, N.; Larive, C. K.; Asunskis, D.; Sherwood, P. M. A. Methanobactin, A copper-acquisition compound from methaneoxidizing bacteria. Science 2004, 305, 1612–1615. (42) Choi, D. W.; Do, Y. S.; Zea, C. J.; McEllistrem, M. T.; Lee, S.-W.; Semrau, J. D.; Pohl, N. L.; Kisting, C. J.; Scardino, L. L.; Hartsel, S. C.; Boyd, E. S.; Geesey, G. G.; Riedel, T. P.; Shafe, P. H.; Kranski, K. A.; Tritsch, J. R.; Antholine, W. E.; DiSpirito, A. A. Spectral and thermodynamic properties of Ag(I), Au(III), Cd(II), Co(II), Fe(III), Hg(II), Mn(II), Ni(II), Pb(II), U(IV), and Zn(II) binding by methanobactin from Methylosinus trichosporium OB3b. J. Inorg. Biochem. 2006, 100, 2150–2161. (43) Jenkins, M. B.; Chen, J.-H.; Kadner, D. J.; Lion, L. W. Methanotrophic bacteria and facilitated transport of pollutants in aquifer material. Appl. Environ. Microbiol. 1994, 60, 3491–3498.

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Journal of Basic Microbiology 2008, 48, 135 – 139

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Short Communication Removal of chromium (VI) through biosorption by the Pseudomonas spp. isolated from tannery effluent Jatin Srivastava1, Harish Chandra2, Kirti Tripathi3, Ram Naraian3 and Ranjeev K Sahu1 1

2 3

Department of Environmental Sciences, Chatrapati Shahu Ji Maharaj University, Kalyanpur – Kanpur – 208024 UP, India Department of Microbiology, Gayatri College of Biomedical Sciences, Dehradun (UK), India Department of Microbiology, Chatrapati Shahu Ji Maharaj University, Kalyanpur – Kanpur – 208024 UP, India

Heavy metal contamination of the rivers is a world wide environmental problem and its removal is a great challenge. Kanpur and Unnao two closely located districts of Uttar Pradesh India are known for their leather industries. The tanneries release their treated effluent in the near by water ways containing Cr metal that eventually merges with the river Ganges. Untreated tannery effluent contains 2.673 ± 0.32 to 3.268 ± 0.73 mg l–1 Cr. Microbes were isolated, keeping the natural selection in the view, from the tannery effluent since microbes present in the effluent exposed to the various types of stresses and metal stress is one of them. Investigations include the exposure of higher concentrations of Cr(VI) 1.0 to 4.0 mg l–1 to the bacteria (presumably the Pseudomonas spp.) predominant on the agar plate. The short termed study (72 h) of biosorption showed significant reduction of metal in the media especially in the higher concentrations with a value from 1.0 ± 0.02, 2.0 ± 0.01, 3.0 ± 0, and 4.0 ± 0.09 at zero h to 0.873 ± 0.55, 1.840 ± 1.31, 2.780 ± 0.03 and 3.502 ± 0.68 at 72 h respectively. The biosorption of metal show in the present study that the naturally occurring microbes have enough potential to mitigate the excessive contamination of their surroundings and can be used to reduce the metal concentrations in aqueous solutions in a specific time frame. Keywords: Natural selection / Tannery effluent / Pseudomonas spp. / Cr (VI) / Biosorption Received: October 16, 2007: accepted December 07, 2007 DOI 10.1002/jobm.200700291

Introduction* Heavy metal contamination of the rivers is a world wide environmental problem and its removal is a great challenge. The di-chromate compounds are used as oxidizing agents in quantitative analysis of various water quality parameters such as chemical oxygen demand (COD) and in tanning processes. Chromium compounds are used in the textile and aircraft industry as mordents and anodizing agents respectively. The fate of chromium in the environment is strongly dependent on its valence state [1]. However, chromium levels in air, water, and food are generally very low; the major Correspondence: Dr. Jatin K. Srivastava, Department of Environmental Sciences, Chatrapati Shahu Ji Maharaj University, Kanpur – 208024 UP, India E-mail: [email protected] © 2008 WILEY-VCH Verlag GmbH & Co. KGaA, Weinheim

human exposure is occupational [2]. Cr(VI) is reactive and a potent carcinogenic species [3]. Wastewater containing chromium must be treated before being discharged into the environment. The most commonly used method to remove Chromium from liquid effluents is alkaline precipitation, but the method is expensive, therefore cheaper and effective bioremediation techniques using bacteria [4, 5], soils [6], algae [7] and plants [8] are being studied all over the world. Microorganisms can physically remove heavy metals from solution through either bioaccumulation or biosorption. In bioaccumulation, metals are transported from the outside of the microbial cell, through the cellular membrane, and into the cell cytoplasm, where the metal is sequestered. Earlier reports of Wong and So [9] suggested the accumulation of Cu(II) ions by the isolated Pseudomonas pudida II-11, from electroplating effluent. www.jbm-journal.com

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crease in 24 h followed by 48 h however; almost no change was observed in the mass collected after 72 h of exposure.

Journal of Basic Microbiology 2008, 48, 135 – 139

study least cysteine residues were found in the microbes (Fig. 2).

Conclusion Discussion Metals play an integral role in the life processes of micro-organisms [18]. Some metals such as Ca, CO, Cr, Cu, K, Mg, Zn and Na are required nutrients and are essential for the growth of microbes however; the higher concentrations of these metals are toxic to every living cell. Micro-organisms are highly effective in sequestering heavy metals. In the present study bacterial mass was isolated from tannery effluent which is a source pollutant of Cr(VI) in the local area of district Unnao. The presence of microbial population in the effluent can be considered as the tolerant strains. Out of 9 different types of microbes the bacterium (in the present study) was chosen on the basis of growth performance in laboratory condition and convenient in handling. The bacterial colony was exposed to different concentration of Cr metal which showed significant reduction along with accumulation in the cells. The study showed that bacterial biomass in broth could adsorb the metal on the outer surface initially and died as soon as it gets inside the cell of bacterium. The biosorption of metals does not consume cellular energy. Positively charged metal ions are sequestered primarily through the adsorption of metals to the negative ionic groups on cell surfaces, the polysaccharide coating found on most forms of bacteria, or other extra-cellular structures such as capsules or slime layers. Binding sites on microbial cell surfaces usually are carboxyl residues, phosphate residues, SH groups, or hydroxyl groups. Non-essential metals bind with greater affinity to SH group [19]. Bacterial cells which are capable of forming an extra-cellular polysaccharide coating e.g., Pseudomonas sp. bio-adsorbs (biosorp) metal ions and can prevents them from interacting with vital cellular components [20]. The amount of metal biosorbed to the exterior of bacterial cells often exceeds the amount predicted using information about the charge density of the cell surface. In the present study, longer survival of the bacterial cells were found in lower concentration of Cr ions and less in higher concentration. It indicated that Cr metal might have entered into the bacterial cell either by ligand interaction or by active membrane uptake. Rouch et al. [21] demonstrated the presence of the cysteine rich proteins in bacterial cells such as Pseudomonas sp. and Synechococcus spp. which provide resistance to the bacterial cells. However; in the present © 2008 WILEY-VCH Verlag GmbH & Co. KGaA, Weinheim

Significant reduction in the metal concentration was observed in all the sets with a significant accumulation in the cells. However; the bacterial cells could not survive for longer period in the broth exhibiting the toxic response of the metal on the bacterial mass. The findings support the well established fact that living organisms naturally selected and can survive the harsh conditions is naturally selected and can be used for the mitigation of pollution from the water and soil.

References [1] Jeremy, F., Joshua, C., David, F., Ken, K., Bruce, B. and Maxim, B., 2000. Non-metabolic reduction of Cr(VI) by bacterial surfaces under nutrient absent conditions. Journal of Conference & Abstracts, 5, 396. [2] Chou, I. N., 1989. Distinct cytoskeleton injuries induced by As, Cd, Co, Cr, and Ni compounds. Biomedical Environmental Sciences, 2, 358 – 365. [3] Gibbs, H.J., Lees, P.S., Pinsky, P.F. and Rooney, B.C., 2000. Lung Cancer among workers in chromium chemical production. Amer. J. Indust. Medicines, 38, 115 – 126. [4] Ohtake, H, Fujii, E. and Toda, K., 1990. Reduction of toxic chromate reducing strain of Enterobacter cloacae. Environm. Technol., 11, 663 – 668. [5] Coleman, R.N. and Paran, J.H., 1991. Biofilm concentration of chromium. Environm. Technol., 12, 1079 – 1093. [6] Losi, M.E., Amrhein, C. and Frankenberger, W.T., Jr., 1994. Bioremediation of chromate contaminated groundwater by reduction and precipitation in surface soils. J. Environm. Quality, 23, 1141 – 1150. [7] Brady, D., Letebele, B., Duncan, J.R. and, Rose, P.D., 1994. Bioaccumulation of metals by Scenedesmus, Selenastrum and Chlorella algae. Water SA, 20, 213 – 218. [8] Wolverton, B.C. and Mc Donald, R.C., 1979. The water hyacinth: from prolific pest to potential provider. Ambio, 8, 2 – 9. [9] Wong, P.K. and So, C.M., 1993. Copper accumulation by a strain of Pseudomonas putida. Microbios, 73, 113 – 121. [10] Silver, S., 1991. Bacterial heavy metal resistance systems and possibility of bioremediation. In: Biotechnology, Bridging Research and Applications, pp. 265 – 287. Kluwer Academic Publishers, London. [11] Cole, F.A. and Clausen, C.A., 1997. Bacterial biodegradation of CCA treated waste wood. In: Proceedings, Forest Products Society Conference on Use of Recycled Wood and Paper in Building Applications, pp. 201 – 204 (September 9 1996, Madison, WI, USA). Forest Products Society. www.jbm-journal.com

Biochemical Engineering Journal 36 (2007) 54–58

Biosorption of Cr(III) and Cr(VI) onto the cell surface of Pseudomonas aeruginosa So-Young Kang a , Jong-Un Lee b , Kyoung-Woong Kim a,∗ a

Department of Environmental Science and Engineering, Gwangju Institute of Science and Technology (GIST), Gwangju 500-712, South Korea b Department of Civil, Geosystem and Environmental Engineering, Chonnam National University, Gwangju 500-757, South Korea Received 6 September 2005; received in revised form 16 May 2006; accepted 12 June 2006

Abstract Biosorption of the chromium ions Cr(III) and Cr(VI) onto the cell surface of Pseudomonas aeruginosa was investigated. Batch experiments were conducted with various initial concentrations of chromium ions to obtain the sorption capacity and isotherms. It was found that the sorption isotherms of P. aeruginosa for Cr(III) were described well by Langmuir isotherm models, while Cr(VI) appeared to fit Freundlich models. The results of FT-IR analysis suggested that the chromium binding sites on the bacterial cell surface were most likely carboxyl and amine groups. The bacterial surface of P. aeruginosa seemed to engage in reductive and adsorptive reactions with respect to Cr(VI) biosorption. © 2006 Elsevier B.V. All rights reserved. Keywords: Biosorption; Pseudomonas aeruginosa; Chromium; FT-IR spectroscopy; Bioremediation; Wastewater treatment

1. Introduction Toxic heavy metals are frequently contained in wastewaters produced by many industrial processes, such as those employed in the electroplating, metal finishing, metallurgical, tannery, chemical manufacturing, mining, and battery manufacturing industries [1,2]. The existence of heavy metals in the environment represents a very significant and long-term environmental hazard. Even at low concentrations these metals can be toxic to organisms, including humans. In particular, chromium is a contaminant that is a known mutagen, teratogen and carcinogen [3]. Chromium is generally found in electroplating and metal finishing industrial effluents, as well as sewage and wastewater treatment plant discharges [4]. Among the several oxidation states (di, tri, penta and hexa), trivalent chromium, Cr(III), together with the hexavalent state, Cr(VI), can be the main forms present in aquatic environments [5]. Chromate (CrO4 2− ) is the prevalent species of Cr(VI) in natural aqueous environments, and is the major pollutant from chromium-related industries [6]. Although Cr(III) is less toxic than Cr(VI), long-term exposure to Cr(III)

∗

Corresponding author. Tel.: +82 62 970 2442; fax: +82 62 970 2434. E-mail address: [email protected] (K.-W. Kim).

1369-703X/$ – see front matter © 2006 Elsevier B.V. All rights reserved. doi:10.1016/j.bej.2006.06.005

is known to cause allergic skin reactions and cancer [7]. As a result, the total chromium level in effluent is strictly regulated in many countries. In the USA, the concentration of chromium in drinking water has been regulated with a maximum level of 0.1 mg/l for total chromium [8]. The removal of heavy metals from aqueous solutions has therefore received considerable attention in recent years. However, the practical application of physicochemical technology such as chemical precipitation, membrane filtration and ion exchange is sometimes restricted due to technical or economical constraints. For example, the ion exchange process is very effective but requires expensive adsorbent materials [9,10]. The use of low-cost waste materials as adsorbents of dissolved metal ions provides economic solutions to this global problem and can be considered an eco-friendly complementary [11,12]. At present, emphasis is given to the utilization of biological adsorbents for the removal and recovery of heavy metal contaminants. Biomass involving pure microbial strains has shown high capacities for the selective uptake of metals from dilute metalbearing solutions. Several investigations have reported that Pseudomonas aeruginosa displays efficiency for metal uptake [13–15]. Chang and Hong [16] found that the amount of mercury adsorbed by a P. aeruginosa biomass sample (180 mg Hg/g dry cells) was higher than that bound to a cation exchange

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Fig. 4. Biosorption isotherms of Cr(VI) onto P. aeruginosa. The biomass was contacted with metal solution for 10 h at 25 ◦ C and 180 rpm in shaking incubator. The lines were produced by using the MINEQL+.

the following reaction: (4)

Fig. 5. FT-IR spectra of P. aeruginosa prepared in KBr disks: (a) pristine; (b) Cr(III)-loaded; (c) Cr(VI)-loaded bacteria.

A number of previous experimental studies of bacterial Cr(VI) reduction reported the enzymatic reduction as a product of metabolic activity [27,28], and most of these focused on the requirement of external electron donors for reduction to occur [29,30]. However, Fein et al. [31] suggested that the Cr(VI) reduction is not dependent on cell metabolism and that some component of the cell wall serves as the electron donor for the reduction reaction. We have also investigated the reduction of Cr(VI) to Cr(III) in the absence of externally supplied electron donors. In the second process, chromium ions are removed from wastewater using the adsorptive functional groups of P. aeruginosa. The adsorptive property is due to the electrostatic interaction between the charged surfaces of bacteria and chromium ions. The experimental sorption isotherms of Cr(VI) are represented by the Freundlich sorption isotherm in Fig. 4. The linearized form of Freundlich is represented by the following equation [26]:

carried out using chromium-loaded P. aeruginosa. The absorption spectrum of chromium-loaded biomass (at pH 5) was compared with that of pristine biomass. The chromium-loaded biomass was washed, dried and powdered after biosorption of chromium ions under the same conditions used in the preparation of pristine biomass. A change of absorption bands can be seen when comparing the FT-IR spectra of pristine and chromiumloaded biomass (Fig. 5). Fig. 5(b) shows the changes in the spectrum of the biomass after sorption of Cr(III) by P. aeruginosa. An interesting phenomenon was the sharp decrease in the band intensity at 1414 cm−1 corresponding to C O stretching after metal binding. On the basis of the change of the band, it was reasonable to assume that the peak value suggested the chelating (bidentate) character of the Cr(III) biosorption onto carboxyl groups [32]. The structure of the metal bound to carboxyl ligands on the bacteria is likely to take the following form [33]:

Cr2 O7 2− + 16e− + 14H+ ↔ 2Cr 3+ + 7H2 O

log Γ = log m + n log[A]

(5)

where m represents the Freundlich constant and n is the measure of the nonlinearity involved. Values of m and n were, respectively, found to be 80.8 and 1.03 as the total adsorbed chromium ions; 38.6 and 1.02 as the adsorbed Cr(III) in Cr(VI) biosorption to P. aeruginosa. The difference of concentration between total and hexavalent chromium was taken as the concentration of trivalent chromium. These results show that the bacterial functional groups of P. aeruginosa can act as reductive and adsorptive sites in metal biosorption. 3.3. FT-IR spectra of chromium-loaded P. aeruginosa To confirm the difference between functional groups in relation to biosorption of Cr(III) and Cr(VI), the FT-IR study was

In the case of Cr(VI)-loaded bacteria, the spectral analysis of P. aeruginosa before and after metal binding indicated that –NH is involved in Cr(VI) biosorption (Fig. 5(c)). There is a substantial decrease in the absorption intensity of –NH bands at 1660 and 1551 cm−1 . The broad overlapping range for N–H and O–H stretching in the range 3200–3600 cm−1 also presents some changes, but it is difficult to determine the group that causes the shift. These amino groups are protonated at pH 3 [34] and the negatively charged chromate ions become electrostatically attracted to the positively charged amines of the biomass cell wall. Similar to Cr(III)-loaded bacteria, the characteristic peak

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of C O stretching at 1414 cm−1 decreased and indicated Cr(III) binding after reduction of Cr(VI) to Cr(III). 4. Conclusions This study shows that P. aeruginosa can be applied to chromium-contaminated wastewater. The sorption of chromium ions by P. aeruginosa was modeled well by the Langmuir and Freundlich sorption isotherms. The data of potentiometric titration indicated the presence of two major functional groups on the cell wall, corresponding to pKa values of 5.2 and 9.5. FT-IR spectrometry showed bindings of chromium ions were dominated by complexation to the carboxyl and amine groups on the biomass surface. In the case of Cr(VI) biosorption of P. aeruginosa, the reduction and adsorption of Cr(VI) occurred coincidently in an abiotic process. This phenomenon is environmentally significant because most bacteria in the subsurface exist in nutrient-poor or -absent conditions under natural conditions [31]. This study shows the potential for the use of P. aeruginosa for chromium recovery in various water and wastewater treatment applications, and highlights the efficacy of using biological agents for the remediation of polluted aqueous environments. Acknowledgements This research was supported by the Gwangju Institute of Science and Technology (GIST) Research Fund and National Research Laboratory Project (Arsenic Geoenvironment Lab.) to K.-W. Kim. References [1] B.J. Alloway, Heavy Metals in Soils, Kluwer Academic/Plenum Publishers, New York, 1994. [2] C. Polprasert, L.R.J. Liyanage, Hazardous waste generation and processing, Resour. Conserv. Recycl. 16 (1996) 213–226. [3] L.W. Chang, Toxicology of Metals, CRC Press, Boca Raton, FL, 1996. [4] E. Merian, Metals and their Compounds in the Environment. Occurrence, Analysis and Biological Relevance, VCH, Weinheim, 1991. [5] Agency for Toxic Substances and Disease Registry (ATSDR), Toxicological Profile for Chromium, U.S. Public Health Service, U.S. Department of Health and Human Services, Atlanta, GA, 1998. [6] U.S. Environmental Protection Agency, Toxicological Review of Hexavalent Chromium, National Center for Environmental Assessment, Office of Research and Development, Washington, DC, 1998. [7] U.S. Environmental Protection Agency, Toxicological Review of Trivalent Chromium, National Center for Environmental Assessment, Office of Research and Development, Washington, DC, 1998. [8] U.S. Environmental Protection Agency, List of Drinking Water Contaminants and MCLs, EPA 816-F-03-016, Office of Water, Washington, DC, 2003. [9] M. Lehmann, A.I. Zouboulis, K.A. Matis, Removal of metal ions from dilute aqueous solutions: a comparative study of inorganic sorbent materials, Chemosphere 39 (1999) 881–892. [10] B. Volesky, Detoxification of metal-bearing effluents: biosorption for the next century, Hydrometallurgy 59 (2001) 203–216. [11] M.D. Mullen, D.C. Wolf, F.G. Ferris, T.J. Beveridge, C.A. Flemming, G.W. Bailey, Bacterial sorption of heavy metals, Appl. Environ. Microbiol. 55 (1989) 3143–3149.

[12] B. Volesky, Z.R. Holan, Biosorption of heavy metals, Biotechnol. Prog. 11 (1995) 235–250. [13] G.M. Strandberg, S.E. Shumate II, J.R. Parrott, Microbial cells as biosorbents for heavy metals: accumulation of uranium by Saccharomyces serevisiae and Pseudomonas aeruginosa, Appl. Environ. Microbiol. 41 (1981) 237–245. [14] A.-C. Texier, Y. Andr`es, P. Le Cloirec, Selecive biosorption of lanthanide (La, Eu, Yb) ions by Pseudomonas aeruginosa, Environ. Sci. Technol. 33 (1999) 489–495. [15] A.-C. Texier, Y. Andr`es, M. Illemassene, P. Le Cloirec, Characterization of lanthanide ions binding sites in the cell wall of Pseudomonas aeruginosa, Environ. Sci. Technol. 34 (2000) 610–615. [16] J.-S. Chang, J. Hong, Biosorption of mercury by the inactivated cells of Pseudomonas aeruginosa PU21(Rip64), Biotechnol. Bioeng. 44 (1994) 999–1006. [17] M.Z.-C. Hu, J.M. Norman, B.D. Faison, M.E. Reeves, Biosorption of uranium by Pseudomonas aeruginosa strain CSU: characterization and comparison studies, Biotechnol. Bioeng. 51 (1996) 237–247. [18] J.-L. Ramos, Pseudomonas, Kluwer Academic/Plenum Publishers, New York, 2004. [19] S.Y. Kang, J.U. Lee, K.W. Kim, Metal removal from wastewater by bacterial biosorption: kinetics and competition studies, Environ. Technol. 26 (2005) 615–624. [20] S.Y. Kang, J.U. Lee, K.W. Kim, A study of the biosorption characteristics of Co2+ in wastewater using Pseudomonas aeruginosa, Key Eng. Mater. 277–279 (2005) 418–423. [21] A.D. Eaton, L.S. Clesceri, A.E. Greenberg, Standard Methods for the Examination of Water and Wastewater, 17th ed., American Public Health Association (APHA), American Water Works Association (AWWA), Water Pollution Control Federation (WPCF), Washington, DC, 2000. [22] T.J. Beveridge, Role of cellular design in bacterial metal accumulation and mineralization, Annu. Rev. Microbiol. 43 (1989) 147–171. [23] A. Herbelin, J. Westall, FITEQL, A computational program for determination of chemical equilibrium constants from experimental data, Version 4.0, Report 99-01, Department of Chemistry, Oregon St. University, Corvallis, OR, USA, 1999. [24] J.B. Fein, C.J. Daughney, N. Yee, T.A. Davis, A chemical equilibrium model for metal adsorption onto bacterial surfaces, Geochim. Cosmochim. Acta 61 (1997) 3319–3328. [25] C.J. Daughney, J.B. Fein, N. Yee, A comparison of the thermodynamics of metal adsorption onto two common bacteria, Chem. Geol. 144 (1998) 161–176. [26] W. Stumm, J.J. Morgan, Aquatic Chemistry: Chemical Equilibria and Rates in Natural Waters, Wiley/Interscience, New York, 1996. [27] P.C. DeLeo, H.L. Ehrlich, Reduction of hexavalent chromium by Pseudomonas fluorescens LB300 in batch and continuous cultures, Appl. Microbiol. Biotechnol. 40 (1994) 756–759. [28] L. Philip, L. Iyengar, C. Venkobachar, Cr(VI) reduction by Bacillus coagulans isolated from contaminated soils, J. Environ. Eng. 124 (1998) 1165–1170. [29] L.H. Bopp, H.L. Ehrlich, Chromate resistance and reduction in Pseudomonas fluorescens strain LB 300, Arch. Microbiol. 150 (1988) 426– 431. [30] H. Shen, Y.-T. Wang, Biological reduction of chromium by E. coli, J. Environ. Eng. 120 (1994) 560–572. [31] J.B. Fein, D.A. Fowle, J. Cahill, K. Kemner, M. Boyanov, B. Bunker, Nonmetabolic reduction of Cr(VI) by bacterial surfaces under nutrient-absent conditions, Geomicobiol. J. 19 (2002) 369–382. [32] N. Yee, J.B. Fein, Cd adsorption onto bacterial surfaces: a universal adsorption edge? Geochim. Cosmochim. Acta 65 (2001) 2037–2042. [33] M.M. Figueira, B. Volesky, H.J. Mathieu, Instrumental analysis study of iron species biosorption by Sargassum biomass, Environ. Sci. Technol. 33 (1999) 1840–1846. [34] B. Volesky, Sorption and Biosorption, BV Sorbex, Inc., Canada, 2003.

Desalination 211 (2007) 156–163

Biological removal of hexavalent chromium in trickling filters operating with different filter media types E. Dermoua, A. Velissarioub, D. Xenosa, D.V. Vayenasa* a

Department of Environmental and Natural Resources Management,University of Ioannina, Seferi 2, Agrinio 30100, Greece Tel. +30 2641 074117; Fax +30 2641 0739576; email: [email protected] b Hellenic Aerospace Industry S.A., P.O. Box 23, GR 32009, Schimatari, Greece

Received 28 November 2005; revised 13 February 2006; accepted 16 February 2006

Abstract The purpose of the present study was to estimate Cr(VI) removal through biological mechanisms in biofilm reactors operated in SBR operating mode with recirculation using different filter media types. The choice of the support material is of great importance since chromate reduction is followed by the formation of sediments, which causes obstruction of the flow along the filter depth. In order to overcome this awkwardness, we tested the performance of two different support materials, i.e. plastic media and calcitic gravel, in pilot-scale trickling filters. The feed concentrations of Cr(VI) examined were about 5, 30 and 100 mg/l, while the concentration of the organic carbon was constant at 400 mg/l, in order to avoid carbon limitations in the bulk liquid. Plastic media showed better performance at the different Cr(VI) concentrations examined, compared to the gravel media. The removal rates for the plastic media achieved were 4.23±0.18, 3.62±0.1 and 3.3±0.08 g Cr(VI)/d, for feed concentration of Cr(VI) about 5, 30 and 100 mg/l respectively, while for the gravel media the corresponding removal rates were 4.11±0.09, 3.52±0.06 and 2.5±0.07 g Cr(VI)/d. Keywords: Hexavalent chromium; Biological reduction; Trickling filter; Plastic media; Gravel

*Corresponding author.

Presented at the 9th Environmental Science and Technology Symposium, September 1–3, 2005, Rhodes, Greece. Organized by the Global NEST organization and prepared with the editorial help of the University of Aegean, Mytilene, Greece and the University of Salerno, Fisciano (SA), Italy. 0011-9164/07/$– See front matter © 2007 Published by Elsevier B.V. doi:10.1016/j.desal.2006.02.090

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1. Introduction Hexavalent chromium contamination in the environment is a result of the extensive use of chromate and dichromate in numerous industries including stainless-steel production, leather tanning, electroplating, pigment fabrication, wood preserving, power plants and nuclear facilities [1]. Chromium can occur at several different oxidation states ranging from –2 to 6. However, only Cr(III) and Cr(VI) are the stable forms in the natural environments. Cr(VI) is rarely naturally occurring, is relatively soluble in aqueous systems and is readily transformed in groundwater [2]. Exposure to Cr(VI) poses an acute health risk because it is highly toxic and chronic exposure can lead to mutagenesis and carcinogenesis [3]. On the contrary Cr(III) is naturally occurring, is much less toxic and even essential to human glucidic metabolism, contributing to the glucose tolerance factor necessary for insulin-regulated metabolism [4]. The discharge of Cr(VI) to surface water is regulated to below 0.05 mg/l by the US EPA [5] and the European Union [6], while total Cr, Cr(III), Cr(VI) and its other forms, is regulated to below 2 mg/l. The most commonly used technology for treatment of heavy metals in wastewaters is chemical precipitation, which involves reduction of Cr(VI) to Cr(III) by a reducing agent under low-pH conditions and subsequent adjustment of solution pH to near neutral ranges to precipitate Cr(III) as hydroxides. However, this method results in high costs and secondary waste generation due to various reagents used [7]. Nowadays, microbiological detoxification of hexavalent chromium has gained importance due to the emphasis placed on protection of the environment [8]. Several bacterial species are capable of transforming Cr(VI), into the much less toxic and less mobile Cr(III), including both Gram-positive and Gram-negative species [9,10]. Bacteria may protect themselves from toxic substances in the environment by transforming toxic com-

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pounds through oxidation, reduction or methyliation into more volatile, less toxic or readily precipitating form [11]. Microorganisms interact with toxic metals and mediate their removal through different processes like bioaccumulation, biosorption and enzymatic reduction [12]. Metabolic enzymatic Cr(VI) reduction has been observed both in the presence and absence of O2 and Cr(VI) reduction rates are strongly dependent on pH, temperature, electron donor and Cr(VI) concentrations. The optimum pH conditions for metabolic enzymatic Cr(VI) reduction by bacteria appears to be close to pH 7 [13]. Biochemical studies of enzymatic Cr(VI) reduction reveal that Cr(VI)-reducing mechanisms are likely associated with bacterial electron transport systems [12]. Therefore, several researchers have proposed that Cr(VI) reduction is mediated by enzymes that are not substrate specific for Cr(VI) and that “chromate reductases” may be serendipitous contributors to Cr(VI) reduction while engaged in other primary physiological functions [1]. In literature, most of the previous studies on biological reduction of Cr(VI) were conducted under sterilized conditions (pure cultures of microorganisms) and resulted in very small removal rates [8,14,15]. There are many microorganisms which can reduce Cr(VI) and have been identified; examples include Shewanella putrefaciens MR-1, Pseudomonas fluorescens LB300, which reduces Cr(VI) under aerobic and anaerobic conditions, Agrobacterium radiobacter EPS-916, which is capable of reducing Cr(VI) aerobically, Enterobacter clocae strain HO1, which only reduces Cr(VI) under anaerobic conditions using a limited selection of electron donors. Also, a member of the obligatory anaerobic sulfate bacteria (D. vulgaris NCIMB 8303) has a well-established ability to reduce metals, including Cr(VI), if a complexing agent is incorporated to chelate Cr(III) [16,17]. However, there are a few recent studies investigating Cr(VI) reduction using mixed cultures of microorganisms [18–20].

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Batch reactors, continuous-flow and fixed film bioreactors were also used for biological reduction of Cr(VI) [8,13,21–23]. In a recent study [11] a pilot-scale trickling filter was constructed and tested for biological chromium (VI) removal from industrial wastewater. Three different operating modes were used to investigate the optimal performance and efficiency of the filter, i.e. batch, continuous and SBR with recirculation. The latter one was found to achieve removal rates up to 3.5 g Cr(VI)/d while aeration was taking place naturally without the use of any external mechanical means. In this study, an attempt was made to estimate the effect of the support media on Cr(VI) reduction rate. Mixed culture of microorganisms, originating from an industrial sludge, were used in pilot-scale trickling filters using two different filter media types, i.e. calcitic gravel and hollow plastic tubes, under SBR operation with recirculation, suggesting the most efficient support media type. 2. Materials and methods 2.1. Media The influent feed to the bioreactor was prepared by dissolving 1g NH4Cl, 0.2 g MgSO47H2O, 0.001 g FeSO47H2O, 0.001 g CaCl22H2O, 2.5 g CH3COONa3H2O and 0.5 g K2HPO4 in 1.0 l of tap water. 2.2. Reagents Stock Cr(VI) solution (500 mg/l) was prepared by dissolving 141.4 mg of 99.5% K2Cr2O7, previously dried at 103°C for 2 h, in Milli-Q water and diluting to 100 ml. Diphenyl carbazide solution was prepared by dissolving 250 mg of 1,5-diphenylcarbazide in 50 ml of HPLC-grade acetone and storing in a brown bottle, for Cr(VI) concentration measurements. 1,5-diphenylcarbazide was purchased from Fluka Chemical, potassium dichromate was purchased from Sigma Chemical Co.

2.3. Analytical methods During all experiments, hexavalent chromium concentration, pH, temperature, dissolved oxygen concentration and TOC measurements were made on a daily basis. Samples were filtered through 0.45 µm–Millipore filters (GN-6 Metricel Grid 47 mm, Pall Corporation). Hexavalent chromium concentration was determined by the 3500-Cr D Colorimetric method according to Standard Methods for the Examination of Water and Wastewater [24]. Total Organic Carbon measurements (TOC) were conducted in order to determine the feed sodium acetate concentration both in the batch reactor and the bulk liquid of the bioreactor, following the methods described in Standard Methods for the Examination of Water and Wastewater [24] by using, Total Organic Carbon Analyzer (TOC-VCSH, SHIMAZDU Corporation, JAPAN). The cell density of liquid culture was determined as optical density at 600 nm on a Jasco V-530, spectrophotometer. 2.4. Isolation and enrichment of indigenous bacteria Samples of industrial sludge were taken from the Hellenic Aerospace Industry S.A. In order to grow bacterial strains able to reduce hexavalent chromium, a sludge sample of 10 grams was added in a 2 l Erlenmeyer flask and was diluted in an acetate-minimal medium and concentrated chromium solution (in the form of K2Cr2O7 ) resulting in a final hexavalent chromium concentration of 50 mg/l. The final volume of the solution was 1 l. Acetate-minimal medium (AMM) was comprising (per litre) 1g NH4Cl, 0.2 g MgSO47H2O, 0.001 g FeSO 4 7H 2 O, 0.001 g CaCl 2 2H 2 O, 2.5 g CH3COONa3H2O, 0.5 g yeast extract and 0.5 g K2HPO4 in 1.0 l of tap water (micronutrients were supplied using tap water as diluent), while the final pH of the nutrient solution was adjusted to 7. The solution was kept under oxic conditions through aeration and mixing while nutrients and hexavalent chromium were added according to the

E. Dermou et al. / Desalination 211 (2007) 156–163 [3]

[4]

[5]

[6] [7]

[8]

[9]

[10]

[11]

[12]

[13]

[14]

[15]

[16]

S. Viamajala, B.M. Peyton, W.A. Apel and J.N. Petersen, Chromate reduction in Shewanella oneidensis MR1, Is an inducible process associated with anaerobic growth. Biotechnol. Prog., 18 (2002) 290–295. A.M. Zayed and N. Terry, Chromium in the environment: factors affecting biological remediation. Plant and Soil, 249 (2003) 139–156. A. Baral and R.D. Engelken, Chromium-based regulations and greening in metal finishing industries in the USA. Environ. Sci. Policy, 5(2) (2002) 121–133. EC-Official Journal of the European Communities, L330/32, December 12, 1998. T.E. Higgins, A.R. Halloran and J.C. Petura, Traditional and innovative treatment methods for Cr (VI) in soil. J. Soil Contam., 6 (1997) 767–797. A.R. Shakoori, M. Makhdoom and R.U. Haq, Hexavalent chromium reduction by a dichromate-resistant gram-positive bacterium isolated from effluents of tanneries. Appl. Microbiol. Biotechnol., 53 (2000) 348– 351. F.A.O. Camargo, B.C. Okeke, F.M. Bento and W.T. Frankenberger, Diversity of chromium-resistant bacteria isolated from soils contaminated with dichromate. Applied Soil Ecology, 29 (2005) 193–202. A. Pal and A.K. Paul, Aerobic chromate reduction by chromium-resistant bacteria isolated from serpentine soil. Microbiol. Res., 159 (2005) 347–354. E. Dermou, A. Velissariou, D. Xenos and D.V. Vayenas. Biological chromium (VI) reduction using a trickling filter, J. Hazard. Mater., B126 (2005) 78–85. T. Srinath, T. Verma, P.W. Ramteke and S.K. Garg, Chromium (VI) biosorption and bioaccumulation by chromate resistant bacteria. Chemosphere, 48 (2002) 427–435. J.B. Fein, D.A. Fowle, J. Cahill, K. Kemner, M. Boyanov and B. Bunker, Nonmetabolic reduction of Cr (VI) by bacterial surfaces under nutrient-absent conditions. Geomicrob. J., 19 (2002) 369–382. Y.T. Wang and C. Xiao, Factors affecting hexavalent chromium reduction in pure cultures of bacteria. Wat. Res., 29(11) (1995) 2467–2474. Y. Wang and H. Shen, Modelling Cr(VI) reduction by pure bacterial cultures, Wat. Res., 31(4) (1997) 727– 732. C.R. Myers, B.P. Carstens, W.E. Antholine and J.M. Myers, Chromium (VI) reductase activity is associated with the cytoplasmic membrane of anaerobically grown Shewanella putrefaciens MR-1. J. Appl. Microbiol., 88 (2000) 98–106.

163

[17] A.C. Humphries, K.P. Nott, L.D. Hall and L.E. Macaskie, Reduction of Cr (VI) by immobilized cells of Desulfovibrio vulgaris NCIMB 8303 and Microbacterium sp. NCIMB 13776. Biotech. Bioeng., 90 (2005) 589–596. [18] A.S. Stasinakis, N.S. Thomaidis, D. Mamais and T.D. Lekkas, Investigation of Cr(VI) reduction in continuous-flow activated sludge systems. Chemosphere, 57 (2004) 1069–1077. [19] A.S. Stasinakis, N.S. Thomaidis, D. Mamais, M. Karivali and T.D. Lekkas, Chromium species behaviour in the activated sludge process. Chemosphere, 52 (2003) 1059–1067. [20] Y. Chen and G. Gu, Preliminary studies on continuous chromium (VI) biological removal from wastewater by anaerobic–aerobic activated sludge process. Biores. Technol., 96 (2005) 1713–1721. [21] H. Shen and Y. Wang, Hexavalent chromium removal in two-stage bioreactor system. J. Environ. Eng., 121(11) (1995) 798–804. [22] E.M.N. Chirwa and Y. Wang, Hexavalent chromium reduction by Bacillus sp. in a packed-bed bioreactor. Environ. Sci. Technol., 31 (1997)1446–1451. [23] M. Megharaj, S. Avudainayagam and R. Naidu, Toxicity of hexavalent chromium and its reduction by bacteria isolated from soil contaminated with tannery waste. Current Microbiol., 47 (2003) 51–54. [24] Standard Methods for the Examination of Water and Wastewater, 17th ed., American Public Health Association, APHA/AWWA/WPCF, Washington, DC, 1989. [25] R. Marchesi, T. Sato, A.J. Weightman, T.A. Martin, J.C. Fry, S.J. Hiom and W.G. Wade, Design and evaluation of useful bacterium-specific PCR primers that amplify genes coding for bacterial 16S rRNA. Appl. Environ. Microbiol., 64 (1998) 795–799. [26] J.S.F. Altschul, T.L. Maden, A.A. Schaffer, J. Zhang, Z. Zhang, W. Miller and D.J. Lipman, Gapped BLAST and PSI-BLAST: a new generation of protein database search programs. Nucleic Acids Res., 25 (1997) 3389– 402. [27] P. Pattanapipitpaisal, N.L. Brown and L.E. Macaskie, Chromate reduction and 16S rRNA identification of bacteria isolated from a Cr(VI)-contaminated site. Appl. Microbiol. Biotechnol., 57 (2001) 257–261. [28] R.Francisco, M.C. Alpoim and P.V. Morais, Diversity of chromium-resistant and reducing bacteria in a chromium-contaminated sludge. J. Appl. Microbiol., 92 (2002) 837–843.

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Biosorption of chromium by Termitomyces clypeatus Sujoy K. Das 1 , Arun K. Guha ∗ Department of Biological Chemistry, Indian Association for the Cultivation of Science, 2A&B, Raja S.C. Mullick Road, Jadavpur, Kolkata 700032, India Received 14 May 2007; received in revised form 24 May 2007; accepted 25 May 2007 Available online 2 June 2007

Abstract The manuscript describes removal of chromium from aqueous solution by biomass of different moulds and yeasts. The biomass of Termitomyces clypeatus (TCB) is found to be the most effective of all the fungal species tested. The sorption of hexavalent chromium by live TCB depends on the pH of the solution, the optimum pH value being 3.0. The process follows Langmuir isotherm (regression coefficient 0.998, χ2 -square 5.03) model with uniform distribution over the surface which gets strong support from the X-ray elemental mapping of chromium adsorbed biomass. The amino, carboxyl, hydroxyl, and phosphate groups of the biomass are involved in chemical interaction with the chromate ion forming a cage like structure depicted by scanning electron microscopic (SEM) and Fourier transform infrared spectroscopic (FTIR) results. Desorption and FTIR studies also exhibited that Cr6+ is reduced to trivalent chromium on binding to the cell surface. The level of chromium concentration present in the effluent of tannery industries’ is reduced to a permissible limit using TCB as adsorbent. © 2007 Elsevier B.V. All rights reserved. Keywords: Sorption; Termitomyces clypeatus; FESEM; FTIR; Chromium; Isotherm model

1. Introduction Chromium, a toxic heavy metal, dissipates into environment as a result of various industrial activities [1,2] such as steel manufacturing, metal plating, mining, leather tanning, textile dying, cement industries etc. Of all the different oxidation states trivalent and hexavalent chromium exist as stable species. The hexavalent chromium species exists in aqueous solution as oxyanionic entities like chromate (CrO4 −2 ), bichromate (HCrO4 − ) and dichromate (Cr2 O7 −2 ), the relative distribution of which depends on the solution pH [3,4]. Two other forms of chromium, e.g. Cr3 O10 −2 , and Cr4 O13 −2 have also been detected in highly acidic medium [4]. The oxyanionic entities of Cr6+ do not bind to the negatively charged mineral surfaces, e.g. silica or clay, become highly mobile in the environment and soluble in a solution of neutral pH. In comparison, Cr3+ forms stable hydroxo complexes [e. g. Cr(OH)n (3−n)+ ] and the cationic Cr3+ having strong affinity for particle surfaces yields ∗

Corresponding author. Tel.: +91 33 2473 4971/5904x502; fax: +91 33 2473 2805. E-mail addresses: sujoy [email protected] (S.K. Das), [email protected], [email protected] (A.K. Guha). 1 Tel.: +91 33 2473 4971/5904x502; fax: +91 33 2473 2805. 0927-7765/$ – see front matter © 2007 Elsevier B.V. All rights reserved. doi:10.1016/j.colsurfb.2007.05.021

insoluble Cr(OH)3 at neutral pH, and becomes almost immobile in the environment [5,6]. Cr3+ is also an essential micronutrient compared with the toxic, mutagenic and carcinogenic hexavalent chromium [7] in addition to its reduced toxicity due to its low bioavailability. In view of toxicity and related environmental hazards the level of chromium in wastewater must be reduced to a permissible limit (5.0 mg/L and 0.5 mg/L for trivalent and hexavalent chromium, respectively) [8] before discharging into water bodies. The removal of chromium employing conventional methodologies [9,10] like ion exchange, chemical precipitation or reverse osmosis suffer from limitations like high operating cost, incomplete precipitation, sludge generation, etc. On the other hand biosorption is receiving increasing attention as an emerging technology for the removal of heavy metals from contaminated effluents. The process is based on the adsorption behavior of certain biological materials towards organic or inorganic substances from their solution. Different types of adsorbents [8,11–17] including activated carbon, sawdust, cactus leaves, lignin, spent grain, chitin, chitosan, jacobsite (MnFe2 O4 ), etc. used for the removal of chromium results in low removal requiring prolonged equilibrium time. Recently many efforts have been directed towards the development of specific biosorbents, the performance of which depends on the biomass characteristics and microenvironment of target metal ion solu-

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Fig. 1. Effect of pH on chromium adsorption by Termitomyces clypeatus biomass at 30 ◦ C (A), (䊉) Cr6+ uptake, () Cr3+ , () Zeta potential; adsorption isotherm of Cr6+ on T. clypeatus biomass (B), (䊉) experimental value, () non-linear Langmuir model, () non-linear Freundlich model; adsorption isotherm following the linear form of Langmuir model (C); adsorption isotherm following the linear form of Freundlich model (D). Data represent an average of four independent experiments ± S.D. shown by error bar.

solution, while maximum adsorption of Cr3+ took place at pH value 6.0 (Fig. 1A). Precipitation of Cr3+ as Cr(OH)3 at pH value > 6.0 restricted the experiment for solution having pH values beyond 6.0. Zeta potential measurements indicate that the overall surface charge remained positive at low pH values, which facilitated the attraction of negatively charged oxyanionic chromate ions. On the other hand, decreased adsorption of Cr3+ at pH 2.0–3.0 is attributed to the competition of the binding sites for H+ ions. From the experimental data, it is noted that at pH 2.0 adsorption of Cr6+ was less than 20% compared with that at pH 3.0. Similar behavior of hexavalent chromium adsorption on a variety of biosorbents including fungal biomasses was also reported earlier [33–35]. However, the rationale behind such influence of pH on the sorption process has not yet been adequately discussed. A number of factors are to be considered to understand the role of pH in the sorption behavior of chromium on TCB. Hexavalent chromium can be reduced to trivalent state both chemically [36] or enzymatically [37]. The cell wall of TCB contains different electron donors and their close proximity to aqueous chromate ions result in the formation of Cr3+ species. This reduction is facilitated at low pH values. On the other hand,

optimum pH of chromate reductase is around pH 7.0 and activity decreases significantly at low or high pH values [37]. So, enzymatic reduction of Cr6+ to Cr3+ can be ruled out at low pH. Accordingly we have measured the amount of both hexavalent and trivalent chromium in the solution after adsorption with TCB at pH 2.0, 3.0, and 4.0. The results presented in Table 2 show the presence of both hexavalent and trivalent chromium in the aqueous solution after sorption of chromate with TCB. This indicates that chromate being a strong oxidizing agent oxidizes some functional groups of biomass and in the process gets reduced to trivalent state. It was also observed that proportion of trivalent chromium to hexavalent chromium decreased with Table 2 Influence of pH on reduction of Cr6+ pH of the solution

Uptake (mg/g)

Total Cr (mg/L)

Cr6+ mg/L

Cr3+ mg/L

Cr3+ /Cr6+

2 3 4

8.77 11.15 7.03

25.69 7.25 39.61

3.25 1.51 24.36

22.44 5.74 15.25

6.9 3.8 0.63

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increasing in pH of the solution. Additionally, to confirm the presence of both the forms of chromium on the biomass, we have desorbed the chromium from the loaded biomass with 0.5 M HCl or 0.5 M NaOH and measured the amount of both forms of chromium in the eluent. The results show (data not shown) the presence of both trivalent and hexavalent chromium but the percentage of Cr3+ being much higher than that of Cr6+ in the HCl eluent while in NaOH eluent opposite results were observed. This further confirms our hypothesis that chromate ions after initial binding to the cell surface were reduced to Cr3+ . Thus, the optimum pH value (3.0) can be explained as; at low pH value positively charged functional groups adsorbed chromate ions through anion exchange mechanism, however, as soon as the pH was lowered below the optimum pH value, the chromate oxidize the biomass and produced Cr3+ ions [36,38]. These Cr3+ ions then competed for the binding sites with the protons via the cation exchange reaction resulting in low adsorption. The desorption of Cr3+ from the biomass at that low pH value also lead to low uptake of chromium [38]. Consequently, the sorption of Cr6+ from its aqueous solution by TCB was maximized at a pH value which was high enough for anion exchange as well as redox reaction to proceed simultaneously. Thus, we presumed that the influence of hydrogen ion concentration on the present sorption process is complex in nature and the noted optimum pH value is the resultant of all these factors. However, 25% of total adsorption also took place at a higher pH (6.0–7.0) value, although cell surface contained negative charges. Based on this experimental data we conclude that in addition to electrostatic forces of attraction, other factors such as reduction, precipitation, chemical interaction and physical forces such as hydrogen bonding and or ion–dipole interaction also involved in the sorption process. 3.4. Sorption isotherm Sorption isotherm is a prerequisite to describe the stoichiometric solute–solid interaction. A few parameters, such as maximum sorption capacity, are important for optimizing the design of sorption system, and analyzed using linearized forms of the isotherm models of Langmuir [39] and Glastone [40]. The result shows that the present sorption process of chromium on the live TCB followed the Langmuir model (Fig. 1, panels B and C) indicating chemisorption and monolayer coverage of sorbate on the sorbent. The theoretical monolayer saturation capacity (Qmax ) of the sorbate on the sorbent calculated (54.05 mg/g) from the slope of the linearized curve of Langmuir model (Fig. 1C) was very close to that obtained experimentally from isotherm study (53.95 mg/g) (Fig. 1B), which is higher than those reported for other types of biosorbents [12–14,16,41]. Sorption capacity of live TCB towards chromium was found to be higher than that reported for activated carbon (15.47 mg/g) [18], though it covers much higher BET surface area (500–3000 m2 /g) [42] compared to that of TCB (15.4 m2 /g). Therefore, we summarized that besides BET surface area other properties of the sorbent such as functional groups of the biomass played important roles in the studied sorption process. The sorption phenomenon did not fit well with the Freundlich isotherm

model (Fig. 1, panels B and D) exhibiting deviation from linearity over the entire concentration range. 3.5. Model analysis Linear coefficients of determination, r2 , and non-linear χ2 square test, χ2 may be used to evaluate the ‘goodness of fit’ of curves to the data they summarize and may be used to estimate the probability of obtaining any series of deviations of observed values from predicted values [43]. The value of linear coefficient of determination, r2 , represents the percentage of variability in the dependent variable that has been explained by the regression line and may vary from 0 to 1. If there is no relationship between the predicted values and actual values, the coefficient of determination is zero or very low, a perfect fit gives a coefficient of 1.0. On the other hand χ2 -square test, χ2 is basically the sum of the squares of the differences between the experimental data and data obtained from the models, with each square difference divided by the corresponding data obtained by calculation from the models. This can be represented mathematically as χ2 =

 (qe − qe,m )2 qe,m

where qe,m is equilibrium uptake obtained by calculation from the model (mg/g) and qe is the experimental data of the equilibrium uptake (mg/g). χ2 -square, χ2 will be small number if the experimental data are close to that obtained from model and will be bigger if they differ. Therefore, it is necessary to analyze the data sets using both linear coefficient of determination, and non-linear χ2 -square test to establish the best fit isotherm model for the adsorption system. The coefficient of determination, r2 , and Chi square test, χ2 , were determined in the range of the whole metal ion concentration. The linear regression coefficient, r2 , values were 0.998 and 0.965, respectively for Langmuir and Freundlich models. On the other hand the χ2 -square, χ2 , values of 5.03 and 18.53 for Langmuir and Freundlich, isotherm, respectively with 9 degrees of freedom (φ) corresponds to the significance level between 75–90% and 2.5–5% according Fischer and Yates chart [44]. Thus, Langmuir isotherm model appears to be the best fit for the present adsorption process. X-ray elemental mapping of the biomass after chromium sorption (Fig. 2) depicts a uniform distribution of chromium over the entire surface area which further supports the observed Langmurian behavior of sorbate on the sorbent surface. 3.6. Chemical characterization of metal ion adsorption In living cells the sorption mechanisms include both metabolism dependent and independent processes. Metabolism independent uptake process essentially involves cell surface binding through ionic and chemical interaction, while dependent process deals with the binding of both the surfaces followed by intracellular accumulation [21,45]. The cell wall of the fungal biomass generally contains large amounts of chitin, chitosan, glucan and mannan as well as small amounts of glycoprotein

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and Prof. A. Dasgupta (Department of Biochemistry, Calcutta University, Kolkata) for providing Atomic Absorption Spectroscopic analysis and Zeta potential measurement facility, respectively. We specially thank Dr. P.C. Banerjee (Indian Institute of Chemical Biology, Kolkata) and Dr. A.R. Das (Polymer Science Unit of our Institute) for their cooperation in various ways throughout the work. References [1] [2] [3] [4] [5] [6] [7]

[8] [9] [10] [11] [12] [13] [14] [15] [16] [17] [18] [19] [20] [21] [22]

[23] [24] [25]

S.A. Abbasi, R. Soni, J. Inst. Eng. (Env. Eng. Div.) 65 (1985) 113. M.J. Udy, Chromium, Reinhold Publishing Corporation, New York, 1956. R.K. Tandon, P.C. Crisp, J. Ellis, R.S. Baker, Talanta 31 (1984) 227. M. Cieslak-Golonka, Coord. Chem. Rev. 109 (1991) 223. C.F. Baes, R.E. Mesmer, The Hydrolysis of Cations, John Wiley & Sons, New York, 1976. D. Rai, B.M. Sass, D.A. Moore, Inorg. Chem. 26 (1987) 345. J.N. Hamilton, K.E. Wetterhan, in: H.G. Seiller, H. Sigel (Eds.), Handbook on Toxicity of Inorganic Compounds, Marcel Dekker, Inc., New York, 1988, p. 239. F.N. Acar, E. Malkoc, Bioresour. Technol. 94 (2004) 13. X. Zhou, T. Korenaga, T. Moriwake, S. Shinoda, Water Res. 27 (1993) 1049. G. Tiranvanti, D. Petruzzelli, R. Passino, Water Sci. Technol. 36 (1997) 197. P-G.G. Burnett, C.J. Daughney, D. Peak, Geochim. Cosmochim. Acta 70 (2006) 5253. M. Dakiky, M. Khamis, A. Manassra, M. Mer’eb, Adv. Environ. Res. 6 (2002) 533. J. Hu, I.M.C. Lo, G. Chen, Langmuir 21 (2005) 11173. S.B. Lalvani, A. Hubner, T.S. Wiltowski, Energ. Source. 22 (2000) 45. R. Maruca, B.J. Suder, J.P. Weightman, J. Appl. Polym. Sci. 27 (1982) 4827. B. Sandhya, A.K. Tonni, Chemosphere 54 (2004) 951. M. Tsezos, X. Wang, J. Chem. Technol. Biotechnol. 50 (1991) 507. Y. Andress, A.C. Texier, P. Le Cloirec, Environ. Technol. 24 (2003) 1367. G.M. Gadd, C. White, Trends Biotechnol. 11 (1993) 353. C. Lacina, G. Germain, A.N. Spiros, Afr. J. Biotechnol. 2 (2003) 620. S.P. Mishra, G. Roy Chowdhury, Min. Process Extract Metall. Rev. 14 (1995) 111. A.D. Eaten, L.S. Clesceri, A.E. Greenberg, Standard Methods for the Examination of Water and Wastewater, American Public Health Association (APHA), AWWA, WPCF, Washington, DC, 1995, p. 4. S.K. Das, J. Bhowal, A.R. Das, A.K. Guha, Langmuir 22 (2006) 7265. L.M. He, B.M. Tebo, Appl. Environ. Microbiol. 64 (1998) 1123. N. Li, R. Bai, Sep. Purif. Technol. 42 (2005) 237.

[26] R.L. Smith, F.E. Strohmair, R.S. Oremland, Arch. Microbiol. 141 (1985) 8. [27] N. Goyal, S.C. Jain, U.C. Banerjee, Adv. Environ. Res. 7 (2003) 311. [28] Z. Aksu, Process Biochem. 40 (2005) 997. [29] R.O. James, G.A. Parks, Surf. Colloid Sci. 12 (1982) 119. [30] B.R. Sudha, T.E. Abraham, Bioresour. Technol. 79 (2001) 73. [31] F. Vegio, F. Beolchini, Hydrometallurgy 44 (1997) 301. [32] L.H. Tan, J.P. Chen, Y.P. Ting, Biosorption of chromium (VI) by algal biomass., in: Proceedings of the 15th International Biohydrometallurgy Symposium, Athens, Hellas, 2003, p. 807. [33] S.R. Bai, T.E. Abraham, Bioresour. Technol. 79 (2001) 73. [34] D.C. Sharma, C.F. Forster, Water Res. 27 (1993) 1201. [35] Y. Sag, T. Kutsal, Process Biochem. 33 (1998) 571. [36] J.B. Fein, D.A. Fowle, J. Cahil, K. Kemner, M. Boyanov, B. Bunker, Geomicrobiol. J. 19 (2002) 369. [37] P.-C. Wang, T. Mori, K. Komori, M. Sasatsu, K. Toda, H. Ohtake, Appl. Environ. Microbiol. 55 (1989) 1665. [38] D. Kratochvil, P. Pimentel, B. Volesky, Environ. Sci. Technol. 32 (1998) 2693. [39] I. Langmuir, J. Am. Chem. Soc. 38 (1916) 2221. [40] S. Glastone, Textbook of Physical Chemistry, 2nd ed., MacMillan Publishing Co., New York, 1962, p. 1196. [41] F.-K. Zeljka, L. Sipos, F. Briski, Food Technol. Biotechnol. 38 (2000) 211. [42] B. Tansel, P. Nagarajan, Adv. Environ. Res. 8 (2004) 411. [43] A.E. Lewis, Biostatistics, East-West Press Student Edition, East-West Press Pvt. Ltd., New Delhi, 1971. [44] R.A. Fisher, F. Yates, Statistical Tables for Biological, Agricultural and Medical Research, 6th ed., Oliver and Boyd, Edinburgh, 1963. [45] S.E. Shumate II., G.N. Standberg, in: M. Moo-Young (Ed.), Comprehensive Biotechnology. The Principles, Applications and Regulations of Biotechnology in Industry, Agriculture and Medicine, Pergamon Press, 1983, p. 235. [46] J. Ruiz-Herrera, Fungal Cell Wall: Structure, Synthesis and Assembly, CRC Press, Boca Raton, FL, 1992. [47] J.M. Tobin, D.G. Cooper, R.J. Neufeld, Appl. Environ. Microbiol. 47 (1984) 821. [48] J.R. Heber, R. Stevenson, O. Boldman, Science 116 (1952) 111. [49] J. Schmitt, H.C. Flemming, Int. Biodeter. Biodegr. Sci. 41 (1998) 1. [50] G. Naja, C. Mustin, J. Berthelin, B. Volesky, J. Colloid Int. Sci. 292 (2005) 537. [51] R. Aravindam, B. Madhan, J.R. Rao, B.N. Nair, T. Ramasami, Environ. Sci. Technol. 38 (2004) 300. [52] Y.G. Ko, U.S. Choi, T.K. Kim, D.J. Ahn, Y.J. Chun, Macromol. Rapid Commun. 23 (2002) 535. [53] J.N. Steel, J.L. Atwood, Supramolecular Chemistry, 1st ed., John Wiley & Sons, Chichester, 2000, p. 9. [54] M.E. Huston, E.U. Akkaya, A.W. Czarnik, J. Am. Chem. Soc. 111 (1989) 8735.

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Low temperature reduction of hexavalent chromium by a microbial enrichment consortium and a novel strain of Arthrobacter aurescens Rene' N Horton*1, William A Apel2, Vicki S Thompson2 and Peter P Sheridan1 Address: 1Department of Biological Sciences, Idaho State University, Campus Box 8007, Pocatello, ID USA 83209-8007 and 2Idaho National Laboratory, P.O. Box 1625, Idaho Falls, ID USA 83415 Email: Rene' N Horton* - [email protected]; William A Apel - [email protected]; Vicki S Thompson - [email protected]; Peter P Sheridan - [email protected] * Corresponding author

Published: 25 January 2006 BMC Microbiology 2006, 6:5

doi:10.1186/1471-2180-6-5

Received: 26 May 2005 Accepted: 25 January 2006

This article is available from: http://www.biomedcentral.com/1471-2180/6/5 © 2006 Horton et al; licensee BioMed Central Ltd. This is an Open Access article distributed under the terms of the Creative Commons Attribution License (http://creativecommons.org/licenses/by/2.0), which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited.

Abstract Background: Chromium is a transition metal most commonly found in the environment in its trivalent [Cr(III)] and hexavalent [Cr(VI)] forms. The EPA maximum total chromium contaminant level for drinking water is 0.1 mg/l (0.1 ppm). Many water sources, especially underground sources, are at low temperatures (less than or equal to 15 Centigrade) year round. It is important to evaluate the possibility of microbial remediation of Cr(VI) contamination using microorganisms adapted to these low temperatures (psychrophiles). Results: Core samples obtained from a Cr(VI) contaminated aquifer at the Hanford facility in Washington were enriched in Vogel Bonner medium at 10 Centigrade with 0, 25, 50, 100, 200, 400 and 1000 mg/l Cr(VI). The extent of Cr(VI) reduction was evaluated using the diphenyl carbazide assay. Resistance to Cr(VI) up to and including 1000 mg/l Cr(VI) was observed in the consortium experiments. Reduction was slow or not observed at and above 100 mg/l Cr(VI) using the enrichment consortium. Average time to complete reduction of Cr(VI) in the 30 and 60 mg/l Cr(VI) cultures of the consortium was 8 and 17 days, respectively at 10 Centigrade. Lyophilized consortium cells did not demonstrate adsorption of Cr(VI) over a 24 hour period. Successful isolation of a Cr(VI) reducing organism (designated P4) from the consortium was confirmed by 16S rDNA amplification and sequencing. Average time to complete reduction of Cr(VI) at 10 Centigrade in the 25 and 50 mg/l Cr(VI) cultures of the isolate P4 was 3 and 5 days, respectively. The 16S rDNA sequence from isolate P4 identified this organism as a strain of Arthrobacter aurescens, a species that has not previously been shown to be capable of low temperature Cr(VI) reduction. Conclusion: A. aurescens, indigenous to the subsurface, has the potential to be a predominant metal reducer in enhanced, in situ subsurface bioremediation efforts involving Cr(VI) and possibly other heavy metals and radionuclides.

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could be useful in bioremediation using nutrient addition.

6+ reduction Figure 10°C 2 by isolate P4 under aerobic conditions at Cr Cr6+ reduction by isolate P4 under aerobic conditions at 10°C.

A number of studies suggest both growth-dependent and growth-independent chromium reduction [20,29,40]. In either case, chromium reduction does seem to be biomass dependent in our study as well as in others [21,41]. The lag at the beginning of the consortium reduction experiments as well as observations of increased turbidity throughout the experiments suggests that adequate cell biomass must be produced before reduction begins in earnest. Bopp and Ehrlich [28] showed that higher concentrations (1000 mg/l) of Cr6+ produced a much longer lag phase and a significantly lower final yield of biomass than lower concentrations. The reduced biomass would also contribute to the lack of complete reduction found at higher concentrations in many studies [22,25] as well as in the higher concentrations tested in our lab (data not shown). Previous studies using cellular biomass grown on uncontaminated substrates to test Cr6+ reduction greatly decreased the amount of time required to completely reduce Cr6+ [21,39], similar to our findings with the isolate P4 reduction experiments (Figure 2). Increased turbidity after only 24 hours in R2 broth at 10°C (grown aerobically) and the achievement of stationary phase (as determined by absorbance readings, 1:10 dilution in R2 broth, OD = 0.16) after 3 days suggests that P4 is relatively fast growing. P4 grew at 10, 18, and 25°C but not at 37°C suggesting the isolate is a true psychrophile. Growth appeared fastest at 18°C. The ability to increase biomass in a short time given the proper nutrients suggests that P4

The enrichment culture and isolate P4 consistently reduced Cr6+ in VB medium up to concentrations of 60 mg/l Cr6+. Higher concentrations seemed to inhibit reduction, although growth was slower but still observed as turbidity in the enrichments (data not shown). Dilution of the Cr6+ at 1000 mg/L may have affected the limited range of the diphenyl carbazide assay, causing the appearance of the lack of reduction at higher concentrations. Both the consortium and isolate P4 showed significant tolerance of Cr6+ up to concentrations of 1000 mg/l (data not shown) as well as measurable reduction over short periods of time at concentrations up to 60 mg/l Cr6+. This tolerance is greater than or comparable to most mesophilic microorganisms tested, such as Pseudomonas fluorescens at 53.5 mg/l [27] and Bacillus sp. at 500 mg/l [42]. Furthermore, the isolate P4 and consortium reductions presented here occurred at temperatures close to 30°C lower than in the studies using mesophilic organisms, suggesting that the enzyme(s) responsible for the reduction are truly coldactive. Complete reduction was observed in all experiments (both consortium and isolate P4) with concentrations of Cr6+ up to 60 mg/l (Figures 1&2) suggesting that complete reduction in the environment is also possible. The lack of reduction in the sterile controls along with the lack of Cr6+ adsorption to cell biomass in the three adsorption experiments suggests that the members of the enrichment community (which included isolate P4) were responsible for the reduction of Cr6+. Since most aquifers contaminated with Cr6+ have levels below 60 mg/l, these experiments would also suggest remediation of the lower levels of Cr6+ contamination present in aquifers is possible. Bioremediation literature suggests low levels of contamination are very difficult to completely remediate. Lack of induction of enzyme systems at low contaminant concentrations and problems with availability of contaminants bound to organics and sequestered in other matrices all contribute to persistence of contaminants in the environment. It has also been suggested that indigenous microorganisms may be more successful in reducing low contaminant concentrations [8]. The complete reduction of Cr6+ at 10°C in this study using an indigenous member of the Hanford microbial community and past studies with indigenous mesophilic microorganisms suggest that there are environmental candidates for reduction of the low levels of contamination usually found in aquifers [23,43]. Studies have shown Arthrobacter species adsorbing Fe, Cd, and Cu, but not Cr [44,45]. Chromium has, however, been shown to adsorb to both Shewanella and Bacillus species [46]. Adsorption studies performed on the Hanford

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39. 40.

41. 42. 43. 44.

45. 46. 47.

48.

49. 50.

51. 52.

53.

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Shen H, Wang YT: Modeling hexavalent chromium reduction in Escherichia coli 33456. Biotechnology and Bioengineering 1994, 43:293-300. Shen H, Wang YT: Simultaneous chromium reduction and phenol degradation in a coculture of Escherichia coli ATCC 33456 and Pseudomonas putida DMP-1. Appl Environ Microbiol 1995, 61:2754-2758. Phillip L, Venkobachar C, Iyengar L: Immobilized microbial reactor for the biotransformation of hexavalent chromium. International Journal of Environment and Pollution 1999, 11:202-210. Chirwa EMN, Wang Y: Hexavalent Chromium reduction by Bacillus sp. in a packed-bed bioreactor. Environ Sci Technol 1997, 31:1446-1451. Bader JL, Gonzalez G, Goodell PC, Ali AS, Pillai SD: Aerobic reduction of hexavalent chromim in soil by indigenous microorganisms. Bioremediation Journal 1999, 3:201-211. Pagnanelli F, Petrangeli Papini M, Toro L, Trifoni M, Veglio F: Biosorption of metal ions on Arthrobacter sp.: Biomass characterization and biosorption modeling. Environ Sci Technol 2000, 34:2773-2778. Beolchini F, Pagnanelli F, Veglio F: Modeling of copper biosorption by Arthrobacter sp. in a UF/MF membrane reactor. Environ Sci Technol 2001, 35:3048-3054. Fein JB, Fowle DA, Cahill J, Kemner K, Boyanov M, Bunker B: Nonmetabolic reduction of Cr(VI) by bacterial surfaces under nutrient-absent conditions. Geomicrobiol J 2002, 19:369-382. Holman HN, Perry DL, Martin MC, Lamble GM, McKinney WR, Hunter-Cevera JC: Real-time characterization of biogeochemical reduction of Cr(VI) on basalt surfaces by SR-FTIR imaging. Geomicrobiology 1999, 16:307-324. Benyehuda G, Coombs J, Ward PL, Balkwill D, Barkay T: Metal resistance among aerobic chemoheterotrophic bacteria from the deep terrestrial subsurface. Can J Microbiol 2003, 49:151-156. Margesin R, Schinner F: Bacterial heavy metal-toleranceextreme resistance to nickel in Arthrobacter spp. strains. J Basic Microbiol 1996, 36:269-282. Fries MR, Zhou J, Chee-Sanford J, Tiedje JM: Isolation, characterization, and distribution of denitrifying toluene degraders from a variety of habitats. Appl Environ Microbiol 1994, 60:2802-2810. Greenberg A, Clescerl L, Eaton A: Standard Methods for the Examination of Water and Wastewater. 18th Edition edition. Washington, D.C., American Public Health Association; 1992. Turick CE, Apel WA, Carmiol NS: Isolation of hexavalent chromium-reducing anaerobes from hexavalent-chromium-contaminated and noncontaminated environments. Appl Microbiol Biotechnol 1996, 44:683-688. Sheridan PP, Loveland-Curtze J, Miteva VI, Brenchley JE: Rhodoglobus vestalii gen. nov., sp. nov., a novel psychrophilic organism isolated from an Antarctic Dry Valley lake. Int J Syst Evol Microbiol 2003, 53:985-994.

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Journal of Chromatography A, 1137 (2006) 180–187

Analysis of bacteria degradation products of methyl parathion by liquid chromatography/electrospray time-of-flight mass spectrometry and gas chromatography/mass spectrometry Jie Liu a , Ling Wang b , Li Zheng c , Xiaoru Wang a,c , Frank S.C. Lee c,∗ a

Department of Chemistry and The Key Laboratory of Analytical Science of MOE, College of Chemistry and Chemical Engineering, Xiamen University, Xiamen 361005, China b College of Physical and Environmental Oceanography, Ocean University of China, Qingdao 266003, China c Key Lab on Analytical Technology Development and Standardization of Chinese Medicines, First Institute Oceanography of SOA, Qingdao 266061, China Received 9 March 2006; received in revised form 31 July 2006; accepted 6 October 2006

Abstract The biodegradation of the organophosphorus insecticide methyl parathion (MP) in aqueous environment by bacteria isolated from river sediment has been studied. Two species of bacteria which show strong MP degradation ability are identified as Shewanella and Vibrio parahaemolyticus. The biodegradation of MP proceeded rapidly with the formation of a series of intermediate products, which were analyzed using a combination of GC/MS and HPLC/ESI-TOFMS techniques. The major products tentatively identified include a series of reduced products of MP. Results demonstrate that the coupling of TOFMS to HPLC enhances further the capability of LC–MS in the identification of polar organic species in complex environmental samples. Degradation pathways leading to the formation of these products are proposed which involves first the reduction of nitro to amino group in MP, followed by combination with some intrinsic matters of bacteria. The mechanism and products from biodegradation are quite different from those of photocatalytic process for which the main intermediates included methyl paraoxon and 4-nitrophenol. © 2006 Elsevier B.V. All rights reserved. Keywords: Methyl parathion; Biodegradation; Bacteria; ESI-TOFMS

1. Introduction Organophosphorus compounds (OPs) are widely used as pesticides, insecticides in agricultural as well as non-agricultural practices. They have mostly replaced organochlorine (OCs) compounds in such applications because compared to the OCs, OPs are less bioaccumulative and more readily degradable in the environment [1]. Currently OPs account for about one-third of the total pesticide consumption in the world [2]. OPs are known to inhibit the activity of acetylcholinesterase (AChE), with subsequent accumulation of acetylcholine at nerve endings, causing major acute toxic effect [3]. The widespread contamination of soils, sediments and aquatic environment by OPs thus creates different set of environmental problems associated

∗

Corresponding author. Tel.: +86 532 88963253; fax: +86 532 88963253. E-mail address: [email protected] (F.S.C. Lee).

0021-9673/$ – see front matter © 2006 Elsevier B.V. All rights reserved. doi:10.1016/j.chroma.2006.10.091

with their acute toxicities and poorly understood degradation pathways. Our focus in this paper is the fate of OPs in the aqueous environment, in which they are known to degrade spontaneously through different pathways including hydrolysis, photolytic oxidation, microbial transformations and other biological processes. Among these processes, biodegradation is particularly attractive because of its effectiveness and low cost. OPs can be biodegraded by plants, algae, fungi and bacteria, and the key players in these processes are enzymes, such as hydrolase and oxidase [4–7]. The analysis of OPs in water samples is generally performed by solid-phase extraction (SPE) [8–11] followed by gas chromatography (GC) with different types of detectors [1,12,13]. However, some OPs and especially their transformation products are unsuitable for GC analysis because of their thermally labile, highly polar and non-volatile properties. Liquid chromatography–mass spectrometry (LC–MS) in these cases

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metric analysis. The temperatures of the ion source and the quadrupole mass analyzer were 230 and 150 ◦ C, respectively. Scan mode was chosen and the range of m/z was from 50 to 450. Split injection (20:1) was used, and the MS data acquisition was set at 3 min post-injection after the elution of the solvent peak. 2.5. HPLC/DAD and LC/ESI-TOFMS analysis HPLC/DAD analysis was carried out using an Agilent 1100 series HPLC coupled with a G1315B UV–vis DAD. The analytical column was a Alltech C18 reversed-phase column (4.6 mm × 150 mm, 5 ␮m). The DAD detector was set at the wavelength range from 190 to 400 nm. Gradient elution was used in HPLC runs. The gradient was formed by varying the proportion of water (A) and acetonitrile (B), each containing 0.1% (g/g) ammonium formate in order to improve the ionization of analytes and make them amenable to MS detection. The solvent programming involved first an isocratic run from 0 to 5 min with a mixture of 10% of B in A; followed by solvent program from 10 to 50% B in A in the next 5–20 min. The flow rated was constant at 0.8 mL/min, of which one half of the flow was directed into the MS ionization chamber. The MS detector was a Agilent G1969A TOFMS combined with electrospray (ESI) as the ion source. The drying gas temperature was set at 300 ◦ C and the velocity of the drying gas was 10 L/min. The ion polarity was positive, and the capillary voltage was set to 4000 V. Two different collision induced dissociation (CID) voltages, 50 and 150 V, were used in order to provide more structural information by MS analysis. The m/z acquisition range was set at 10–1000. This TOFMS possesses high mass resolving power of more than 10,000, which means that the deviation in a measured mass can be controlled to be less than 5 ppm in terms of m/m. The acquisition rate for this TOFMS is 10000 times per second. Ions with m/z 121.0509 and 922.0098 were used as reference ions in mass measurement in order to eliminate systematic errors. 3. Results and discussion 3.1. Identification of bacteria strains From the results of the degradation experiment, two bacteria strains showed significant MP degrading activity. They were marked as L-10 and S-2. After extraction, PCR amplification and sequencing of their 16S rRNA genes [24], they were identified to belong to the genus Shewanella and Vibrio parahaemolyticus, respectively. Some Shewanella species have been confirmed to be a kind of metal ion reducing bacteria capable of detoxicating metals from high valence to low valence state, e.g., U(VI) to U(III) or Cr(VI) to Cr(III) [25–27], and V. parahaemolyticus is known to cause gastrointestinal illness in humans [28]. 3.2. LC/ESI-TOFMS analysis of degradation products As described earlier in Section 2, three fractions were separated from the samples after biodegradation and they were: the soluble components extracted by SPE, the supernatant clear

Fig. 2. HPLC/DAD chromatograms of MP: (a) standard and (b) degradation without bacteria present for a week at room temperature.

solution after adsorption by SPE cartridge, and the acetonitrile extract of the solids portion from centrifugation. Preliminary screening of the three samples showed that most of the OP degradation products were in the first fraction, i.e., the soluble components extracted by SPE. Thus, later analytical effort was focused primarily on this product fraction. In Fig. 2, the HPLCUV chromatograms of a fresh standard MP solution and the same standard solution after 1 week standing at room temperature are compared. In the absence of bacteria, the MP is shown to be stable without the production of detectable degradation products. The MP peak (18.33 min) in the chromatograms was identified by both standard calibration and TOFMS spectrum. The process described below exemplifies a typical peak identification process in TOFMS analysis. The exact mass of peak 1 was measured as m/z 264.0089 (Fig. 3a), resulting from the protonation adduction of the molecular ion (mass 263). Data were processed using Analyst QS software from Agilent TOFMS. First, parameters, such as elemental composition and number limit, mass error tolerance and number of charges must be pre-set based on the background knowledge. For this TOFMS, mass error tolerance can be set at 5 ppm, which represents the uncertainties imposed by the systematic error. For analysis of MP and its degradation products, elements C, H, O, N, P, S were chosen as the possible elements. Besides mass measurement, the isotopic mass distributions also provide information about the variety and number of elements. The software in TOFMS is capable of calculating all possible molecular formula satisfying the observed data. Compared with other possible assignments, C8 H11 NO5 PS has the least mass error. Furthermore, its theoretical isotopic mass distributions match excellently with the one measured by TOFMS in both intensity and m/z position (Fig. 3b). Combining this information with double bond equivalent (DBE) and the MS fragment pattern in GC/MS (Table 2), the analyte can be unequivocally identified as MP. The other m/z 281.0352 (Fig. 3a) also can be identified as formula C8 H14 N2 O5 PS (Fig. 3c), which is the ammonium adduct of MP because of the presence of 0.1% ammonium formate in the mobile phase. Fig. 4 shows the results of MP degraded by bacteria for 1 week. Here MP peak (peak 1) disappears almost completely. The estimated % of parent species degraded based on relative

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3.4. Mechanism pathways of OPs degradation The bio-transformation of MP generally proceed in two phases as same as those commonly observed in the metabolism of extrinsic matters in vivo [31]. In phase one, oxidation, reduction and hydrolysis are the main reactions. Afterwards, the metabolites will combine with some intrinsic matters through hydroxyl, amino or carboxyl linkage. During the whole metabolic process, the molecular backbone of these extrinsic compounds maintain their integrity, with changes occur only in the substituent functionalities. Combining information obtained from GC/MS and LC/TOFMS analysis, the biodegradation pathways for MP can be proposed as is depicted in Fig. 7. In the proposed scheme, reduction reactions are the most important process in phase one, and the major reactions in MP involve the reduction of the nitro to amino group, followed with the reactions of amino group with other intrinsic matters. The major acute toxic effect of OPs is the inhibition of acetylcholinesterase, a serine hydrolase containing a catalytic triad—serine, histidine and an acidic residue [32]. The OPs undergo nucleophilic attack by a reactive serine assisted by a general base (histidine) in the enzyme active site to render the phosphoryl-AChE, causing the inhibition of AChE [33]. The structure of P S or P O is the most important to keep the P atom electrophilic, which is a necessary condition for the toxicity of OPs. The MP intermediates we observed here all contain the thiophosphoric moiety, and are therefore likely to remain toxic. However, their toxicities are likely to be reduced somewhat from the parent MP because of the conversion of nitrobenzene moiety to aniline moiety linking to the P atom during bio-transformation. This is because the resulted aniline moiety is more electro-donating than the original nitrobenzene moiety, making the P atom less likely to be attacked by a nucleophilic Ser in the enzyme active site, causing less inhibition of AChE. 4. Conclusion Our findings demonstrated that the two strains of bacteria Shewanell and V. parahaemolyticus are able to degrade MP with high efficiency. After 1 week degradation, MP disappeared almost completely with the formation of several degradation products. The molecular formulas and structures of these biodegradation intermediates have been analyzed via a combination of LC/TOFMS and GC/MS methods, and the major ones are tentatively identified as a series of reduced products of MP. The high precision mass measurements provided by TOFMS are able to suggest reasonable structural assignments for these species, which are of great value in their eventual unequivocal identification supplemented by other spectroscopic analyses, such as NMR. Degradation pathways leading to the formation of these intermediates are proposed. These degradation products are likely to remain toxic because of the persistence of the thiophosphoric moiety. However, the toxicity could be somewhat reduced because the dominant processes involved in the degradation processes is the conversion of nitro to amino

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groups in the MP parent molecule, and thus the weakening of the nucleophilicity of the P atom, which is known to be responsible for MP toxicity. Acknowledgement The authors thank Qingdao “2004 JiangCai Plan” (04-3JJ-11) and Agilent Technologies Co., Ltd. (China) to provide HPLC/TOFMS and GC/MS system for our lab. References [1] [2] [3] [4] [5]

[6] [7] [8] [9] [10] [11] [12] [13] [14] [15] [16] [17] [18] [19] [20] [21]

[22] [23] [24] [25] [26] [27]

[28] [29] [30] [31] [32] [33]

M. Bavcon, P. Trebse, Chemosphere 50 (2003) 595. C. Zamy, P. Mazellier, B. Legube, Water Res. 38 (2004) 2304. M. Jokanovic, Toxicology 166 (2001) 139. J.W. Kim, E.I. Rainina, W.W. Mulbry, C.R. Engler, J.R. Wild, Biotechnol. Prog. 18 (2002) 429. G. Amitai, R. Adani, G. Sod-Moriah, I. Rabinovitz, A. Vincze, H. Leader, B. Chefetz, L. Leibovitz-Persky, D. Friesem, Y. Hadar, FEBS Lett. 438 (1998) 195. K. Lai, N.J. Stolowich, J.R. Wild, Arch. Biochem. Biophys. 318 (1995) 59. Y. Lei, A. Mulchandani, W. Chen, Biotechnol. Prog. 21 (2005) 678. S. Chiron, A. Rodriguez, A. Fernandez-Alba, J. Chromatogr. A 823 (1998) 98. I. Tolosa, J.W. Readman, L.D. Mee, J. Chromatogr. A 725 (1996) 93. S. Lacorte, D. Barcelo, J. Chromatogr. A 712 (1995) 104. M. Castillo, R. Domingues, M.F. Alpendurada, D. Barcelo, Anal. Chim. Acta 353 (1997) 134. V.A. Sakkas, D.A. Lambropoulou, T.M. Sakellarides, T.A. Albanis, Anal. Chim. Acta 467 (2002) 235. B. Albero, C. Sanchez-Brunete, J.L. Tadeo, J. Agric. Food Chem. 51 (2003) 6915. M. Fernandez, Y. Pico, S. Girotti, J. Manes, J. Agric. Food Chem. 49 (2001) 3540. N.H. Spliid, B. Koppen, J. Chromatogr. A 736 (1996) 106. J. Slobodnik, A.C. Hogenboom, U.A.Th. Brinkman, J. Chromatogr. A 741 (1996) 60. M. Guilhaus, V. Mlynski, D. Selby, Rapid Commun. Mass Spectrom. 11 (1997) 951. R.J. Cotter, Anal. Chem. 71 (1999) 446A. K.F. Blom, Anal. Chem. 73 (2001) 715. P. Eichhorn, D.S. Aga, Anal. Chem. 76 (2004) 6007. U.S. EPA, Summary of the Risks and Uses of Organophosphate Methyl Parathion, 2000, http://www.epa.gov/pesticides/op/methyl parathion/methylsum.htm. Y.F. Tang, Y.X. Wang, H.S. Cai, B.J. Merkel, J. Wuhan Inst. Chem. Technol. 26 (2004) 11. K.M. Kelly, A.Y. Chistoserdov, FEMS Microbiol. Ecol. 35 (2001) 86. L. Zheng, H.M. Chen, X.T. Han, W. Lin, X.J. Yan, World J. Microbiol. Biotechnol. 21 (2005) 202. J.M. Tiedje, Nat. Biotechnol. 20 (2002) 1093. J.B. Fein, D.A. Fowle, J. Cahill, K. Kemner, M. Boyanov, B. Bunker, Geomicrobiol. J. 19 (2002) 369. R. Bencheikh-Latmani, S.M. Williams, L. Haucke, C.S. Criddle, L.Y. Wu, J.Z. Zhou, B.M. Tebo, Appl. Environ. Microbiol. 71 (2005) 7453. R.K. Ellison, E. Malnati, A. Depaola, J. Bowers, G.E. Rodrick, J. Food Prot. 64 (2001) 682. S.V. Dzyadevych, A.P. Soldatkin, J. Chovelon, Anal. Chim. Acta 459 (2002) 35. S.V. Dzyadevych, J. Chovelon, Mater. Sci. Eng. C 21 (2002) 56. C.X. Liu (Ed.), Practice Pharmacokinetics, Chinese Medical Technology Publish, Beijing, 2003, p. 69. H. Soreq, S. Seidman, Nat. Rev. Neurosci. 2 (2001) 294. C.B. Millard, G. Koellner, A. Ordentlich, A. Shafferman, I. Silman, J.L. Sussman, J. Am. Chem. Soc. 121 (1999) 9883.

SYNCHROTRON RADIATION INFRARED SPECTROMICROSCOPY: A NONINVASIVE CHEMICAL PROBE FOR MONITORING BIOGEOCHEMICAL PROCESSES H.‐Y. N. Holman1,2 and M. C. Martin3 1

Ecology Department, Earth Sciences Division, Lawrence Berkeley National Laboratory, University of California, Berkeley, California 94720 2 Virtual Institute for Microbial Stress and Survival, Lawrence Berkeley National Laboratory, University of California, Berkeley, California 94720 3 Advanced Light Source Division, Lawrence Berkeley National Laboratory, University of California, Berkeley, California 94720

I. Introduction II. SR‐FTIR Spectromicroscopy A. Background B. Synchrotron IR Light Sources C. Synchrotron IR Spectromicroscopy of Biogeochemical Systems III. Biogeochemical Processes Measured by SR‐FTIR Spectromicroscopy A. Instrumentation B. Spectral Analysis C. Application Examples IV. Future Possibilities and Requirements Acknowledgments References

A long‐standing desire in biogeochemistry is to be able to examine the cycling of elements by microorganisms, as the processes are happening on surfaces of earth and environmental materials. Over the past decade, physics, engineering, and instrumentation innovations have led to the introduction of synchrotron radiation‐based infrared (IR) spectromicroscopy. Spatial resolutions of less than 10 micrometers (mm) and photon energies of less than an electron volt make synchrotron IR spectromicroscopy noninvasive and useful for following the course of biogeochemical processes on complex heterogeneous surfaces of earth and environmental materials. In this chapter, we will first briefly describe the technology and then present 79 Advances in Agronomy, Volume 90 Copyright 2006, Elsevier Inc. All rights reserved. 0065-2113/06 $35.00 DOI: 10.1016/S0065-2113(06)90003-0

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surfaces of earth and environmental materials. This surface biogeochemistry can be highly variable at a microscopic level because of the small‐scale (ranging from one micron to hundreds of microns) surface heterogeneity, which involves the distributions of clusters of mineral‐inhabiting microorganisms and reactive molecules of metal oxides and organic molecules. The methodology commonly employed to study this type of heterogeneous biogeochemical phenomenon is a combination of microscopic imaging and synchrotron radiation (SR)‐based X‐ray spectroscopy techniques. The interested readers can read reviews (Brown and Parks, 2001; Gordon and Sturchio, 2002) and other relevant studies (Amonette et al., 2003; Arnesano et al., 2003; Benison et al., 2004; Benzerara et al., 2005; Cooper et al., 2005; De Stasio et al., 2001; Fein et al., 2002; Foriel et al., 2004; Francis et al., 2004; Jones et al., 2003; Jurgensen et al., 2004; Khijniak et al., 2005; Lack et al., 2002; Li et al., 2003; Lieberman et al., 2003; Neal et al., 2004a,b; Nesterova et al., 2003; Panak et al., 2002; Pickering et al., 2001; Prange et al., 2002a,b; Renshaw et al., 2005; Saita and Maenosono, 2005; Sarret et al., 2005; Suzuki et al., 2003; Tebo et al., 2004, 2005; Templeton et al., 2005; Toner et al., 2005; Twining et al., 2004; Vogt et al., 2003; Watson and Ellwood, 2003; Wildung et al., 2004; Zouboulis and Katsoyiannis, 2005). SR‐based X‐ray spectromicroscopy studies have provided important and unique information about how microorganisms interact with earth and environmental materials. However, the energy range associated with SR‐based X‐ray spectromicroscopy techniques is between tens and thousands of electron volts (eV), which can adversely aVect, harm, or even kill the microorganisms. Consequently, it has limited the use of these techniques to measuring the biogeochemical actions only at single time points. Being able to measure real‐time sequential molecular changes in a biogeochemical system, as they are happening on surfaces of earth and environmental surfaces, has been a long‐standing scientific desire in biogeochemistry. The new availability of SR‐based infrared (IR) sources to the scientific community in the 1990s provided this opportunity. Our group began developing an SR‐based Fourier transform infrared (SR‐FTIR) spectromicroscopy approach in 1998 for studying biogeochemical transformation of environmental pollutants, choosing the reduction of hexavalent chromium by living microorganisms on mineral surfaces as the initial application (Holman et al., 1999). Prior to the availability of SR‐based IR facilities, these type of in vivo and in situ measurements were very diYcult for two reasons. First, earth materials inherently have low IR reflectivity surfaces. High‐quality IR spectroscopy measurements of earth and environmental materials require a high‐IR photon flux on small surface areas. Without an SR‐based source, one often needs to coadd thousands to tens of thousands of spectral scans, which can be prohibitively time consuming. Second, the IR measurements of live microorganisms had been problematic.

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Fein, J. B., Fowle, D. A., Cahill, J., Kemner, K., Boyanov, M., and Bunker, B. (2002). Nonmetabolic reduction of Cr(VI) by bacterial surfaces under nutrient‐absent conditions. Geomicrobiol. J. 19, 369–382. Foriel, J., Philippot, P., Susini, J., Dumas, P., Somogyi, A., Salome, M., Khodja, H., Menez, M., Fouquet, Y., Moreira, D., and Lopez‐Garcia, P. (2004). High‐resolution imaging of sulfur oxidation states, trace elements, and organic molecules distribution in individual microfossils and contemporary microbial filaments. Geochim. Cosmochim. Acta 68, 1561–1569. Francis, A. J., Dodge, C. J., Gillow, J. B., and Papenguth, H. W. (2000). Biotransformation of uranium compounds in high ionic strength brine by a halophilic bacterium under denitrifying conditions. Environ. Sci. Technol. 34, 2311–2317. Francis, A. J., Gillow, J. B., Dodge, C. J., Harris, R., Beveridge, T. J., and Papenguth, H. W. (2004). Uranium association with halophilic and non‐halophilic bacteria and archaea. Radiochim. Acta 92, 481–488. Fredrickson, J. K., Zachara, J. M., Balkwill, D. L., Kennedy, D., Li, S. M. W., Kostandarithes, H. M., Daly, M. J., Romine, M. F., and Brockman, F. J. (2004). Geomicrobiology of high‐ level nuclear waste‐contaminated vadose sediments at the Hanford site, Washington State. Appl. Environ. Microbiol. 70, 4230–4241. Furukawa, K. (2000). Biochemical and genetic bases of microbial degradation of polychlorinated biphenyls (PCBs). J. Gen. Appl. Microbiol. 46, 283–296. Furukawa, K. (2003). ‘‘Super bugs’’ for bioremediation. Trends Biotechnol. 21, 187–190. Furukawa, K., Hirose, J., Suyama, A., Zaiki, T., and Hayashida, S. (1993). Gene components responsible for discrete substrate‐specificity in the metabolism of biphenyl (Bph operon) and toluene (Tod operon). J. Bacteriol. 175, 5224–5232. Furukawa, K., Suenaga, H., and Goto, M. (2004). Biphenyl dioxygenases: Functional versatilities and directed evolution. J. Bacteriol. 186, 5189–5196. Gez, S., Luxenhofer, R., Levina, A., Codd, R., and Lay, P. A. (2005). Chromium(V) complexes of hydroxamic acids: Formation, structures, and reactivities. Inorg. Chem. 44, 2934–2943. Ghiorse, W. C., and Chapnick, S. D. (1983). Metal‐depositing bacteria and the distribution of manganese and iron in swamp waters. Ecol. Bull. 35, 367–376. Ghiorse, W. C., and Hirsch, P. (1979). Ultrastructural‐study of iron and manganese deposition associated with extracellular polymers of pedomicrobium‐like budding bacteria. Arch. Microbiol. 123, 213–226. Ghosh, U., Talley, J. W., and Luthy, R. G. (2001). Particle‐scale investigation of PAH desorption kinetics and thermodynamics from sediment. Environ. Sci. Technol. 35, 3468–3475. Goodhue, L. D., Hamilton, S., and Southam, G. (2005). The geomicrobiology of surficial geochemical anomalies. Geochim. Cosmochim. Acta 69, A367. Gordon, G. E., and Sturchio, N. C. (2002). An overview of synchrotron radiation applications to low temperature geochemistry and environmental science. In ‘‘Reviews in Mineralogy & Geochemistry’’ (P. A. Fenter, M. L. Rivers, N. C. Sturchio, and S. R. Sutton, Eds.), The Mineralogical Society of America, Vol. 69, pp. 1–116, Washington DC. Gore, R. C. (1949). Infrared spectrometry of small samples with the reflecting microscope. Science 110, 710–711. GriYth, W. P., Lewis, J., and Wilkinson, G. (1959). Infrared spectra of transition metal‐nitric oxide complexes.4. The pentacyanonitrosyl‐complexes of chromium and molybdenum. J. Am. Chem. Soc. (MAR) 872–875. Guilhaumou, N., Dumas, P., Carr, G. L., and Williams, G. P. (1998). Synchrotron infrared microspectrometry applied to petrography in micrometer‐scale range: Fluid chemical analysis and mapping. Appl. Spectrosc. 52(8), 1029–1034.

Geochimica et Cosmochimica Acta, Vol. 69, No. 3, pp. 553–577, 2005 Copyright © 2005 Elsevier Ltd Printed in the USA. All rights reserved 0016-7037/05 $30.00 ⫹ .00

doi:10.1016/j.gca.2004.07.018

Hydrous ferric oxide precipitation in the presence of nonmetabolizing bacteria: Constraints on the mechanism of a biotic effect DENIS G. RANCOURT,1,* PIERRE-JEAN THIBAULT,1,2 DENIS MAVROCORDATOS,3,† and GILLES LAMARCHE1 1

Lake Sediment Structure and Evolution (LSSE) Group, Department of Physics, University of Ottawa, Ottawa, ON K1N 6N5, Canada 2 Department of Earth Sciences, University of Ottawa, Ottawa, ON K1N 6N5, Canada ¨ berlandstrasse 133, CH-8600 Particle Laboratory, Swiss Federal Institute for Environmental Science and Technology, Postfach 611, U Dübendorf/Zurich, Switzerland

3

(Received February 20, 2004; accepted in revised form July 14, 2004)

Abstract—We have used room temperature and cryogenic 57Fe Mössbauer spectroscopy, powder X-ray diffraction (pXRD), mineral magnetometry, and transmission electron microscopy (TEM), to study the synthetic precipitation of hydrous ferric oxides (HFOs) prepared either in the absence (abiotic, a-HFO) or presence (biotic, b-HFO) of nonmetabolizing bacterial cells (Bacillus subtilis or Bacillus licheniformis, ⬃108 cells/mL) and under otherwise identical chemical conditions, starting from Fe(II) (10⫺2, 10⫺3, or 10⫺4 mol/L) under open oxic conditions and at different pH (6 –9). We have also performed the first Mössbauer spectroscopy measurements of bacterial cell wall (Bacillus subtilis) surface complexed Fe, where Fe(III) (10⫺3.5–10⫺4.5 mol/L) was added to a fixed concentration of cells (⬃108 cells/mL) under open oxic conditions and at various pH (2.5– 4.3). We find that non-metabolic bacterial cell wall surface complexation of Fe is not passive in that it affects Fe speciation in at least two ways: (1) it can reduce Fe(III) to sorbed-Fe2⫹ by a proposed steric and charge transfer effect and (2) it stabilizes Fe(II) as sorbed-Fe2⫹ against ambient oxidation. The cell wall sorption of Fe occurs in a manner that is not compatible with incorporation into the HFO structure (different coordination environment and stabilization of the ferrous state) and the cell wall-sorbed Fe is not chemically bonded to the HFO particle when they coexist (the sorbed Fe is not magnetically polarized by the HFO particle in its magnetically ordered state). This invalidates the concept that sorption is the first step in a heterogeneous nucleation of HFO onto bacterial cell walls. Both the a-HFOs and the b-HFOs are predominantly varieties of ferrihydrite (Fh), often containing admixtures of nanophase lepidocrocite (nLp), yet they show significant abiotic/biotic differences: Biotic Fh has less intraparticle (including surface region) atomic order (Mössbauer quadrupole splitting), smaller primary particle size (magnetometry blocking temperature), weaker Fe to particle bond strength (Mössbauer center shift), and no six-line Fh (6L-Fh) admixture (pXRD, magnetometry). Contrary to current belief, we find that 6L-Fh appears to be precipitated directly, under a-HFO conditions, from either Fe(II) or Fe(III), and depending on Fe concentration and pH, whereas the presence of bacteria disables all such 6L-Fh precipitation and produces two-line Fh (2L-Fh)-like biotic coprecipitates. Given the nature of the differences between a-HFO and b-HFO and their synthesis condition dependences, several biotic precipitation mechanisms (template effect, near-cell environment effect, catalyzed nucleation and/or growth effect, and substrate-based coprecipitation) are ruled out. The prevailing present view of a template or heterogeneous nucleation barrier reduction effect, in particular, is shown not to be the cause of the large observed biotic effects on the resulting HFOs. The only proposed mechanism (relevant to Fh) that is consistent with all our observations is coprecipitation with and possible surface poisoning by ancillary bacteriagenic compounds. That bacterial cell wall functional groups are redox active and the characteristics of biotic (i.e., natural) HFOs compared to those of abiotic (i.e., synthetic) HFOs have several possible biogeochemical implications regarding Fe cycling, in the photic zones of water columns in particular. Copyright © 2005 Elsevier Ltd al., 1999; Fowle and Fein, 1999; Fein et al., 2001; Yee and Fein, 2001; Kulczycki et al., 2002) rather than direct physical methods such as local coordination environment and valence state sensitive spectroscopies. This followed early direct observations of nonmetabolic metal sorption onto bacterial cell walls in sediments (Degens et al., 1970; Degens and Ittekkot, 1982) and early laboratory investigations of the phenomenon (e.g., Starkey, 1945; Pringsheim, 1949; Beveridge and Murray, 1980). The importance of passive (i.e., nonmetabolic) cell wall metal complexation probably cannot be overstated (Morel and Morel-Laurens, 1981). It is desirable therefore to study surface complexation of metals with molecular resolution probes (e.g., Cr, using XANES; Fein et al., 2002a), as we have done for the first time for Fe, using Mössbauer spectroscopy. This directly gives bonding environments, electron transfers, steric re-

1. INTRODUCTION

1.1. Bacterial Metal Sorption and Bacterium-HFO Associations Bacteria are ubiquitous in aquatic and humid environments, where they constitute a quantitatively important metal sorbing compartment (Beveridge and Doyle, 1989; Ledin, 2000; Warren and Haack, 2001) that is actively being studied, usually by wet chemical methods such as titration (e.g., Fein et al., 1997; Daughney and Fein, 1998; Daughney et al., 1998; Cox et

* Author to whom correspondence ([email protected]). † Deceased.

should

be

addressed

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559

Fig. 1. LNT Mössbauer spectra and their fits for two sorbed-Fe experiments: (a) spectrum SB1-2-LNT (Tables 1 and 2) showing a cell wall sorbed Fe3⫹ doublet contribution (dotted line) and a cell wall sorbed Fe2⫹ doublet contribution (solid line), and (b) spectrum SB3-9N-LNT (Tables 1 and 2) showing a 2L-Fh Fe3⫹ doublet (dotted line) and no Fe2⫹ contribution. The individual dots are the raw folded spectral data. The theoretical fit (not shown) consists in the sum of the assumed spectral contributions.

sorbed to HFOs as it would be magnetically polarized by the magnetic order of the HFO at 4.2 K, (2) cannot be sorbed in dense patches of Fe2⫹ that would have intercation magnetic superexchange bonds, as would be required if they were to form a template for oxide growth by re-oxidation, and (3) most probably do not form a separate Fe2⫹-rich solid phase because most such phases would be magnetically ordered at 4.2 K. Since TEM observations of similar samples (Warren and Ferris, 1998) show either only sorbed-Fe or sorbed-Fe with HFO nanogranules, depending on pH and total Fe concentration, we conclude that some of the originally sorbed Fe3⫹ must have been reduced to sorbed-Fe2⫹, presumably by either oxidizing the functional group or by transferring an electron from some other component of the bacterium. A 2⫹D contribution is always detected in spectra that display measurable amounts of sorbed-Fe3⫹ (Table 2) and the detected sorbed-Fe2⫹ amount always increases with increasing durations of wet aging (under bottle top atmosphere, at 4°C, and with some exposure to higher temperatures) (Tables 1 and 2). There appears to be a maximum amount of sorbed-Fe2⫹ of ⬃20% of total sorbed-Fe. This saturation amount occurs in at least three samples

(Tables 1 and 2): (1) 21 ⫾ 2% in sample SB1 after 4 d of wet aging; (2) 18 ⫾ 6% in sample SB6 after 1d of wet aging; and (3) 20 ⫾ 10% in sample SB5, counting only the sorbed fraction and excluding the 80% of total Fe that is HFO, after 1 d of wet aging. Based on the larger total Fe concentration used in a fourth sample (SB2, Table 1), we expect that 16.6% of the total Fe in this sample is sorbed-Fe, since 10⫺4.5 mol/L gives full sorption coverage at this pH (sample SB1 and Warren and Ferris, 1998). Therefore, the observed 5 ⫾ 2% of total Fe that is sorbed-Fe2⫹ in sample SB2 after 13 d of wet aging corresponds to 30 ⫾ 12% of sorbed-Fe that is sorbed-Fe2⫹. The observed production of sorbed-Fe2⫹ may have been enhanced by metabolic consumption of oxygen during wet aging, using dead cells and ancillary organics as electron donors. It appears to be a form of nonmetabolic reduction not unlike the reduction of Cr(VI) to Cr(III) observed by Fein et al. (2002a). The observed sorbed-Fe2⫹ is compatible with our a/b-HFO experiments described below (section 5.3), where Fe(II) was employed and was stabilized up to ⬃100% of available bacterial sorption sites, in the b-HFO preparations.

Biotic precipitation of hydrous ferric oxide ing the making of the SB samples and sharing the costs of the Mössbauer measurements. We thank Dr. Nagina Parma (University of Toronto) for performing all the surface complexation (SB) syntheses. We thank Prof. Danielle Fortin for her help and guidance in initiating the a-b-HFO part of this project and for access to her laboratory for performing the a-b-HFO syntheses. We thank Gabriella Giustiniano, Milena Kushnir, and Natalie Roux for helping to develop the project with preliminary syntheses and characterizations. We thank Marie Wang for extensive literature work. DGR thanks Prof. Jeremy B. Fein for several instructive discussion sessions in the early stages of this project. We thank Associate Editor Fein and two anonymous reviewers for extensive comments. We gratefully acknowledge financial support from the Natural Sciences and Engineering Research Council of Canada. Associate editor: J. B. Fein

REFERENCES Appenzeller B. M. R., Duval Y. B., Thomas F., and Block J.-C. (2002) Influence of phosphate on bacterial adhesion onto iron oxyhydroxide in drinking water. Environ. Sci. Technol. 36, 646 – 652. Banfield J. F., Welch S. A., Zhang H., Thomsen Ebert T., and Penn R. L. (2000) Aggregation-based crystal growth and microstructure development in natural iron oxyhydroxide biomineralization products. Science 289, 751–754. Barbeau K. and Moffett J. W. (2000) Laboratory and field studies of colloidal iron oxide dissolution as mediated by phagotrophy and photolysis. Limnol. Oceanogr. 45, 827–235. Barbeau K., Rue E. L., Bruland K. W., and Butler A. (2001) Photochemical cycling of iron in the surface ocean mediated by microbial iron(III)-binding ligands. Nature 413, 409 – 413. Barker W. W., Welch S. A., Chu S., and Banfield J. F. (1998) Experimental observations of the effects of bacteria on aluminosilicate weathering. Am. Mineral. 83, 1551–1563. Bauminger E. R., Harrison P. M., Hechel D., Nowik I., and Treffry A. (1991) Iron (III) can be transferred between ferritin molecules. Proc. R. Soc. London B 244, 211–217. Bergh Ø., Børsheim K. Y., Bratbak G., and Heldal M. (1989) High abundance of viruses found in aquatic environments. Nature 340, 467– 468. Beveridge T. J. and Murray R. G. E. (1980) Sites of metal deposition in the cell wall of Bacillus subtilis. J. Bacteriol. 141, 876 – 887. Beveridge T. J. and Doyle R. J. (1989) Metal Ions and Bacteria. John Wiley, New York. Beveridge T. J., Meloche J. D., Fyfe W. S., and Murray R. G. E. (1983) Diagenesis of metals chemically complexed to bacteria: Laboratory formation of metal phosphates, sulfides and organic condensates in artificial sediments. Appl. Environ. Microbiol. 45, 1094 –1108. Brown D. A., Choudori Kamieni D., Sawicki J. A., and Beveridge T. J. (1994) Minerals associated with biofilms occurring on exposed rock in granitic underground research laboratory. Appl. Environ. Microbiol. 60, 3182–3191. Châtellier X., Fortin D., West M. M., Leppard G. G., and Ferris F. G. (2001) Effect of the presence of bacterial surfaces during the synthesis of Fe oxides by oxidation of ferrous ions. Eur. J. Mineral. 13, 705–714. Châtellier X., West M. M., Rose J., Fortin D., Leppard G. G., and Ferris F. G. (2004) Characterization of iron-oxides formed by oxidation of ferrous ions in the presence of various bacterial species and inorganic ligands. Geomicrobiol. J. 21, 99 –112. Coleman A. W. (1980) Enhanced detection of bacteria in natural environments by fluorochrome staining of DNA. Limnol. Oceanogr. 25, 948 –951. Cornell R. M. and Schwertmann U. (1979) Influence of organic anions on the crystallization of ferrihydrite. Clays Clay Miner. 27, 402– 410. Cornell R. M. and Schwertmann U. (1996) The Iron Oxides—Structure, Properties, Reactions, Occurrence and Uses. VCH, Weinheim, Germany.

575

Cowen J. P. and Bruland K. W. (1985) Metal deposits associated with bacteria: Implications for Fe and Mn marine biogeochemistry. Deep-Sea Res. 32, 253–272. Cox J. S., Smith D. S., Warren L. A., and Ferris F. G. (1999) Characterizing heterogeneous bacterial surface functional groups using discrete affinity spectra for proton binding. Environ. Sci. Technol. 33, 4514 – 4521. Daughney C. J. and Fein J. B. (1998) The effect of ionic strength on the adsorption of H⫹, Cd2⫹, Pb2⫹ and Cu2⫹ by Bacillus subtilis and Bacillus licheniformis: A surface complexation model. J. Colloid Interface Sci. 198, 53–77. Daughney C. J., Fein J. B., and Yee N. (1998) A comparison of the thermodynamics of metal adsorption onto two common bacteria. Chem. Geol. 144, 161–176. Decho A. W. (1990) Microbial exopolymer secretions in ocean environments: Their role(s) in food webs and marine processes. In Oceanography and Marine Biology (ed. M. Barnes), pp. 73–153. Aberdeen University Press, Aberdeen, UK. Degens E. T. and Ittekkot V. (1982) In situ metal-staining of biological membranes in sediments. Nature 298, 262–264. Degens E. T., Watson S. W., and Remsen C. C. (1970) Fossil membranes and cell wall fragments from a 7000-year-old Black Sea sediment. Science 168, 1207–1208. Domes H., Leupold O., Nagy D. L., Ritter G., Spiering H., Molnár B., and Szücs I. S. (1986) Mossbauer study of short range order in frozen aqueous solutions of Fe(CIO4)2. J. Chem. Phys. 85, 7294 – 7300. Doyle R. J. (1989) How cell walls of gram-positive bacteria interact with metal ions. In Metal Ions and Bacteria (eds. T. J. Beveridge and R. J. Doyle), pp. 275–293. John Wiley, New York. Droppo I. G. (2001) Rethinking what constitutes suspended sediment. Hydrol. Proc. 15, 1551–1564. Ehrlich H. L. (1999) Microbes as geologic agents: Their role in mineral formation. Geomicrobiol. J. 16, 135–153. Fein J. B., Daughney C. J., Yee N., and Davis T. A. (1997) A chemical equilibrium model for metal adsorption onto bacterial surfaces. Geochim. Cosmochim. Acta 61, 3319 –3328. Fein J. B., Martin A. M., and Wightman P. G. (2001) Metal adsorption onto bacterial surfaces: Development of a predictive approach. Geochim. Cosmochim. Acta 65, 4267– 4273. Fein J. B., Fowle D. A., Cahill J., Kemner K., Boyanov M., and Bunker B. (2002a) Nonmetabolic reduction of Cr(VI) by bacterial surfaces under nutrient-absent conditions. Geomicrobiol. J. 19, 369 –382. Fein J. B., Scott S., and Rivera N. (2002b) The effect of Fe on Si adsorption by Bacillus subtilis cell walls: Insights into non-metabolic bacterial precipitation of silicate minerals. Chem. Geol. 182, 265–273. Ferris F. G., Beveridge T. J., and Fyfe W. S. (1986) Iron-silica crystallite nucleation by bacteria in a geothermal sediment. Nature 320, 609 – 611. Ferris F. G., Fyfe W. S., and Beveridge T. J. (1987) Bacteria as nucleation sites for authigenic minerals in a metal-contaminated lake sediment. Chem. Geol. 63, 225–232. Ferris F. G., Konhauser K. O., Lyvén B., and Pedersen K. (1999) Accumulation of metals by bacteriogenic iron oxides in a subterranean environment. Geomicrobiol. J. 16, 181–192. Ferris F. G., Schultze S., Witten T. C., Fyfe W. S., and Beveridge T. J. (1989a) Metal interactions with microbial biofilms in acidic and neutral pH environments. Appl. Environ. Microbiol. 55, 1249 – 1257. Ferris F. G., Shotyk W., and Fyfe W. S. (1989b) Mineral formation and decomposition by microorganisms. In Metal Ions and Bacteria (eds. T. J. Beveridge and R. J. Doyle), pp. 412– 441. John Wiley, New York. Fleisch H. (1975) Physico chemical and biological effects of pyrophosphate and diphosphonates. Colloques internationaux du Centre national de la recherche scientifique. Physico-chimie et cristallographie des apatites d’intérêt biologique. 230, 263–273. Paris, Éditions du Centre national de la recherche scientifique. Fortin D. and Beveridge T. J. (1997) Role of the bacterium Thiobacillus in the formation of silicates in acidic mine tailings. Chem. Geol. 141, 235–250.

APPLIED AND ENVIRONMENTAL MICROBIOLOGY, Dec. 2005, p. 7679–7689 0099-2240/05/$08.00⫹0 doi:10.1128/AEM.71.12.7679–7689.2005 Copyright © 2005, American Society for Microbiology. All Rights Reserved.

Vol. 71, No. 12

Soil Microbial Community Responses to Additions of Organic Carbon Substrates and Heavy Metals (Pb and Cr) Cindy H. Nakatsu,1* Nadia Carmosini,1,2 Brett Baldwin,1 Federico Beasley,1 Peter Kourtev,2 and Allan Konopka2 Department of Agronomy1 and Department of Biological Sciences,2 Purdue University, West Lafayette, Indiana 47907

Microcosm experiments were conducted with soils contaminated with heavy metals (Pb and Cr) and aromatic hydrocarbons to determine the effects of each upon microbial community structure and function. Organic substrates were added as a driving force for change in the microbial community. Glucose represented an energy source used by a broad variety of bacteria, whereas fewer soil species were expected to use xylene. The metal amendments were chosen to inhibit the acute rate of organic mineralization by either 50% or 90%, and lower mineralization rates persisted over the entire 31-day incubation period. Significant biomass increases were abolished when metals were added in addition to organic carbon. The addition of organic carbon alone had the most significant impact on community composition and led to the proliferation of a few dominant phylotypes, as detected by PCR-denaturing gradient gel electrophoresis of bacterial 16S rRNA genes. However, the community-wide effects of heavy metal addition differed between the two carbon sources. For glucose, either Pb or Cr produced large changes and replacement with new phylotypes. In contrast, many phylotypes selected by xylene treatment were retained when either metal was added. Members of the Actinomycetales were very prevalent in microcosms with xylene and Cr(VI); gene copy numbers of biphenyl dioxygenase and phenol hydroxylase (but not other oxygenases) were elevated in these microcosms, as determined by real-time PCR. Much lower metal concentrations were needed to inhibit the catabolism of xylene than of glucose. Cr(VI) appeared to be reduced during the 31-day incubations, but in the case of glucose there was substantial microbial activity when much of the Cr(VI) remained. In the case of xylene, this was less clear. unclear since no studies have addressed this point. If species richness is reduced in sites contaminated with complex mixtures, the communities may be less resilient because the probability that an ecotype is capable of a specific required function is reduced (13). Previous research in our laboratory showed that the microbial community structure in a long-term mixedwaste contaminated site might reflect both metal and aromatic hydrocarbon concentrations in the soil. For individual microbes to persist under complex conditions, they must tolerate both local metal and hydrocarbon contaminants. Shi et al. (42) found a very broad distribution of metal tolerances within the microbial communities in these soils. This is consistent with a heterogeneous distribution of microbes in both highly contaminated and noncontaminated microsites. For this study, we used microcosms to segregate the effects of two metals, Pb and Cr, and also to study the impact of aromatic hydrocarbons on microbial community structure and activity. By using microcosms, soils could be homogenized to evenly distribute both the microbial populations and toxicants and thereby reduce spatial variability. This allowed us to test experimental additions of Pb2⫹ or Cr6⫹ and to assess their effects on the microbial community. One of two energy substrates (glucose or xylene) was added to provide the necessary force for selection to operate and drive changes in community composition. Glucose is broadly utilized by microorganisms; xylene catabolism is more restricted among microbes, and xylene mimics aromatic compounds present in these soils. The changes in community activity were related to molecular analyses of community composition and functional gene levels.

The use and release of heavy metals to air, water, and soils has created a significant number of contaminated sites across the United States and the world. Thus, the effect of metal contamination on the microbial community has been extensively studied over the past several decades. The acute effects of short-term exposure to toxic heavy metals upon a broad array of microbial processes have been well documented (9, 10, 21, 40, 47). More recently, investigators have examined habitats exposed to anthropogenic or natural metal contamination over an extended period of time (5, 20, 22, 24, 39, 45). Studies that focused on the culturable fraction of the microbial community indicated that as few as 10 to almost 100% of the bacteria in habitats contaminated for extended periods were metal resistant. Thus, there may be substantial variability in community responses to metal exposure between locations. As non-cultivation-based methods have become available, researchers have begun examining the impact of metal exposure on the entire indigenous community (6, 24, 25, 26) and have tried to address the impact of these exposures on community diversity (30) and resiliency (15). A confounding factor in metal-contaminated sites is the frequent co-occurrence of organic contaminants. These organic molecules may be metabolizable energy sources, toxicants, or both. The combined effect of metals and organic carbon pollutants on microbial activity and community composition is

* Corresponding author. Mailing address: Department of Agronomy, Purdue University, West Lafayette, IN 47907-2054. Phone: (765) 496-2997. Fax: (765) 496-2629. E-mail: [email protected]. 7679

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Received 26 November 2004/Accepted 24 July 2005

VOL. 71, 2005

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centrations than those required when glucose was the added energy source. As noted above, this difference may reflect the narrower pool of metal-tolerant species that could catabolize xylene than of those that could catabolize glucose. A final physiological consequence of metal addition was the reduction in net biomass increases at the expense of added organic carbon. As noted above, growth dynamics were still present to produce changes in DGGE fingerprints, culturable bacterial numbers, and catabolic gene copies. However, the lack of biomass increase may also reflect an energy expenditure to implement metal tolerance, as is true in the case of ATPdependent efflux systems (43). The ratio of mineralized to assimilated organic C was found to increase in other metalcontaminated soils (6), which is consistent with the increased energy expenditure under these conditions. The BPH4-type biphenyl dioxygenase and PHE genes were routinely detected in control microcosms and most xyleneamended microcosms, rather than the xylene monooxygenase and toluate or benzoate dioxygenases usually associated with xylene metabolism (The University of Minnesota Biocatalysis/ Biodegradation Database [http://umbbd.ahc.umn.edu/]). The presence of aromatic oxygenase genes in the control microcosms was not unexpected because the soil sample was taken from a site contaminated with aromatic hydrocarbons. BPH4 and PHE gene copies increased in xylene microcosms and corresponded to changes in CO2 evolution, and in some cases, biomass; this suggests that there was an enrichment of strains harboring these oxygenase genes. Phenol hydroxylase catalyzes the continued oxidation of hydroxylated intermediates in xylene and toluene catabolism (3) and has been detected in other xylene-amended-microcosm studies (B. Baldwin, personal communication). The selection of BPH4 oxygenase genes following the addition of xylenes may seem counterintuitive; however, previous reports have noted the sequence similarity and functional overlap of biphenyl and alkyl-benzene dioxygenases, including toluene dioxygenase (17, 44). The BPH4 subfamily of biphenyl dioxygenase genes, in particular, is closely related to isopropylbenzene dioxygenase genes. Although Cr(VI) inhibits microbes, there are biotic and abiotic detoxification mechanisms in soil (4). One biotic mechanism (under both aerobic and anaerobic conditions) occurs via Cr(VI) reduction to less toxic and less mobile Cr(III) (14, 48). Cr(VI) reduction to nontoxic levels could have been a precondition for the onset of microbial activity in the microcosms. However, this was not the case in systems to which glucose was added; these could tolerate relatively large additions of Cr(VI), and microbes were mineralizing glucose when most of the Cr(VI) remained. In the case of xylene amendments, the relationship is less clear. Because only low levels of Cr(VI) could be added to retain any xylene degradation at all, even a modest rate of biological reduction resulted in very small residual Cr(VI) concentrations. Coupled with the lag in detectable xylene mineralization even in the absence of Cr(VI), these data do not convincingly show that xylene catabolism occurred while Cr(VI) was present. More sensitive analyses (for example, with radiotracers) would be required to resolve this issue. Rarefaction analysis of a 16S rRNA gene clone library created from several soils at this site indicated that they contain a relatively low diversity of microbes (21a), and the provision of a single carbon source in microcosms produced a further re-

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more sweeping replacements of community members in response to metal additions than did xylene-amended ones. The number of different organic substrates and metals tested was small; however, this result may reflect the distinction between cases where there are large numbers of organisms that might be recruited (glucose catabolizers) and those (xylene) where potential responders (and their levels of metal resistance) are more limited. The Cr(VI) levels were different for each substrate and were titrated to reduce catabolism by a predetermined value; the amount added to glucose microcosms to reduce activity in the glucose microcosms by 90% [4 mg g⫺1of Cr(VI)] resulted in no metabolic activity in microcosms with xylene. This experimental approach allowed us to identify some of the indigenous soil populations that responded to these conditions. Intense bands generally only occurred in 16S rRNA PCR-DGGE profiles for microcosms in which carbon mineralization occurred. We believe these intense bands correspond to organisms that have multiplied at the expense of the added organic substrate and are present as a significant fraction of the total population. This argument is strengthened by the significantly higher copy numbers of the catabolic BPH4 and PHE genes in the microcosms with xylene addition. Furthermore, xylene-degrading bacteria that were isolated from the microcosms contained BPH4 or PHE genes (Beasley et al., in preparation) and corresponded to intense bands in the community PCR-DGGE profiles. Other approaches, such as heavy isotope additions (29, 35), have been suggested as an approach to link populations responsible for functions in soil communities, but the approach used here is much simpler and less expensive and can be used on a broader scale in the field. The combination of PCR-DGGE of 16S rRNA genes with quantitative PCR of functional genes does have limitations because an individual band (organism) is not unequivocally linked to a specific function and because not all gene variants for specific catabolic functions are known. However, at the field scale, the correlation between stimulation of specific rRNA gene sequences and specific functional genes does provide a good initial basis to stimulate detailed analyses of the consequences of changes in community structure upon community function. The phylogenetic identity of many of the bacteria selected under these conditions is consistent with in situ analyses of these (21a) and other metal-contaminated soils. The Actinomycetales have been reported to be important in metal-impacted soils (18). In cases where glucose only was added or xylene plus Cr(VI) was added, organisms from several genera of the Actinomycetales commonly responded to the stimuli. Although the primary objective of this study was to analyze changes in community composition, activities were monitored to provide some basis for understanding community dynamics. In general, activity responses were similar to what has been observed in a number of other microcosm studies (8, 19, 37, 38). In the absence of metals, carbon additions stimulated increases in carbon mineralization, microbial biomass, and the number of culturable bacteria. Heavy metals generally suppressed these community responses, as lower carbon mineralization rates and longer lag phases were observed. On a molar basis, Cr was more toxic than Pb; this may reflect its greater mobility and bioavailability in soil (34). In addition, heavy metals inhibited the xylene-degrading community at lower con-

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ACKNOWLEDGMENTS This work was supported by a grant from the DOE Natural and Accelerated Bioremediation Research (NABIR) program (grant DEFG02-98ER62681). We thank the Indiana Department of Transportation, and in particular Bill Jervis, for giving us access to the site, Judy Lindell for technical assistance, and Leon Toussaint and Joanne Becker for assisting in soil collection. REFERENCES 1. Acosta-Martı´nez, V., Z. Reicher, M. Bischoff, and R. F. Turco. 1999. The role of tree leaf mulch and nitrogen fertilizer on turfgrass soil quality. Biol. Fert. Soils 29:55–61. 2. Altschul, S. F., T. L. Madden, A. A. Schaffer, J. Zhang, Z. Zhang, W. Miller, and D. J. Lipman. 1997. Gapped BLAST and PSI-BLAST: a new generation of protein database search programs. Nucleic Acids Res. 25:3389–3402. 3. Arenghi, F. L. G., D. Berlanda, E. Galli, G. Sello, and P. Barbieri. 2001. Organization and regulation of meta cleavage pathway genes for toluene and o-xylene derivative degradation in Pseudomonas stutzeri OX1. Appl. Environ. Microbiol. 67:3304–3308. 4. Avudainayagam, S., A. Megharaj, G. Owens, R. S. Kookana, D. Chittleborough, and R. Naid. 2003. Chemistry of chromium in soils with emphasis on tannery waste sites. Rev. Environ. Contam. Toxicol. 178:53–91. 5. Baath, E. 1992. Measurement of heavy metal tolerance of soil bacteria using thymidine incorporation into bacteria extracted after homogenization-centrifugation. Soil Biol. Biochem. 24:1167–1172. 6. Baath, E., M. Diaz-Ravina, A. Frostegard, and C. D. Campbell. 1998. Effect of metal-rich sludge amendments on the soil microbial community. Appl. Environ. Microbiol. 64:238–245. 7. Baldwin, B. R., C. H. Nakatsu, and L. Nies. 2003. Detection and enumeration of aromatic oxygenase genes by multiplex and real-time PCR. Appl. Environ. Microbiol. 69:3350–3358. 8. Bardgett, R. D., and S. Saggar. 1994. Effects of heavy-metal contamination on the short-term decomposition of labeled [C-14] glucose in a pasture soil. Soil Biol. Biochem. 26:727–733. 9. Barnhart, C. L., and R. Vestal. 1983. Effect of environmental toxicant on metabolic activity of natural microbial communities. Appl. Environ. Microbiol. 46:970–977. 10. Capone, D. G., D. Reese, and R. P. Kiene. 1983. Effects of metals on methanogenesis, sulfate reduction, carbon dioxide evolution, and microbial biomass in anoxic salt marsh sediments. Appl. Environ. Microbiol. 45:1586–1591. 11. Cassel, D. K., and D. R. Nielsen. 1986. Field capacity and available water, p. 901–926. In A. Klute (ed.), Methods of soil analysis. I. Physical and mineralogical methods, 2nd ed. American Society for Agronomy, Soil Science Society of America, Madison, Wis. 12. Clesceri, L. S., A. E. Greenberg, and A. D. Eaton. 1999. Standard methods for examination of water and wastewater, 20th ed. American Public Health Association, Washington, D.C. 13. Ekschmitt, K., and B. S. Griffiths. 1998. Soil biodiversity and its implications for ecosystem functioning in a heterogeneous and variable environment. Appl. Soil Ecol. 10:201–215. 14. Fein, J. B., D. A. Fowle, J. Cahill, K. Kemner, M. Boyanov, and B. Bunker. 2002. Nonmetabolic reduction of Cr(VI) by bacterial surfaces under nutrient-absent conditions. Geomicrobiol. J. 19:369–382. 15. Feris, K. P., P. W. Ramsey, M. Rillig, J. N. Moore, J. E. Gannon, and W. E. Holbert. 2004. Determining rates of change and evaluating group-level re-

siliency differences in hyporheic microbial communities in response to fluvial heavy-metal deposition. Appl. Environ. Microbiol. 70:4756–4765. 16. Findlay, R. H. 1996. The use of phospholipid fatty acids to determine microbial community structure, p. 1–17. In A. K. Akkermans, J. D. van Elsas, and F. de Bruijn (ed.), Molecular microbial ecology manual. Kluwer Academic Publishers, Dordrecht, The Netherlands. 17. Furukawa, K., J. Hirose, A. Suyama, T. Zaiki, and S. Hayashida. 1993. Gene components responsible for discrete substrate specificity in the metabolism of biphenyl (bph operon) and toluene (tod operon). J. Bacteriol. 175:5224– 5232. 18. Gremion, F., A. Chatzinotas, and H. Harms. 2003. Comparative 16S rDNA and 16S rRNA sequence analysis indicates that Actinobacteria might be a dominant part of the metabolically active bacteria in heavy metal-contaminated bulk and rhizosphere soil. Environ. Microbiol. 5:896–907. 19. Gremion, F., A. Chatzinotas, K. Kaufmann, W. V. Sigler, and H. Harms. 2004. Impacts of heavy metal contamination and phytoremediation on a microbial community during a twelve-month microcosm experiment. FEMS Microbiol. Ecol. 48:273–283. 20. Hutchinson, T. C., and M. S. Symington. 1997. Persistence of metal stress in a forested ecosystem near Sudbury, 66 years after closure of the O’Donnell roast bed. J. Geochem. Explor. 58:323–330. 21. Jonas, R. B., C. G. Gilmour, D. L. Stoner, M. M. Weir, and J. H. Tuttle. 1984. Comparison of methods to measure acute metal and organometal toxicity to natural aquatic microbial communities. Appl. Environ. Microbiol. 47:1005– 1011. 21a.Joynt, J., M. Bischoff, R. Turco, A. Konopka, and C. H. Nakatsu. Microbial community analysis of soils contaminated with lead, chromium and petroleum hydrocarbons. Microb. Ecol., in press. 22. Kamaludeen, S. P. B., M. Megharaj, R. Naidu, I. Singleton, A. L. Juhasz, B. G. Hawke, and N. Sethunathan. 2003. Microbial activity and phospholipid fatty acid pattern in long-term tannery waste-contaminated soil. Ecotox. Environ. Safety 56:302–310. 23. Konopka, A., D. Knight, and R. F. Turco. 1989. Characterization of a Pseudomonas sp. capable of aniline degradation in the presence of secondary carbon sources. Appl. Environ. Microbiol. 55:385–389. 24. Konopka, A., T. Zakharova, M. Bischoff, L. Oliver, C. Nakatsu, and R. F. Turco. 1999. Microbial biomass and activity in lead-contaminated soil. Appl. Environ. Microbiol. 65:2256–2259. 25. Konstantinidis, K. T., N. Isaacs, J. Fett, S. Simpson, D. T. Long, and T. L. Marsh. 2003. Microbial diversity and resistance to copper in metal-contaminated lake sediment. Microb. Ecol. 45:191–202. 26. Kozdroj, J., and J. D. van Elsas. 2001. Structural diversity of microorganisms in chemically perturbed soil assessed by molecular and cytochemical approaches. J. Microbiol. Methods 43:197–212. 27. Lane, D. J. 1991. 16S/23S rRNA sequencing, p. 115–175. In E. Stackebrandt and M. Goodfellow (ed.), Nucleic acid techniques in bacterial systematics. John Wiley & Sons, New York, N.Y. 28. LaPara, T. M., C. H. Nakatsu, L. Pantea, and J. E. Alleman. 2000. Phylogenetic analysis of bacterial communities in mesophilic and thermophilic bioreactors treating pharmaceutical wastewater. Appl. Environ. Microbiol. 66:3951–3959. 29. Manefield, M., A. S. Whiteley, R. I. Griffiths, and M. J. Bailey. 2002. RNA stable isotope probing, a novel means of linking microbial community function to phylogeny. Appl. Environ. Microbiol. 68:5367–5373. 30. Moffett, B. F., F. A. Nicholson, N. C. Uwakwe, B. J. Chambers, J. A. Harris, and T. C. J. Hill. 2003. Zinc contamination decreases the bacterial diversity of agricultural soil. FEMS Microbiol. Ecol. 43:13–19. 31. Muyzer, G., E. C. de Waal, and A. G. Uitterlinden. 1993. Profiling of complex microbial populations by denaturing gradient gel electrophoresis analysis of polymerase chain reaction-amplified genes coding for 16S rRNA. Appl. Environ. Microbiol. 59:695–700. 32. Muyzer, G., S. A. Hottentrager, A. Teske, and C. Wawer. 1996. Denaturing gradient gel electrophoresis of PCR-amplified 16S rDNA—a new molecular approach to analyse the genetic diversity of mixed microbial communities, p. 1–23. In A. Akkermans, J. D. van Elsas, and F. J. de Bruijn (ed.), Molecular microbial ecology manual 3.4.4. Kluwer Academic Publishers, Nowell, Mass. 33. Nakatsu, C. H., V. Torsvik, and L. Øvreås. 2000. Soil community analysis using DGGE of 16S rDNA polymerase chain reaction products. Soil Sci. Soc. Am. J. 64:1382–1388. 34. Nieboer, E., and A. A. Jusys. 1988. Biologic chemistry of chromium, p. 21–79. In J. O. Nriagu and E. Nieboer (ed.), Chromium in natural and human environments. Wiley, New York, N.Y. 35. Padmanabhan, P., S. Padmanabhan, C. DeRito, A. Gray, D. Gannon, J. R. Snape, C. S. Tsai, W. Park, C. Jeon, and E. L. Madsen. 2003. Respiration of 13 C-labeled substrates added to soil in the field and subsequent 16S rRNA gene analysis of 13C-labeled soil DNA. Appl. Environ. Microbiol. 69:1614– 1622. 36. Pielou, E. C. 1969. An introduction to mathematical ecology. Wiley-Interscience, New York, N.Y. 37. Rajapaksha, R. M. C. P., M. A. Tobor-Kaplon, and E. Baath. 2004. Metal toxicity affects fungal and bacterial activities in soil differently. Appl. Environ. Microbiol. 70:2966–2973.

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duction in diversity. These effects are explicable in light of the resource heterogeneity hypothesis (46), i.e., the site is relatively uniformly barren with respect to resource availability due to the lack of plant vegetation. As a result, heterotrophic productivity and diversity are low. The addition of a single C source to microcosms results in even lower organic resource heterogeneity, and there is strong selection for a few types that use that dominant C source. When an additional selection (heavy metals) factor was imposed, the community dynamics were found to be much greater for components expected to contain a large degree of functional redundancy (glucose-catabolizing bacteria) than a more restricted catabolic function (xylene degradation). Consistent with this broader capacity was the greater robustness of glucose catabolism, that is, it recovered and proceeded at much higher metal concentrations than did xylene catabolism.

APPL. ENVIRON. MICROBIOL.

THE ADSORPTION OF CHEMICAL CONTAMINANTS ONTO ENVIRONMENTAL SURFACES WITH SPECIAL CONSIDERATION OF THE BACTERIAL SURFACE

A Dissertation

Submitted to the Graduate School of the University of Notre Dame in Partial Fulfillment of the Requirements for the Degree of

Doctor of Philosophy

by

Drew Gorman-Lewis, Ph.D.

_________________________________ Jeremy B. Fein, Director

Graduate Program in Civil Engineering and Geological Sciences Notre Dame, Indiana November 2005

© Copyright by Drew Gorman-Lewis 2005 All rights reserved

THE ADSORPTION OF CHEMICAL CONTAMINANTS ONTO GEOLOGIC SURFACES WITH SPECIAL CONSIDERATION OF THE BACTERIAL SURFACE

Abstract by Drew Gorman-Lewis

Adsorption reactions are one of many processes to consider when attempting to predict and understand the movement of contaminants through the subsurface. This dissertation presents the work of four individual but related studies that measured and quantified adsorption reactions of chemical contaminants onto a variety of particulate subsurface media with special consideration of the reactivity of the bacterial surface. Chapter 2 describes the adsorption of an ionic liquid onto mineral oxides, clay, and bacteria. The experimental results reveal that 1-Butyl, 3-methylimidazolium chloride (Bmim Cl) is unstable in water below pH 6 and above pH 10 and that it exhibits pH independent and ionic strength dependent adsorption onto Na-montmorillonite with 0.4, 0.8, 1.0, 1.2, and 2.0 g/L of clay.

We observed no adsorption of the Bmim Cl onto Bacillus subtilis (3.95

or 7.91 g (dry weight) bacteria/L) at pH 5.5 to 8.5 or onto gibbsite (500 or 1285 g/L) or

Drew Gorman-Lewis

quartz (1000 and 2000 g/L) over the pH range 6-10. The measured adsorption was subsequently quantified using a distribution coefficient approach.

Chapter 3 focuses

specifically on the reactivity of the bacterial surface using the new technique of combining titration calorimetry with surface complexation modeling to produce sitespecific enthalpies and entropies of proton and Cd adsorption.

Our results provide

mechanistic details of these adsorption reactions that are impossible to gain from previous techniques used to study the bacterial surface. Chapters 4 and 5 present work measuring and quantifying the adsorption of U and Np onto B. subtilis under a variety of conditions. Np adsorption exhibited a strong ionic strength dependence and unusual behavior under low pH high ionic strength conditions that was consistent with reduction of Np(V) to Np(IV). U adsorption, in constrast to Np adsorption, was extensive under all conditions studied. Thermodynamic modeling of the data suggests that uranyl-hydroxide, uranyl-carbonate and calcium-uranyl-carbonate species each can form stable surface complexes on the bacterial cell wall.

These studies investigate a variety of adsorption

reactions and provide parameters to quantify adsorption that may aid in integration of these reactions into geochemical models to predict contaminant transport in the subsurface.

CONTENTS

FIGURES............................................................................................................ v TABLES .......................................................................................................... vii ACKNOWLEDGMENTS ............................................................................... viii CHAPTER 1: INTRODUCTION ........................................................................ 1 CHAPTER 2: EXPERIMENTAL STUDY OF THE ADSORPTION OF AN IONIC LIQUID ONTO BACTERIAL AND MINERAL SURFACES ............................ 7 2.1 Introduction................................................................................................... 7 2.2 Experimental Procedures .............................................................................. 9 2.3 Results and Discussion ............................................................................... 14 2.4 Conclusions ................................................................................................ 27 CHAPTER 3: ENTHALPIES AND ENTROPIES OF PROTON AND CADMIUM ADSORPTION ONTO BACILLUS SUBTILIS FROM CALORIMETRIC MEASUREMENTS ......................................................................................... 28 3.1 Introduction ................................................................................................ 28 3.2 Experimental Procedures ............................................................................ 31

ii

3.2.1 Cell Prepartion ......................................................................................... 31 3.2.2 Bulk Adsorption Experiments .................................................................. 32 3.2.3 Titrations Calorimetry .............................................................................. 33 3.3 Results and Discussion ............................................................................... 38 3.3.1 Protonation Reactions .............................................................................. 38 3.3.2 Cd Adsorption Reactions .......................................................................... 49 3.4 Implications of Derived Enthalpies and Entropies........................................ 52 3.5 Conclusions ................................................................................................ 57 CHAPTER 4: EXPERIMENTAL STUDY OF NEPTUNYL ADSORPTION ONTO BACILLIS SUBTILIS ........................................................................................ 58 4.1 Introduction ................................................................................................ 58 4.2 Methods and Materials ................................................................................ 60 4.2.1 Cell Preparation ........................................................................................ 60 4.2.2 pH and Ionic Strength Dependent Adsorption Experiments....................... 61 4.2.3 Concentration Dependent Adsorption Experiments ................................... 62 4.2.4 Desorption Experiments............................................................................ 63 4.2.5 Kinetics Experiments ............................................................................... 63 4.3 Results ........................................................................................................ 60 4.3.1 Kinetics Experiments ............................................................................... 63 4.3.2 Adsorption Experiments ........................................................................... 65 4.3.3 Desorption Experiments ........................................................................... 68 4.4 Discussion ................................................................................................... 70 4.4.1 Thermodynamic Modeling ....................................................................... 70

iii

4.4.2 Adsorption Experiment Modeling ........................................................... 72 4.4.3 pH and Ionic Strength Effects .................................................................. 78 4.4.4

Linear Free-Energy Correlation ............................................................ 80

4.5 Conclusions ................................................................................................ 84 CHAPTER 5: THE ADSORPTION OF AQUEOUS URANYL COMPLEXES ONTO BACILLUS SUBTILIS CELLS .......................................................................... 86 5.1 Introduction ................................................................................................ 86 5.2 Experimental Section .................................................................................. 89 5.2.1 Bacterial Growth ..................................................................................... 89 5.2.2 Control Experiments ................................................................................ 91 5.2.3 U Adsorption Experiments........................................................................ 91 5.2.4 U Desorption Experiments ....................................................................... 93 5.2.5 Ca Adsorption Experiments ...................................................................... 94 5.3 Modeling of Metal-Bacteria Adsorption ...................................................... 95 5.4 Results and Discussion ............................................................................... 97 5.4.1 Adsorption and Desorption Results........................................................... 97 5.4.2 Modeling Results.....................................................................................106 5.5 Conclusions ...............................................................................................115 CHAPTER 6: CONCLUSIONS ...................................................................... 117 BIBLIOGRAPHY ...........................................................................................120

iv

FIGURES

Figure 2.1. Generic structure of an imidazolium based ionic liquid .................... 10 Figure 2.2. UV-Vis Spectra of Bmim Cl at various pH values............................ 15 Figure 2.3. Percent Bmim Cl (9.3 X 10-4 M) adsorbed onto gibbsite, quartz, and Bacillus subtilis ................................................................................................. 17 Figure 2.4. Percent Bmim Cl (5.0 X 10-4 M) adsorbed onto 0.4 and 2.0 g / L SWy-1 ionic strength of 0.0001 M................................................................................. 19 Figure 2.5. Percent Bmim Cl (9.3 X 10-4 M) adsorbed onto 0.8, 1.0 and 1.2 g / L SWy-1 ionic strength of 0.0001 M..................................................................... 20 Figure 2.6. Percent Bmim Cl (9.3 X 10-4 M) adsorbed onto 0.8, 1.0 and 1.2 g / L SWy-1 ionic strength of 0.1 M .......................................................................... 21 Figure 2.7. Percent Bmim Cl (9.3 X 10-4 M) adsorbed as a function of SWy-1 concentration with an ionic strength of 0.1 and 0.0001 M .................................. 22 Figure 3.1. Typical calorimetric raw data for low pH proton adsorption............. 39 Figure 3.2. Corrected heat evolved (mJ) from three low pH proton adsorption titrations versus pH ........................................................................................................ 40

v

Figure 3.3. Corrected heat evolved (mJ) from two high pH proton adsorption titrations versus pH ........................................................................................................ 41 Figure 3.4. −

∑Q n

x

corr

versus total Cd added (mM) for the Cd adsorption titrations

x =1

at pH 5.9 and 5.3 ............................................................................................... 52 Figure 4.1. Percent of Np adsorbed as a function of time .................................. 64 Figure 4.2. Percent of Np adsorbed as a function of pH with I = 0.1 M ............. 67 Figure 4.3. Percent of Np adsorbed as a function of Np concentration .............. 76 Figure 4.4. Percent of Np adsorbed as a function of pH with I = 0.0001 M ....... 77 Figure 4.5. Correlation plots showing calculated and previously published metal-carboxyl stability constants for B. subtilis as functions of aqueous metal-organic acid anion stability constants for acetate (A), oxalate (B), and citrate (C) .......................... 81 Figure 5.1. Ca released by B. subtilis (10 g/L wet weight) as a function of pH .. 90 Figure 5.2. U adsorption kinetics ...................................................................... 93 Figure 5.3. U adsorbed by B. subtilis as a function of pH and bacterial concentration, (a) 0.125, (b) 0.25, and (c) 0.5 g/L (wet mass), in a closed system (no CO2) ...... 98 Figure 5.4. U adsorbed by B. subtilis as a function of pH and bacterial concentration, (a) 0.125, (b) 0.25, and (c) 0.5 g/L (wet mass), in a open system ......................100 Figure 5.5. U adsorbed by B. subtilis as a function of pH and bacterial concentration, (a) 0.125 and (b) 0.25 g/L (wet mass), in a open system with 10mM Ca ..........102 Figure 5.6. Ca adsorbed by B. subtilis (10 g/L wet weight) as a function of pH 105

vi

TABLES

Table 2.1 Sorbent Properties.............................................................................. 24 Table 3.1 Site-specific thermodynamic parameters for the reaction of H+ and Cd2+ with the surface of B. subtilis in 0.1 M NaClO4 at 25.0 °C................................. 45 Table 4.1 Experimental Conditions for Np Datasets........................................... 66 Table 4.2 Np-Bacterial Stability Constants ........................................................ 79 Table 5.1 Aqueous U Complexation Stability Constants ..................................107 Table 5.2 U-Bacterial Stability Constants, Ca-Bearing Aqueous Complexation Reactions, and Ca-Bacterial Stability Constants ...............................................110

vii

ACKNOWLEDGMENTS

First I would like to thank Dr. Jeremy B. Fein for his guidance, patience, and friendship. I don’t think it’s possible to find a better advisor. A large portion of the research in this dissertation was conducted at Argonne National Lab in the Actinide Facility. I would like to thank Dr. Lynne Soderholm for making that collaboration possible. A special thanks to Dr. Mark P. Jensen from whom I have learned so much and truly enjoyed working with. You are a gifted scientist and an exceptional teacher. I would also like to thank the members of my committee Drs. Patricia A. Maurice, Peter C. Burns, and Clive R. Neal. Each of you have brought a unique perspective to my work and graduate career.

I would not have made it to graduate school if not for Drs. Michael M. Haley, Mark H. Reed, Karen Kelskey, and Nancy Deans. Thank you all for your support and taking the time to prepare me for my graduate career.

Drs. Jo Trigilio and Sal Johnston gave me the first glimpse of what my future may hold and I thank them for their friendship.

viii

I received a GAANN fellowship and additional funding through the Environmental Molecular Science Institute during my graduate career. This funding provided me with travel and research experiences that otherwise would have been unavailable. Numerous portions of the research contained in this dissertation would not have been possible without the equipment and staff at the Center for Environmental Science and Technology at the University of Notre Dame.

ix

CHAPTER 1 INTRODUCTION

Biogeochemical cycles, activities related to industry, weapons production, mining, nuclear energy, and other processes have introduced chemical contaminants into the environment. As contaminants migrate through the subsurface they encounter a range of geologic surfaces that may retard the mobility of the contaminants through a variety of chemical reactions. Adsorption reactions are one type of interaction to consider when examining the migration of contaminants through the subsurface. The mobility of chemical contaminants in the subsurface can be highly influenced by adsorption onto geological surfaces (Beveridge and Murray, 1976; Sposito, 1984; Waite et al., 1994; Fein et al., 1997; Macaskie and Basnakova, 1998; Rheinlander et al., 1998; Stipp et al., 2002; Stewart et al., 2003; Garelick et al., 2005). Since adsorption reactions can have such an impact on the mobility of pollutants, it is necessary to be able to quantify adsorption onto common geologic surfaces. This dissertation encompasses four individual research projects describing the adsorption and quantification of chemical contaminants onto geologic surfaces with special consideration of the reactivity of the bacterial surface.

1

initial data point of the desorption experiment. We depict these data as separate from the adsorption edge above pH 4.5, and neglect them in subsequent adsorption modeling, because it is likely that a similar irreversible reaction controls the Np removal under each of these low pH, high ionic strength experimental conditions.

The above observations suggest that Np removal from solution at high ionic strength and low pH is not purely an adsorption process, and is likely influenced by reduction of Np(V) to Np(IV). The decrease in Np removal from pH 2.5 to 4.5 (Figure 4.2 data shown enclosed by the oval) and the increasing Np removal with time for the pH 2.5 system (open circles in Figure 4.1) are both consistent with reduction of Np(V) to Np(IV) and continued reduction over the course of the experiments. If indeed this reduction does occur, our experiment cannot provide information on whether the reduction leads to enhanced adsorption of Np+4 relative to NpO2+, or whether precipitation of a Np(IV) phase causes the enhanced Np removal at low pH. Nevertheless, Cr(VI) exhibits similar behavior under comparable experimental conditions (Fein et al., 2002). Non-metabolic B. subtilis cells, grown and harvested in a similar manner to the cells used in our experiments, reduce Cr(VI) to Cr(III) in the absence of an external electron donor. Fein et al. (2002) observed continuously increasing Cr(VI) removal from solution by B. subtilis at pH 2.3 for at least 100 hours. Fein et al. (2002) also determined that the removal was irreversible and that the kinetics of removal increased with decreasing pH. X-ray adsorption near edge spectroscopy (XANES) confirmed the reduction of Cr(VI) to Cr(III) by the bacterial cell wall in these experimental systems. While we do not have spectroscopic confirmation of Np reduction

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Fein J. B., Boily J.-F., Yee N., Gorman-Lewis D., and Turner B. F. (2005) Potentiometric titrations of Bacillus subtilis cells to low pH and a comparison of modeling approaches. Geochimica et Cosmochimica Acta 69(5), 1123-1132. Fein J. B., Daughney C. J., Yee N., and Davis T. A. (1997) A chemical equilibrium model for metal adsorption onto bacterial surfaces. Geochimica et Cosmochimica Acta 61(16), 3319-3328. Fein J. B. and Delea D. (1999) Experimental study of the effect of EDTA on Cd adsorption by Bacillus subtilis: a test of the chemical equilibrium approach. Chemical Geology 161(4), 375-383. Fein J. B., Fowle D. A., Cahill J., Kemner K., Boyanov M., and Bunker B. (2002) Nonmetabolic reduction of Cr(VI) by bacterial surfaces under nutrient-absent conditions. Geomicrobiology Journal 19(3), 369-382. Fein J. B., Martin A. M., and Wightman P. G. (2001) Metal adsorption onto bacterial surfaces: development of a predictive approach. Geochimica et Cosmochimica Acta 65(23), 4267-4273. Ferris F. G., Schultze S., Witten T. C., Fyfe W. S., and Beveridge T. J. (1989) Metal interactions with microbial biofilms in acidic and neutral pH environments. Applied and Environmental Microbiology 55(5), 1249-57. Fowle D. A. and Fein J. B. (1999) Competitive adsorption of metal cations onto two gram positive bacteria: testing the chemical equilibrium model. Geochimica et Cosmochimica Acta 63(19/20), 3059-3067. Fowle D. A. and Fein J. B. (2000) Experimental measurements of the reversibility of metalbacteria adsorption reactions. Chemical Geology 168(1-2), 27-36. Fowle D. A., Fein J. B., and Martin A. M. (2000) Experimental study of uranyl adsorption onto Bacillus subtilis. Environmental Science and Technology 34(17), 3737-3741. Garelick H., Dybowska A., Valsami-Jones E., and Priest N. D. (2005) Remediation technologies for arsenic contaminated drinking waters. Journal of Soils and Sediments 5(3), 182-190. Giblin A. M., Batts B. D., and Swaine D. J. (1981) Laboratory simulation studies of uranium mobility in natural waters. Geochimica et Cosmochimica Acta 45(5), 699-709. Gorman-Lewis D., Elias P. E., and Fein J. B. (2005a) Adsorption of Aqueous Uranyl Complexes onto Bacillus subtilis Cells. Environmental Science and Technology 39(13), 4906-4912. Gorman-Lewis D., Fein Jeremy B., Soderholm L., Jensen M. P., and Chiang M. H. (2005b) Experimental Study of Neptunyl Adsorption onto Bacillus Subtilis. Geochimica et Cosmochimica Acta 29(20), 4837-4844.

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Journal of Hazardous Materials B126 (2005) 78–85

Biological chromium(VI) reduction using a trickling filter E. Dermou a , A. Velissariou b , D. Xenos a , D.V. Vayenas a,∗ a

Department of Environmental and Natural Resources Management, University of Ioannina, Seferi 2, 30100 Agrinio, Greece b Hellenic Aerospace Industry S.A., P.O. Box 23, GR 32009, Schimatari, Greece Received 27 January 2005; received in revised form 2 June 2005; accepted 2 June 2005 Available online 27 July 2005

Abstract A pilot-scale trickling filter was constructed and tested for biological chromium(VI) removal from industrial wastewater. Indigenous bacteria from industrial sludge were enriched and used as inoculum for the filter. Sodium acetate was used as carbon source and it was found to inhibit chromate reduction at high concentrations. Three different operating modes were used to investigate the optimal performance and efficiency of the filter, i.e. batch, continuous and SBR with recirculation. The latter one was found to achieve removal rates up to 530 g Cr(VI)/m2 d, while aeration was taking place naturally without the use of any external mechanical means. The low operating cost combined with the high hexavalent chromium reduction rates indicates that this technology may offer a feasible solution to a very serious environmental problem. © 2005 Elsevier B.V. All rights reserved. Keywords: Chromate; Biological removal; Trickling filter; Sequencing batch reactor; Recirculation

1. Introduction Chromium is one of the most toxic heavy metals discharged into the environment through various industrial wastewaters, and has become a serious health problem. Metal plating, tanneries and industrial processes using catalysts discharge worldwide huge amounts of chromium every year. The effluents of these industries contain Cr(VI) and Cr(III) at concentrations ranging from tenths to hundreds of milligrams/liter. While Cr(VI) is highly toxic and is known to be carcinogenic and mutagenic to living organisms [1], Cr(III) is generally only toxic to plants at very high concentrations and is less toxic or non-toxic to animals [2]. The discharge of Cr(VI) to surface water is regulated to below 0.05 mg/l by the US EPA [3] and the European Union [4], while total Cr, including Cr(III), Cr(VI) and its other forms, is regulated to below 2 mg/l [3]. At present, the most commonly used technology for treatment of heavy metals in wastewaters is chemical precipitation. Conventional chemical treatment involves reduction of Cr(VI) to Cr(III) by a reducing agent under ∗

Corresponding author. Tel.: +30 26410 39517; fax: +30 26410 39576. E-mail address: [email protected] (D.V. Vayenas).

0304-3894/$ – see front matter © 2005 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2005.06.008

low-pH conditions and subsequent adjustment of solution pH to near neutral ranges to precipitate Cr(III) as hydroxides [5]. However, this method is not completely satisfactory because of the large amount of secondary waste products due to various reagents used in the above-mentioned processes. Biological treatments arouse great interest because of their lower impact on the environment as opposed to chemical treatments. Recent studies have shown that certain species of bacteria are capable of transforming hexavalent chromium, Cr(VI), into the much less toxic and less mobile trivalent form, Cr(III) [6,7]. Bacteria may protect themselves from toxic substances in the environment by transforming toxic compounds through oxidation, reduction or methylation into more volatile, less toxic or readily precipitating forms. The processes by which microorganisms interact with toxic metals enabling their removal/and recovery are bioaccumulation, biosorption and enzymatic reduction [8]. Microbial heavy metal accumulation often comprises of two phases. An initial rapid phase involving physical adsorption or ion exchange at cell surface and by a subsequent slower phase involving active metabolism-dependent transport of metal into bacterial cells [9]. Biosorption is a metabolism-independent process and thus can be performed by both living and dead microorganisms. This adsorption is

E. Dermou et al. / Journal of Hazardous Materials B126 (2005) 78–85

based on mechanisms such as complexation, ion exchange, coordination, adsorption, chelation and microprecipitation, which may be synergistically or independently involved [10]. Enzymatic reduction of Cr(VI) into Cr(III) is believed to be one of the defense mechanisms employed by microorganisms living in Cr(VI)-contaminated environments. The reduced Cr(III) may precipitate as chromium hydroxide in neutral pH range [11]. Most of the previous studies on biological reduction of Cr(VI) were conducted in batch reactors (flasks) using mainly pure cultures. For instance, Wang and Xiao [12] studied several factors affecting hexavalent chromium reduction in pure cultures of bacteria in flasks. Wang and Shen [5] studied the kinetics of Cr(VI) reduction by pure bacterial cultures in flasks. Shakoori et al. [13] isolated a dichromate-resistant Gram-positive bacterium from effluent of tanneries and used flasks as batch reactors. Fein et al. [14] used pure bacterial cultures in flasks to study the non-metabolic reduction of Cr(VI) by bacteria under nutrient-absent conditions. Srinath et al. [8] studied Cr(VI) biosorption and bioaccumulation by pure cultures of chromate resistant bacteria in flasks. Megharaj et al. [15] studied hexavalent chromium reduction in flasks, by pure cultures of bacteria isolated from soil contaminated with tannery waste. Recently, continuous-flow and fixed-film bioreactors were also used for biological reduction of Cr(VI). Shen and Wang [16] demonstrated Cr(VI) reduction in a two-stage, continuous-flow suspended growth bioreactor system. Escherichia coli cells grown in the first-stage completely mixed reactor were pumped into the second-stage plug-flow reactor to reduce Cr(VI). Chirwa and Wang [11] demonstrated the potential of fixed-film bioreactors for Cr(VI) reduction. This was the first report on Cr(VI) reduction through biological mechanisms in a continuous-flow laboratory-scale biofilm reactor without the need to constantly resupply fresh Cr(VI)-reducing cells. Bacillus sp. was used in this work for the transformation of Cr(VI) into Cr(III). Virtually all the previous studies on biological reduction of Cr(VI) were conducted in laboratory scale apparatus (reactors), using sterilized conditions and pure cultures of microorganisms. The present study is the first to report on Cr(VI) biological reduction in a pilot-scale trickling filter using mixed culture of microorganisms, originating from an industrial sludge. The operation of the trickling filter as a Sequencing Batch Reactor (SBR) with recirculation led to significantly high Cr(VI) reduction rates, thus promising a feasible technological solution to a serious environmental problem.

2. Materials and methods 2.1. Media The influent feed to the bioreactor was prepared by dissolving 1 g NH4 Cl, 0.2 g MgSO4 ·7H2 O, 0.001 g FeSO4 ·

79

7H2 O, 0.001 g CaCl2 ·2H2 O, 5 g CH3 COONa·3H2 O and 0.5 g K2 HPO4 in 1.0 l of tap water. 2.2. Reagents Stock Cr(VI) solution (500 mg/l) was prepared by dissolving 141.4 mg of 99.5% K2 Cr2 O7 , previously dried at 103 ◦ C for 2 h, in Milli-Q water and diluting to 100 ml. Diphenyl carbazide solution was prepared by dissolving 250 mg of 1,5-diphenylcarbazide in 50 ml of HPLC-grade acetone and storing in a brown bottle. Potassium hydrogen phthalate standard (KHP) was prepared by dissolving 425 mg in distilled water and diluting to 1000 ml. Digestion solution was prepared by dissolving 10.216 g K2 Cr2 O7 , previously dried at 103 ◦ C for 2 h, in 500 ml distilled water, 167 ml conc. H2 SO4 and 33.3 g HgSO4 and diluting to 1000 ml (for the determination of the COD values). 1,5-Diphenylcarbazide was purchased from Fluka Chemical, potassium dichromate was purchased from Sigma Chemical Co. All the others chemicals were purchased from Riedel-de Haen. 2.3. Analytical methods During all experiments, hexavalent chromium concentration, pH, temperature, dissolved oxygen concentration and TOC measurements were made on a daily basis. Samples were filtered through 0.45 ␮m –Millipore filters (GN-6 Metricel Grid 47 mm, Pall Corporation). Hexavalent chromium concentration was determined by the 3500-Cr D Colorimetric method according to Standard Methods for the Examination of Water and Wastewater [17]. Total organic carbon measurements (TOC) were conducted in order to determine the feed sodium acetate concentration both in the liquid culture (chemostat) and the liquid volume of the bioreactor, following the methods described in Standard Methods for the Examination of Water and Wastewater [17] by using, Total organic carbon analyzer (TOC-VCSH , SHIMAZDU Corporation, Japan). Total chromium concentration measurements were made according to Standard Methods for the Examination of Water and Wastewater [17] using an atomic absorption spectrophotometer (model AAS700, Perkin-elmer) (results not shown for total chromium concentrations). 2.4. Isolation and enrichment of indigenous bacteria Samples of industrial sludge were taken from the Hellenic Aerospace Industry S.A. In order to grow bacterial strains able to reduce hexavalent chromium, a sludge sample of 10 g was added in a 2 l Erlenmeyer flask and was diluted in an acetate-minimal medium and concentrated chromium solution (in the form of K2 Cr2 O7 ) resulting in a final hexavalent chromium concentration of 50 mg/l. The final volume of the solution was 1 l. Acetate-minimal medium (AMM) was comprising (per litre) 1 g NH4 Cl, 0.2 g MgSO4 ·7H2 O, 0.001 g

E. Dermou et al. / Journal of Hazardous Materials B126 (2005) 78–85 [14] J.B. Fein, D.A. Fowle, J. Cahill, K. Kemner, M. Boyanov, B. Bunker, Nonmetabolic reduction of Cr(VI) by bacterial surfaces under nutrient-absent conditions, Geomicrobiol. J. 19 (2002) 369– 382. [15] M. Megharaj, S. Avudainayagam, R. Naidu, Toxicity of hexavalent chromium and its reduction by bacteria isolated from soil contaminated with tannery waste, Curr. Microbiol. 47 (2003) 51– 54. [16] H. Shen, Y. Wang, Hexavalent chromium removal in two-stage bioreactor system, J. Environ. Eng. 121 (11) (1995) 798–804. [17] APHA, AWWA and WPCF, Standard Methods for the Examination of Water and Wastewater, 17th ed., American Public Health Association, American Water Works Association and Water Pollution Control Federation, Washington, DC, 1989.

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[18] J.R. Marchesi, T. Sato, A.J. Weightman, T.A. Martin, J.C. Fry, S.J. Hiom, W.G. Wade, Design and evaluation of useful bacteriumspecific PCR primers that amplify genes coding for bacterial 16S rRNA, Appl. Environ. Microbiol. 64 (1998) 795–799. [19] S.F. Altschul, T.L. Maden, A.A. Schaffer, J. Zhang, Z. Zhang, W. Miller, D.J. Lipman, Gapped BLAST and PSI-BLAST: a new generation of protein database search programs, Nucl. Acids Res. 25 (1997) 3389–3402. [20] P. Pattanapipitpaisal, N.L. Brown, L.E. Macaskie, Chromate reduction and 16S rRNA identification of bacteria isolated from a Cr(VI)contaminated site, Appl. Microbiol. Biotechnol. 57 (2001) 257–261. [21] R. Francisco, M.C. Alpoim, P.V. Morais, Diversity of chromiumresistant and -reducing bacteria in a chromium-contaminated sludge, J. Appl. Microbiol. 92 (2002) 837–843.

Langmuir 2004, 20, 11433-11442

11433

Elucidation of Functional Groups on Gram-Positive and Gram-Negative Bacterial Surfaces Using Infrared Spectroscopy Wei Jiang,*,† Anuradha Saxena,‡ Bongkeun Song,† Bess B. Ward,† Terry J. Beveridge,‡ and Satish C. B. Myneni†,§ Department of Geosciences, Princeton University, Princeton, New Jersey 08544, Department of Microbiology, University of Guelph, Guelph, Canada, and Earth Sciences Division, Lawrence Berkeley National Laboratory, Berkeley, California 94720 Received April 15, 2004. In Final Form: August 19, 2004 Surface functional group chemistry of intact Gram-positive and Gram-negative bacterial cells and their isolated cell walls was examined as a function of pH, growth phase, and growth media (for intact cells only) using attenuated total reflectance Fourier transform infrared (ATR-FTIR) spectroscopy. Infrared spectra of aqueous model organic molecules, representatives of the common functional groups found in bacterial cell walls (i.e., hydroxyl, carboxyl, phosphoryl, and amide groups), were also examined in order to assist the interpretation of the infrared spectra of bacterial samples. The surface sensitivity of the ATR-FTIR spectroscopic technique was evaluated using diatom cells, which possess a several-nanometers-thick layer of glycoprotein on their silica shells. The ATR-FTIR spectra of bacterial surfaces exhibit carboxyl, amide, phosphate, and carbohydrate related features, and these are identical for both Gram-positive and Gramnegative cells. These results provide direct evidence to the previously held conviction that the negative charge of bacterial surfaces is derived from the deprotonation of both carboxylates and phosphates. Variation in solution pH has only a minor effect on the secondary structure of the cell wall proteins. The cell surface functional group chemistry is altered neither by the growth phase nor by the growth medium of bacteria. This study reveals the universality of the functional group chemistry of bacterial cell surfaces.

1. Introduction Bacteria are ubiquitous in near-surface geological systems, and are known to play important role in different biogeochemical processes, including contaminant transport and degradation,1,2 mineral dissolution and precipitation,3 and metal sorption by minerals and their redox transformations.4-9 Bacteria-water interfacial chemistry is one of the critical variables that play a central role in mediating these bacterial reactions. In addition, bacterial transport through porous media, adhesion to minerals and biological tissue, response to antibiotics, and the formation and chemistry of biofilms are also modified by the bacterial surface chemistry.10-14 The composition and structure of bacterial cell walls, and their variation as a * Corresponding author. Telephone: (609) 258-3827. E-mail: [email protected]. † Princeton University. ‡ University of Guelph. § Lawrence Berkeley National Laboratory. (1) Corapcioglu, M. Y.; Kims. Water Resour. Res. 1995, 31, 26392648. (2) Watanabe, K.; Hamamura, N. Curr. Opin. Biotechnol. 2003, 14, 289-295. (3) Stillings, L. L.; Drever, J. I.; Brantley, S. L.; Sun, Y.; Oxburgh, R. Chem. Geol. 1996, 132, 79-90. (4) Berveridge, T. J.; Forsberg, C. W.; Doyle, R. J. J. Bacteriol. 1982, 150, 1438-1448. (5) Suzuki, T.; Miyata, H.; Kawai, K.; Takmizawa, K.; Tai, Y.; Okazaki, M. J. Bacteriol. 1992, 174, 5340-5345. (6) Jackson, T. A.; West, M. M.; Leppard, G. G. Environ. Sci. Technol. 1999, 33, 3795-3801. (7) Holman, H. N.; Perry, D. L.; Martin, M. C.; Lamble, G. M.; McKinney, W. R.; Hunter-Cevera, J. C. Geomicrobiol. J. 1999, 16, 307324. (8) Klaus-Joerger, T.; Joerger, R.; Olsson, E.; Granqvist, C. G. Trends Biotechnol. 2001, 19, 15-20. (9) Fein, J. B.; Fowle, D. A.; Cahill, J.; Kemner, K.; Boyanov, M.; Bunker, B. Geomicrobiol. J. 2002, 19, 369-382. (10) McWhirter, M. J.; Bremer, P. J.; Lamont, I. L.; McQuillan, A. J. Langmuir 2003, 19, 3575-3577.

function of different environmental variables (e.g., pH, solution and substrate composition), is responsible for most surface interactions of bacteria. While the bulk chemical composition of bacterial cell walls is often known, their ability to complex metals and attach to surfaces as a function of different environmental conditions is not well understood. The Gram-positive cell wall is primarily made up of peptidoglycan (ca. 40-80% of the dry weight of the wall), which is a polymer of N-acetylglucosamine and Nacetylmuramic acid, containing mainly carboxyl, amide, and hydroxyl functional groups.15 The two other important constituents of Gram-positive cell walls are teichoic acid, a polymer of glycopyranosyl glycerol phosphate, and teichuronic acid, which is similar to teichoic acid, but replaces the phosphate functional groups with carboxyls. The cell walls of Gram-negative bacteria are more complex due to the presence of an outer membrane in addition to a thin peptidoglycan layer, but do not contain teichoic or teichuronic acids.16 Instead, the outer membrane contains phospholipids, lipoproteins, lipopolysaccharides, and proteins. Several recent investigations examined the surface chemistry of intact bacterial cells and their cell walls using both macroscopic (e.g., potentiometric titration, ion adsorption) and molecular tools (microscopy and (11) Burdman, S.; Okon, Y.; Jurkevitch, E. Crit. Rev. Microbiol. 2000, 26, 91-110. (12) Nelson, Y. M.; Lion, L. W.; Shuler, M. L.; Ghiorse, W. C. Environ. Sci. Technol. 1996, 30, 2027-2035. (13) Decho, A. W. Continental Shelf Res. 2000, 20, 1257-1273. (14) Suci, P. A.; Geesey, G. G.; Tyler, B. J. J. Microbiol. Methods 2001, 46, 193-208. (15) Berveridge, T. J. Int. Rev. Cytol. 1981, 72, 229-317. (16) Perry, J. J.; Staley, J. T.; Lory, S. Microbial Life; Sinauer Associates, Inc.: Sunderland, MA, 2002; pp 61-100.

10.1021/la049043+ CCC: $27.50 © 2004 American Chemical Society Published on Web 11/16/2004

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8 Treatment Technologies for Chromium(VI)

Elisabeth L. Hawley, Rula A. Deeb, Michael C. Kavanaugh and James Jacobs R.G

CONTENTS 8.1 Treatment Concepts.................................................................................. 274 8.1.1 Introduction: Chemistry of Chromium .....................................274 8.1.2 Chemical Transformations ...........................................................276 8.1.2.1 Oxidation–reduction.......................................................276 8.1.2.2 Sorption ............................................................................277 8.1.2.3 Precipitation.....................................................................278 8.1.3 Biological Transformations ..........................................................279 8.1.4 Physical Remediation Processes .................................................280 8.2 Classification of Treatment Technologies...............................................280 8.2.1 Reduction of Toxicity....................................................................280 8.2.2 Destruction and Removal ............................................................281 8.2.3 Containment...................................................................................281 8.3 Toxicity Reduction Methods ....................................................................281 8.3.1 Chemical Reduction......................................................................282 8.3.2 Microbial Reduction......................................................................283 8.3.3 Phytoremediation ..........................................................................286 8.4 Removal Technologies ..............................................................................288 8.4.1 Ex-situ Technologies .....................................................................288 8.4.1.1 Ion Exchange ...................................................................288 8.4.1.2 Granular Activated Carbon...........................................289 8.4.1.3 Adsorbents .......................................................................290 8.4.1.4 Membrane Filtration.......................................................290 8.4.1.5 Soil Washing and Separation Technologies................292 8.4.2 In-situ Technologies ......................................................................293 8.4.2.1 In-Situ Soil Flushing.......................................................293 8.4.2.2 Electrokinetics..................................................................294

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is useful for increasing Cr(VI) to Cr(III) reaction rates, by Le Chatelier’s Principle. Natural precipitation of Cr(VI) is not a major removal mechanism. CaCrO4 was observed to precipitate naturally during summer months at a hazardous waste site (Palmers et al., 1990). Based on laboratory studies, BaCrO4 and Cr/Al coprecipitates were suggested to occur at other sites (Palmer et al., 1990; Palmer and Wittbrodt, 1991). Plating tank sludge at the first site contained PbCrO4, PbCrO4 ◊ H2O, and K2CrO4. However, the solids are highly soluble and are not a considerable removal mechanism for Cr(VI). These three processes (redox reactions, sorption, and precipitation) form the basis of both chemical and biological treatment processes used to influence the balance between Cr(III) and Cr(VI). 8.1.3

Biological Transformations

Microorganisms often carry out enzymatic redox reactions as part of their metabolic processes. Cr(VI) can also be reduced nonmetabolically by reactions that occur on bacterial surfaces (Fein et al., 2001). This has been postulated by Fein et al. (2001) as the dominant pathway for reduction in natural geologic settings. A third mechanism for Cr reduction involves intra-cellular precipitation (Cervantes et al., 2001). However, most studies have focused on the first mechanism, where Cr is reduced metabolically in the presence of large amounts of electron donors. Chemical reducing compounds that require biological interactions include the use of molasses, lactic acid, and proprietary formulations such as the Hydrogen Release Compound (HRC), and cheese whey. These chemicals provide a carbon source in the environment, but require biological transformations to generate hydrogen in an anaerobic setting. Bacteria can enzymatically reduce Cr(VI) by both aerobic and anaerobic pathways. However, other nonbiological Cr reduction pathways compete with the biological pathways. Under anaerobic conditions, biological reduction is slow so abiotic reduction by Fe(II) or hydrogen sulfide is expected to dominate. Microbial reduction only becomes kinetically important in aerobic environments (Fendorf et al., 2001). Oxygen concentrations in the system are the primary factor influencing reduction rate, followed by pH and geochemical conditions. Microorganisms are always present in the environment. Their role in Cr reduction is still being defined through research. Topics of interest include the role of bacterial surfaces in Cr reduction, new tools for monitoring transformations such as infrared spectromicroscopy (FTIR Beamline) (Holman et al., 1999) and coupled biological remediation/chemical reduction processes. Phytoremediation is the engineered use of plants in the environmental remediation process. Phytoremediation is also a cutting-edge topic in research. There are six basic subsets of phytoremediation: phytoaccumulation (also called phytoextraction or hyperaccumulation), phytodegradation (also called phytotransformation), phytovolatilization, phytostabilization, rhizodegradation (also called phytostimulation or plant-assisted bioremediation), and

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FIGURE 8.3 Injection of liquids using RIP lance equipment.

Cr(III) include bacteria (Psuedomonas, Micrococcus, Escherichia, Enterobacter, Bacillus, Aeromonas, Achromobacter, and Desulfomamaculum) (McLean and Beveridge, 1999), algae (Cervantes et al., 1994), yeasts, and fungi. External reduction reactions that are biologically mediated still require the presence of an external electron donor, such as Fe, Mn, or oxidized organic matter. The process is the same as chemical reduction, but is biologically mediated and is thus kinetically advantageous to nonbiological reactions, particularly under aerobic conditions. Alternatively, sulfur-reducing bacteria are stimulated to produce H2S, which serves as the reductant. Recent work by Fein et al. (2001) has shown that bacterial surfaces can also catalyze Cr reduction. Both eukaryotic and prokaryotic cells can actively transport Cr(VI) across their cell membrane. In yeasts, Cr(VI) may enter via the permease system, a nonspecific method of ion transport for anions such as phosphate (PO43–) and SO42– (Cervantes et al., 2001). Cr is toxic to yeast because it inhibits SO42– uptake. Differences in the retention of Cr(VI) by algae have been studied and reported. Green algae retain more Cr than red or brown algae (Cervantes

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remediation strategies combine multiple technologies and mechanisms. For example, toxicity reduction occurs when Cr(VI) is reduced to Cr(III). However, this will also result in containment, since Cr(III) will usually sorb or precipitate as a solid. Phytoremediation is a broad term that encompasses all three remediation strategies. Plants take up Cr, removing it permanently from the soil once they are harvested. Plants can also be used to stabilize contamination via reduction that occurs at the roots and subsequent precipitation or adsorption. Finally, this process converts Cr(VI) to Cr(III), reducing the Cr toxicity (Fein et al., 2001). Some remediation strategies employ a variety of different mechanisms owing to the complex nature of biological, geological and chemical processes and interactions. Scientists are just beginning to understand the mechanisms that contribute to remediation. For example, constructed wetlands are emerging as a way to reduce contaminants in a low-tech, natural setting. In a wetland, Cr(VI) will be reduced to Cr(III) and sorb to the soil, or be taken up by plants, algae, or bacteria. Cr may associate with organics or soil particles, and undergo colloidal transport. More hybrid technologies are emerging, as advantages of each technology and the collaborative mechanisms under environmental conditions become apparent. For example, electrokinetics is being employed to enhance biological reduction in the LasagnaTM process. Reversing the polarity of the electrodes periodically leads to contaminant migration back and forth through the bioactive zone. New technologies that synthesize multiple treatment approaches will only continue to emerge in the future.

Bibliography Allan, M.L. and Kukacka, L.E., 1995, Blast furnace slag-modified grouts for in situ stabilization of Cr-contaminated soil, Waste Management 15, 3, 193–202. Bailey, S.E., Olin, T.J., Bricka, R.M., and Adrian, D.D., 1999, A review of potentially low-cost sorbents for heavy metals, Water Resources, 33, 11, 2469–2479. Bohdziewicz, J., 2000, Removal of Cr ions (VI) from underground water in the hybrid complexation-ultrafiltration process, Desalination, 129, 227–235. Bryant, P.S., Petersen, J.N., Lee, J.M., and Brouns, T.M., 1992, Sorption of heavy metals by untreated red fir sawdust, Applied Biochemistry Biotechnology 34–35, 777–788. Cadena, F., Rizvi, R., and Peters, R.W., 1990, Feasibility studies for the removal of heavy metals from solution using tailored bentonite, in Hazardous and Industrial Wastes, Proceedings of the Twenty-Second Mid-Atlantic Industrial Waste Conference, Drexel University, pp. 77–94. California Regional Water Quality Control Board (CRWQCB), 2000, Waste Discharge Requirements for In-Situ Pilot-Study for the Chemical Reduction of Cr, Order No. R1–2000–54. Cervantes, C., Campos-Garcia, J., Devars, S., Gutierrez-Corona, F., Loza-Tavera, H., Torres-Guzman, J.C., and Moreno-Sanchez, R., 1994, Interactions of Cr with microorganisms and plants, FEMS Microbiology Reviews, in Heavy Metals, De Filippis, L.F. and Pallaghy, C.K., Sources and Biological Effects, in Advances in Limnology Series, Algae and Water Pollution, pp. 31–37.

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Cervantes, C., Campos-Garcia, J., Devars, S., Gutierrez-Corona, F., Loza-Tavera, H., Torres-Guzman, J.C., and Moreno-Sanchez, R., 2001, Interactions of Cr with microorganisms and plants, FEMS Microbiology Reviews, 25, 335–347. Colorado State University (CSU), 1988, Digital Operation Management Model of the North Boundary System at the Rocky Mountain Arsenal Near Denver, Colorado, Technical Report No. 16, Department of Civil Engineering. Corpapcioglu, M.O. and Huang, C.P., 1987, The adsorption of heavy metals onto hydrous activated carbon, Water Resources, 21, 9, 1031–1044. Dikshit, V.P., 1989, Removal of Cr(VI) by adsorption using sawdust, Nat. Acad. Sci. Lett., 12, 12, 419–421. Dushenkov, V., Kumar, P.B.A.N., Motto, H., and Raskin, I., 1995, Rhizofiltration: the use of plants to remove heavy metals from aqueous streams, Environment Science and Technology, 29, 5, 1239. Electrokinetics, Inc., 1994, Electro-Klean Electrokinetic Soil Processing, SITE technology Profile-Demonstration Program. Evanko, C.R. and Dzombak, D.A., 1997, Remediation of Metals-Contaminated Groundwater, Groundwater Remediation Technologies Analysis Center, Technology Evaluation Report TE–97–01. Evanko, C.R. and Dzombak, D.A., 2000, Remediation of metals-contaminated soil and groundwater, in Standard Handbook of Environmental Health, Science and Technology, Lehr, J., Ed., McGraw-Hill, New York, NY, pp. 14.100–14.134. Fein, J.B., Kemner, K., Fowle, D.A., Cahill, J., Boyanov, M., and Bunker, B., 2001, Non- Au: Please metabolic reduction of Cr(VI) by bacterial surfaces under nutrient-absent con- update this ditions, Geomicrobiology Journal, (in press). ref. Fendorf, S., Wielinga, B.W., and Hansel, C.M., 2001, Reduction of Cr in Surface and Subsurface Environments, Contributions of Biological and Abiological Processes, Eleventh Annual V. M. Goldschmidt Conference. Fruchter, J.S., Cole, C.R., Williams, M.D., Vermeul, V.R., Amonette, J.E., Szecsody, J.E., Istok, J.D., and Humphrey, M.D., 2000, Creation of a subsurface permeable treatment barrier using in situ redox manipulation, Groundwater Monitoring and Remediation Review. Gallinatti, J.D. and Warner, S.D., 1994, Hydraulic design considerations for permeable in-situ groundwater treatment wells, Ground Water 32, 5, 851. Gardea-Torresdey, J.L., Gonzalez, J.H., Tiemann, K.J., Rodriguea, O., and Gamez, G., 1998, Phytofiltration of hazardous cadmium, chromium, lead and zinc ions by biomass of medicago sativa (Alfalfa), Journal of Hazardous Materials, 57, 1–3, 29. Groundwater Resources Association (GRA), 1999, Design and Implementation of Permeable Reactive Barriers for Groundwater Treatment, San Francisco CA. Hafiane, A., Lemordant, D., and Dhahbi, M., 2000, Removal of Cr(VI) by nanofiltration, Desalination, 130, 305–312. Hamadi, N.K., Chen, X.D., Farid, M.M., and Lu, M.G.Q., 2001, Adsorption kinetics for the removal of Cr(VI) from aqueous solution by adsorbents derived from used tyres and sawdust, Chemical Engineering Journal, 84, 95–105. Haq, R.U. and Shakoori, A.R., 1998, Short Communication, Microbiological treatment of industrial wastes containing toxic Cr involving successive use of bacteria, yeast and algae, World Journal of Microbiology and Biotechnology, 14, 583–585. Henshaw Associates, Inc., 1998, Pilot-Scale GAC Treatment Study Results, Former Remco Hydraulics Facility, Draft Memorandum to Mr. John Farr, Ph.D., P.E. from Michael Harrison, P.E.

Cometabolism of Cr(VI) by Shewanella oneidensis MR-1 Produces Cell-Associated Reduced Chromium and Inhibits Growth Sarah S. Middleton,1 Rizlan Bencheikh Latmani,2 Mason R. Mackey,3 Mark H. Ellisman,3 Bradley M. Tebo,2 Craig S. Criddle1 1

Department of Civil and Environmental Engineering, Stanford University, Stanford, California 943051; telephone: (650) 723-9032; fax: (650) 725-3164; e-mail: [email protected]. 2 Marine Biology Research Division and Center for Marine Biotechnology and Biomedicine, Scripps Institution of Oceanography, University of California—San Diego, La Jolla, California 92093-02022 3 National Center for Microscopy and Imaging Research, University of California-San Diego, La Jolla, California 920933 Received 4 November 2002; accepted 1 April 2003 Published online 23 June 2003 in Wiley InterScience (www.interscience.wiley.com). DOI: 10.1002/bit.10725

Abstract: Microbial reduction is a promising strategy for chromium remediation, but the effects of competing electron acceptors are still poorly understood. We investigated chromate (Cr(VI)) reduction in batch cultures of Shewanella oneidensis MR-1 under aerobic and denitrifying conditions and in the absence of an additional electron acceptor. Growth and Cr(VI) removal patterns suggested a cometabolic reduction; in the absence of nitrate or oxygen, MR-1 reduced Cr(VI), but without any increase in viable cell counts and rates gradually decreased when cells were respiked. Only a small fraction (1.6%) of the electrons from lactate were transferred to Cr(VI). The 48-h transformation capacity (Tc) was 0.78 mg (15 µmoles) Cr(VI) reduced ⭈ [mg protein]−1 for high levels of Cr(VI) added as a single spike. For low levels of Cr(VI) added sequentially, Tc increased to 3.33 mg (64 µmoles) Cr(VI) reduced ⭈ [mg protein]−1, indicating that it is limited by toxicity at higher concentrations. During denitrification and aerobic growth, MR-1 reduced Cr(VI), with much faster rates under denitrifying conditions. Cr(VI) had no effect on nitrate reduction at 6 µM, was strongly inhibitory at 45 µM, and stopped nitrate reduction above 200 µM. Cr(VI) had no effect on aerobic growth at 60 µM, but severely inhibited growth above 150 µM. A factor that likely plays a role in Cr(VI) toxicity is intracellular reduced chromium. Transmission electron microscopy (TEM) and electron energy loss spectroscopy (EELS) of

Correspondence to: Craig S. Criddle Contract grant sponsor: the National Institute of Environmental Health Sciences (NIEHS) Superfund Basic Research Program Contract grant numbers: ES04911 (to the Institute for Environmental Toxicology at Michigan State University), ES10337 to the University of California San Diego) Contract grant sponsor: the National Science Foundation Graduate Fellowship Program (to SM) Contract grant sponsor: the Swiss National Science Foundation Post Graduate Fellowship (to RBL).

© 2003 Wiley Periodicals, Inc.

denitrifying cells exposed to Cr(VI) showed reduced chromium precipitates both extracellularly on the cell surface and, for the first time, as electron-dense round globules inside cells. © 2003 Wiley Periodicals, Inc. Biotechnol Bioeng 83: 627–637, 2003.

Keywords: Shewanella oneidensis MR-1; chromate; reduction; denitrification; inhibition

INTRODUCTION Hexavalent chromium [Cr(VI)] is an Environmental Protection Agency (EPA) priority pollutant and a known carcinogen. Due to its widespread industrial use, it is often found in contaminated groundwater. In order to achieve levels of chromium below the EPA maximum contaminant level (100 ␮g/L), remediation strategies focus on the reduction of Cr(VI) to insoluble trivalent forms, which are relatively stable and nontoxic. The only compounds able to oxidize trivalent chromium at any appreciable rate are manganese oxides (Eary and Rai, 1987). Bioremediation may be effective for the removal of Cr(VI) from groundwater, as many aerobic and anaerobic microorganisms reduce Cr(VI) to Cr(III) while utilizing a wide range of electron donors (Bopp and Ehrlich, 1988; Ishibashi et al., 1990; Shen and Wang, 1993; Rege et al., 1997; Tebo and Obraztsova, 1998; Francis et al., 2000; Myers et al., 2000). However, there are few studies that compare the kinetics of Cr(VI) reduction by bacteria under different electron-accepting conditions or that study the effects of Cr(VI) on the reduction of other electron acceptors. Shewanella oneidensis MR-1 (formerly Shewanella putrefaciens MR-1) is a facultative Gram-negative bacterium whose respiratory versatility has prompted interest in its use in bioremediation. A nonfermenting bacterium, MR-1 can

RESULTS AND DISCUSSION Cr(VI) Reduction Without Additional Electron Acceptors Present The capacity of MR-1 to reduce Cr(VI) in the absence of other electron acceptors was investigated. All experiments were performed in the minimal medium defined in the previous section. MR-1 reduced Cr(VI) with lactate as the electron donor (Fig. 1) but not formate (data not shown). Controls prepared without cells, without lactate, and with autoclaved cells showed little decrease in Cr(VI) concentration (approximately 20, 36, and 20 ␮M, respectively, over 10 days). Abiotic reduction of Cr(VI) could have resulted from lactate, yeast extract, bactopeptone, or Fe(II) in the medium. In controls with cells but without lactate, yeast extract and bactopeptone could have initially been used as electron donor. Cr(VI) reduction in the absence of electron donor has also been described by Fein et al. (2002). As expected for an anion such as CrO42−, adsorption was not observed at this pH. Cr(VI) reduction in autoclaved controls did not differ from abiotic controls. MR-1 cultures were unable to grow during the reduction of Cr(VI), as indicated by viable cell count (Fig. 1) and total cellular protein (data not shown). Colony-forming units did not increase after the addition of almost 175 ␮M Cr(VI). In addition, the specific activity of Cr(VI) reduction decreased with each addition of Cr(VI). If Cr(VI) reduction supported growth, the rate of reaction would increase. These results suggest secondary utilization or cometabolism as the mechanism for Cr(VI) reduction. Cr(III) Solubilization Total soluble chromium was measured at the end of experiments in order to probe for the presence of soluble Cr(III). At the pH of most groundwater, Cr(III) formed from Cr(VI)

reduction will precipitate as insoluble Cr(III) hydroxides (Jardine et al., 1999). However, Cr(III) is known to complex with organic ligands (Nieboer and Jusys, 1988). Soluble Cr(III) is commonly computed as the difference between total soluble Cr and total soluble Cr(VI). In medium containing lactate as electron donor, 13–14% of the reduced Cr(III) remained soluble, presumably as a complex to lactate or components of yeast extract or bactopeptone. When 100 mM HEPES was included in the medium as a buffer, the percentage of soluble Cr(III) increased to 47%. This observation is important for bioremediation of Cr(VI) because the EPA maximum contaminant level (MCL) is measured in terms of total soluble Cr, including both soluble Cr(VI) and Cr(III). Cr(VI) Reduction During Denitrification MR-1 can reduce nitrate to nitrite to N2O coupled to growth (Krause and Nealson, 1997). Batch studies were performed in minimal medium to study simultaneous Cr(VI) and nitrate reduction. Cultures were spiked with 5 mM lactate and ∼2 mM nitrate and allowed to grow for several hours to ensure the onset of nitrate reduction. Cr(VI) was then added at ∼6 and 45 ␮M to a subset of the bottles. As shown in Figure 2, in the absence of Cr(VI) lactate was oxidized to acetate, nitrate was reduced to nitrite, and nitrite was almost completely reduced (presumably to N2O). When Cr(VI) was added at 6 ␮M, Cr(VI) was reduced during nitrate and nitrite reduction. Patterns of lactate oxidation, nitrate and nitrite reduction, and growth were nearly identical to controls without Cr(VI). However, when Cr(VI) was added at 45 ␮M, nitrate reduction was immediately inhibited and nitrite reduction stopped. Cr(VI) concentrations were reduced to the MCL (2 ␮M). Controls without lactate showed no nitrate reduction (data not shown). No Cr (VI) reduction was observed in abiotic controls containing nitrate, nitrite, and Cr (VI) (data not shown).

Figure 1. Shewanella oneidensis MR-1 reduces Cr(VI) in the absence of other electron acceptors, but does not grow on Cr(VI). Where error bars are shown, values are the average of duplicates and error bars represent the range. 50 ␮M Cr(VI) was added at day 2 and 5. (⽧), MR-1 + lactate + Cr(VI); (䉭), MR-1 + Cr(VI); (䊉), Cr(VI) + lactate; (䊊), autoclaved MR-1 + lactate + Cr(VI); and (*), viable cell count for MR-1 + lactate + Cr(VI).

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Figure 8. A study of the peak position and the peak ratio relative to the oxidation state indicate that the chromium associated with cells (both intracellularly and extracellularly) does not correspond to Cr(VI). Results suggest the chromium is in a reduced state. (–), Cr inside the cells; (䊉), Cr outside the cells; (⽧), Cr(VI) standard; (䉭), Cr(III) oxide standard; (䊐), Cr(III)Cl3 standard.

tions. Bacteria often preferentially utilize more energetically favorable electron acceptors. For example, in the presence of nitrate, reduction of Mn(IV), thiosulfate, and Fe(III) in Shewanella putrefaciens 200 was inhibited, indicating that nitrate was the preferred electron acceptor for anaerobic respiration (DiChristina, 1992). Although oxygen and nitrate have half-reaction reduction potentials that are higher or comparable (respectively) to Cr(VI), MR-1 is able to reduce Cr(VI) while reducing both of these electron acceptors. These results suggest that in bioremediation applications, the ability of MR-1 to reduce Cr(VI) would not be inhibited by oxygen or nitrate. However, it is clear that kinetic parameters vary significantly between oxic and anoxic conditions. Our results also indicate that Cr(VI) inhibits aerobic growth and denitrification above certain levels. This may influence the sequence in which electron acceptors are reduced in a mixed waste setting. For example, the presence of toxic levels of Cr(VI) may hinder denitrification and prevent the redox potential of the system from decreasing to levels where other electron acceptors are reduced. Toxicity also appears to limit Cr(VI) transformation capacity; sequential spikes of lower levels of Cr(VI) allowed for a larger transformation capacity over a longer duration than single spikes of a higher Cr(VI) concentration. Although the mechanism of Cr(VI) toxicity in MR-1 is not well understood, TEM and EELS studies revealed reduced chromium

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precipitates both within and surrounding cells, including intracellular reduced chromium globules, that may contribute to the inhibitory effects of chromium. We thank James Bouwer from the National Center for Microscopy and Imaging Research for help with the Gatan Imaging Filter. We also thank Jizhong Zhou for supplying the MR-1 strain used in this study, and Alfred Spormann and Weimin Wu for helpful suggestions on the manuscript.

References Alvarez-Cohen L, McCarty PL. 1991. A cometabolic biotransformation model for halogenated aliphatic compounds exhibiting product toxicity. Environ Sci Technol 25:1381–1387. Bopp LH, Ehrlich HL. 1988. Chromate resistance and reduction in Pseudomonas fluorescens strain LB300. Arch Microbiol 150:426–431. Daulton TL, Little BJ, Lowe K, Jones-Meeham J. 2002. Electron energy loss spectroscopy techniques for the study of microbial chromium(VI) reduction. J Microbiol Methods 50:39–54. DiChristina TJ. 1992. The effects of nitrate and nitrite on dissimilatory iron reduction in Shewanella putrefaciens 200. J Bacteriol 174:1891–1896. Eary L, Rai D. 1987. Kinetics of chromium(III) oxidation to chromium(VI) by reaction with manganese dioxide. Environ Sci Technol 21: 1187–1193. Egerton RF. 1996. Electron energy-loss spectroscopy in the electron microscope. New York: Plenum Press. Fein JB, Fowle DA, Cahill J, et al. 2002. Nonmetabolic reduction of Cr(VI) by bacterial surfaces under nutrient-absent conditions. Geomicrobiol J 19:369–382.

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Rev Environ Contam Toxicol 178:93–164

Chromium–Microorganism Interactions in Soils: Remediation Implications Sara P.B. Kamaludeen, Mallavarapu Megharaj, Albert L. Juhasz, Nabrattil Sethunathan, and Ravi Naidu Contents I. Introduction .......................................................................................................... A. Forms of Chromium ....................................................................................... B. Sources of Chromium in Soil ......................................................................... C. Chromium Transformations in Soil ................................................................ II. Physicochemical Factors Governing Chromium Transformations in Soil ........ A. Soil Physical Factors ...................................................................................... B. Soil pH ............................................................................................................ C. Organic Matter ................................................................................................ D. Iron .................................................................................................................. E. Manganese ....................................................................................................... III. Microbiological Factors Governing Chromium Transformations in Soil .......... A. Resistance or Tolerance to Cr(VI) ................................................................. B. Direct Cr(VI) Reduction ................................................................................. C. Indirect Reduction ........................................................................................... D. Biotic–Abiotic Coupling in Mn Oxide-Mediated Oxidation of Chromium(III) ............................................................................................ IV. Implications of Chromium Transformations on Microorganisms and their Activities .............................................................................................. A. Microorganisms .............................................................................................. B. Effects on Soil Microbial Community ........................................................... C. Effect on Soil Microbial Processes and Activities ........................................ V. Remediation of Chromium-Contaminated Water and Soils ............................... A. Remediation Technologies for Wastewater and Solutions ............................ B. Remediation Technologies for Chromium Wastes in Soils .......................... C. Bioremediation ................................................................................................ D. Applicability of Phytostabilization to Cr-Contaminated Soil ........................ VI. Challenges ............................................................................................................ Summary ....................................................................................................................

94 95 95 97 97 97 97 97 98 99 103 103 104 119 121 126 126 130 132 141 141 143 144 147 148 148

Communicated by G.W. Ware. S.P.B. Kamaludeen The University of Adelaide, Department of Soil and Water, Waite Campus, Glen Osmond, SA 5064, Australia and Tamil Nadu Agricultural University, Trichy Campus, Trichy, Tamil Nadu, India. M. Megharaj ( ), A.L. Juhasz, N. Sethunathan, R. Naidu, (formerly CSIRO Land and Water, Adelaide), Australian Centre for Environmental Assessment and Remediation, University of South Australia, Mawson Lakes Campus, Mawson Lakes, SA 5095, Australia.

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Anabaena variabilis

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—

—

—

—

—

Cr tolerance level

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—

—

Source

Chlorella

Algae Spirogyra sp. and Mougeotia sp. Oscillatoria

Bacillus subtilis Sporosarcina ureae Shewanella putrefaciens

Identification

Table 1. (Continued).

Transitory formation of Cr(V) during Cr(VI) reduction Enzymatically reduced Cr(VI) Enzymatically reduced Cr(VI) Chromate reduced in 18 d to Cr(III); 50% of Cr(III) formed accumulated in the cells and 50% in the medium; Cr(VI) reduction associated with heterocysts

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Cr-reducing efficiency Reference

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—

Garnham and Green 1995

Losi et al. 1994b

Losi et al. 1994b

Nonmetabolizing bacte- Fein et al. 2002 rial cells reduced Cr(VI) on cell surface in absence of externally supplied electron donors; not coupled to oxidation of bacterial exudates; fastest under acidic conditions — Liu et al. 1995

Mechanism

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physiological state of the culture, was possibly inducible under anaerobic conditions. Cr(VI) reduction in the anaerobically grown stationary phase of this bacterium is a complex process, possibly involving more than one pathway (Viamajala et al. 2002b). A wide range of organic pollutants such as phenol, 2-chlorophenol, p-cresol, 2,6-dimethylphenol, 3,5-dimethylphenol, 3,4-dimethylphenol, benzene, and toluene can also serve as electron donors for Cr(VI) reduction in cocultures containing E. coli ATCC33456 and P. putida DMP-1 (Shen and Wang 1995). Metabolites produced during phenol degradation by P. putida served as electron donors for Cr(VI) reduction by E. coli. Technology using such cocultures would help to simultaneously detoxify both organic pollutants and the toxic Cr(VI). Nonmetabolizing resting cells of bacteria could reduce Cr(VI), but only in the presence of an added carbon source (Bopp and Ehrlich 1988; Shen and Wang 1994b; Philip et al. 1998). Killed resting cells could not cause Cr(VI) reduction (Shen and Wang 1994b; Wang and Shen 1997). Soluble enzymes in cell extracts can reduce Cr(VI) in the presence (Horitsu et al. 1987; Philip et al. 1998) or absence (Bopp and Ehrlich 1988; Shen and Wang 1994b) of added electron donors. According to very recent evidence, nonmetabolic Cr(VI) reduction can occur on bacterial surfaces even in the absence of externally added electron donors in the medium. Thus, Fein et al. (2002) demonstrated that nonmetabolizing cells of Bacillus subtilis, Sporosarcina ureae, and Shewanella putrefaciens could reduce significant amounts of Cr(VI) in the absence of externally supplied electron donors. The Cr(VI) reduction by the bacterial strains was dependent on solution pH, decreasing with increasing pH, and presumably occurred at the cell wall and independent of the oxidation of bacterial organic exudates. Such nonmetabolizing reduction of Cr(VI) by bacteria in nutrient-poor conditions may be important in the biogeochemical distribution of Cr. Cr(VI) reduction by microorganisms, known to occur under both aerobic and anaerobic conditions (see Table 1), is a redox-sensitive process (Shen and Wang 1994b; Chen and Hao 1996). The ability of washed resting cells of Agrobacterium radiobacter EPS-916 to reduce Cr(VI) was governed by their redox potenial (Llovera et al. 1993). Resting cells of A. radiobacter EPS-916, pregrown under aerobic conditions on glucose, fructose, maltose, lactose, mannitol, or glycerol as the sole carbon and energy source, exhibited similar redox potentials of around −200 mV and completely reduced 0.5 mM chromate. On the other hand, the inability of the resting cells of the bacterium, pregrown on glutamate or succinate, to reduce chromate was associated with relatively high redox potentials of −138 to −132 mV. Moreover, resting cells pregrown under anaerobic conditions on glucose had lower redox potentials (−240 mV) and a more pronounced chromate-reducing activity than did the aerobically grown resting cells on glucose with a redox potential of −200 mV. Likewise, cells pregrown anaerobically on chromate as the electron acceptor effected more rapid reduction of chromate than did the anaeorobically grown cells (−198 mV) on nitrate. Evidence suggested a negative correlation between chromate reduction by the rest-

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copper, nickel, lead and zinc on soil microbial respiration in relation to abiotic soil factors. Plant Soil 79:317–327. Doelman P, Haanstra L (1986) Short- and long-term effect of heavy metals on urease activity in soils. Soil Biol Biochem 2:213–218. Doelman P, Haanstra L (1989) Short- and long-term effects of heavy metals on phosphatase activity in soils: an ecological dose-response model approach. Biol Fertil Soils 8:235–242. Eary LE, Rai D (1987) Kinetics of chromium(III) oxidation to chromium(VI) by reaction with manganese dioxide. Environ Sci Technol 21:1187–1193. Eary LE, Rai D (1988) Chromate removal from aqueous wastes by reduction with ferrous iron. Environ Sci Technol 22:972–977. Edenborn HM, Paquin Y, Chateauneuf G (1985) Bacterial contribution to manganese oxidation in a deep coastal sediment. Estuar Coast Mar Sci 21:801–815. Efstathiou JD, McKay LL (1977) Inorganic salts resistance associated with a lactosefermenting plasmid in Streptococcus lactis. J Bacteriol 13:257–265. Ehrlich HL (1976) Manganese as an energy source for bacteria. In: Nriagu JO (ed) Environmental Biogeochemistry, vol 2. Metals Transfer and Ecological Mass Balances. Ann Arbor Science, Ann Arbor, MI. Ehrlich HL (1981) Geomicrobiology. Dekker, New York. Ehrlich HL (1982) Enhanced removal of Mn2+ from seawater by marine sediments and clay minerals in the presence of bacteria. Can J Microbiol 28:1389–1395. Emerson S, Kalhorn S, Jacobs L (1982) Environmental oxidation rate of manganese (II): bacterial catalysis. Geochim Cosmochim Acta 46:1073–1079. Fein JB, Fowle DA, Cahill J, Kemner K, Boyanov M, Bunker B (2002) Nonmetabolic reduction of Cr(VI) by bacterial surfaces under nutrient-absent conditions. Geomicrobiol J 19:369–382. Fendorf SE, Li G (1996) Kinetics of chromate reduction by ferrous iron. Environ Sci Technol 30:1614–1617. Fendorf S, Zasoski RJ (1992) Chromium(III) oxidation by delta-manganese oxide: 1. Characterization. Environ Sci Technol 26:79–85. Fendorf SE, Zasoski RJ, Burau RG (1993) Competing metal ion influences on chromium(III) oxidation by birnessite. Soil Sci Soc Am J 57:1508–1515. Fendorf SE, Li G, Gunter ME (1996) Micromorphologies and stabilities of chromium (III) surface precipitates elucidated by scanning force microscopy. Soil Sci Soc Am J 60:99–106. Francis CA, Obraztsova AY, Tebo BM (2000) Dissimilatory metal reduction by the facultative anaerobe Pantoea agglomerans SP1. Appl Environ Microbiol 66:543–548. Francisco R, Alpoim MC, Morais PV (2002) Diversity of chromium-resistant and -reducing bacteria in a chromium-contaminated activated sludge. J Appl Microbiol 92:837– 843. Frankenberger WT Jr, Tabatabai MA (1981) Amidase activity in soils: IV. Effects of trace elements and pesticides. Soil Sci Soc Am J 45:1120–1124. Frankenberger WT Jr, Tabatabai MA (1991) Factors affecting L-asparaginase activity in soils. Biol Fertil Soils 11:1–5. Fredrickson JK, Kostandarithes HM, Li SW, Plymale AE, Daly MJ (2000) Reduction of Fe(III), Cr(VI), U(VI), and Te(VII) by Deinococcus radiodurans R1. Appl Environ Microbiol 66:2006–2011. ˚ , Tunlid A, Ba˚a˚th E (1993) Phospholipid fatty acid composition, biomass, Frostega˚rd A