Characterization of Five Chromium-Removing Bacteria Isolated from Chromium-Contaminated Soil

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

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 chromi- Keywords Chromium-removing bacteria . um pollution using bacteria was investigated in this Pseudochrobactrum saccharolyticum . Aerobic process . study. Five chromium-removing bacteria strains were Biotransformations . Bioremediation . Waste treatment 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– 1 Introduction 11) for the five isolates were investigated. The effect of initial Cr(VI) concentrations (0–1,575 mg L−1)onbac- Among heavy-metal pollutants, chromium (Cr) is con- terial growth was also studied. Results showed that sidered to be toxic and one of the main pollutants Pseudochrobactrum saccharolyticum strain W1 had (Yewalkar et al. 2007). Chromium is widely used in high chromium-removing ability and could grow at − industrial operations such as leather tanning, pigment Cr(VI) concentrations from 0 to 1,225 mg L 1.Toour production, electroplating, paints, steel manufacture, knowledge, this is the first report of chromium removal and automobile production (Wang and Xiao 1995; by a member of the Pseudochrobactrum genus. Pattanapipitpaisal et al. 2001). Intensive industrial ap- Sporosarcina saromensis W5 had the highest − − plications of chromium and releases of associated waste chromium-removing rate of 0.79 mg h 1 mg 1 biomass. have caused substantial soil contamination. Chromium Exopolysaccharide (EPS) production and components exists in several oxidation states, from −2 to 6 (Jacobs of the five bacteria strains were also investigated, and a and Testa 2005). However, in the environment, the most positive relationship was found between the bacterial stable and common forms are the trivalent [Cr(III)] and chromium removal and EPS production. hexavalent [Cr(VI)] species (Fendorf 1995). The Cr(VI) : : form is more reactive and harmful than the trivalent Z. He S. Li L. Wang (Francisco et al. 2002) which is, in comparison, less School of Minerals Processing and Bioengineering, Central South University, toxic, less soluble, and less mobile than the hexavalent Changsha 410083, People’s Republic of China form (Stanin 2005). In recent years, more and more : : attention has been focused on the bioremediation of Z. He S. Li L. Wang Cr(VI) contamination with chromate-resistant bacteria Key Laboratory of Biohydrometallurgy of Ministry of Education, Central South University, (Zhu et al. 2008). Changsha, Hunan, China 410083 In this work, plating method was used to isolate Cr(VI)-removing bacteria from chromium- * H. Zhong ( ) contaminated soil samples adjacent to a chromium land- School of Life Science, Central South University, Changsha, Hunan, China 410012 fill in Changsha, Hunan province, China. The growth e-mail: [email protected] conditions, such as pH, temperature, and the initial 1904, Page 6 of 10 Water Air Soil Pollut (2014) 225:1904

A number of previous studies have shown that many Cr(VI) concentration of 1,050 mg L−1.Inverylimited microorganisms possess Cr(VI) tolerance/resistance. As previous studies, Sporosarcina sp. have been found outlined in Table 1, P. saccharolyticum strain W1, capable of growing in Cr(VI) concentration of 5 ppm Oceanobacillus sp. W4, and S. saromensis W5 had (5 mg L−1) (Fein et al. 2002) and tolerate 2,900 μM higher Cr(VI)-resistance ability than most other previ- (=150.8 mg L−1)ofCr(Bafana2011), while ously identified strains, which highlight the potential of S. saromensis W5 could tolerate Cr(VI) concentration these three isolates as bioremediators of Cr(VI) from of up to 1, 400 mg L−1. Most previous studies have chromium-polluted areas. Pseudochrobactrum was re- found that Cr(VI) inhibits bacterial growth at any Cr(VI) cently proposed by Kämpfer et al. (2006) and comprises concentration (He et al. 2009; Middleton et al. 2003; five species to date: P. saccharolyticum (Kampfer et al. Camargo et al. 2003). It has been reported that the 2006), Pseudochrobactrum asaccharolyticum (Kampfer growth of Arthrobacter sp. and Bacillus sp. was stimu- et al. 2006), Pseudochrobactrum kiredjianiae (Kampfer lated by Cr(VI) concentrations of 50 and 5 mg L−1, et al. 2007), Pseudochrobactrum lubricantis (Kampfer respectively (Megharaj et al. 2003). In this study, growth et al. 2009), and Pseudochrobactrum glaciei of the five isolates was promoted by Cr(VI) when under (Romanenko et al. 2008). Pseudochrobactrum sp. certain Cr(VI) concentrations. P. saccharolyticum strain (Kampfer et al. 2006; Kampfer et al. 2007; Kampfer W1 and S. saromensisW5 were stimulated by Cr(VI) et al. 2009; Romanenko et al. 2008) was able to grow at concentration from 0 to 875 mg L−1 and from 175 to 15–40 °C (optimum temperature was 25–30 °C) and the 1,400 mg L−1, respectively. This mechanism was not optimum pH value was about 7.1–7.5. In this study, the explicit and need to be further studied. optimum temperature for P. saccharolyticum strain W1 was 35 °C and it could grow at a pH range of 8.0–11.0, 3.3 Cr(VI) Removal by the Five Isolates with optimum pH value of 9.5. To our knowledge, this is the first report of chromium resistance by a strain from The ability of the five isolates to remove Cr(VI) when Pseudochrobactrum sp. incubated in their respective optimum initial Cr(VI) Oceanobacillus sp. was initially reported to grow at concentrations for growth was studied. Abiotic removal pH 9–10 and at 15–40 °C with the optimum temperature of Cr(VI) has also been evaluated and was found to be at 30–36 °C and was identified as a halotolerant obligate neglectable. The changes in Cr(VI) concentration in the alkaliphile isolated from the skin of a rainbow trout medium with time for the five isolates were shown in (Yumoto et al. 2005). Molokwane et al. (2008) also Fig. 3. A decrease of the chromium concentration in a reported that a mixed culture of bacteria, containing solution was observed with time, with the maximum Oceanobacillus sp., could remove Cr(VI) under anaer- change generally observed within the initial 15 h for all obic condition. In this study, Oceanobacillus sp. W4 isolates. Among the five isolates, S. saromensis W5 had was observed to have the ability to remove chromate the highest chromium-removal rate of 0.79 mg h−1 mg−1 under aerobic conditions and could grow at a wider biomass, as the Cr(VI) concentration decreased from range of temperature and pH value than those previously 1,050 to 165 mg L−1 within the first 15 h. The concen- reported. tration of Cr(VI) decreased from 350 to 111.76 mg L−1 As described by Ilhan et al. (2004), the optimum (a rate of 0.21 mg h−1 mg−1 biomass) in incubations with temperature for chromium removal by a strain of P. saccharolyticum strain W1 within the first 15 h. The S. saprophyticus was 27 °C and the optimum pH value removal rate of Cr(VI) by S. saprophyticus strain W2 was found to be at 2.0. Mistry et al. (2010)reportedthe and Lysinibacillus sp. strain W3 was both about optimum pH value was 7.0 for the chromate removal of 0.07 mg h−1 mg−1 biomass within 15 h at the initial astrainofS. saprophyticus. In this study, Cr(VI) concentration of 175 mg L−1. According to Fein S. saprophyticus W2 could grow at a pH range of 8.0– et al. (2002), a strain of Sporosarc inaureae was capable 10.0 even though the 16S rRNA gene sequences share of removing Cr from a medium containing 5 mg L−1 99 % identity, which may indicate that strain W2 and the Cr(VI); calculation based on data got in the first 20 h strain of S. saprophyticus described in Ilhan et al. (2004) showed a removal rate of 1.15×10−5 mg h−1 mg−1 may belong to different subspecies. S. saromensis W5 biomass. S. saromensis W5 had much greater ability had the highest resistance to chromium among the five to remove Cr(VI) when compared with that. As de- isolates, as it reached the maximum cell density at scribed by Ilhan et al. (2004), the Cr(VI) bio-removing 1904, Page 8 of 10 Water Air Soil Pollut (2014) 225:1904 removal and EPS production was investigated, and the References 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 Badar, U., Ahmed, N., Beswick, A., Pattanapipitpaisal, P., & cell material. P. saccharolyticum strain W1 and Macaskie, L. (2000). Reduction of chromate by microorgan- S. saromensis W5, which had higher chromium- isms isolated from metal contaminated sites of Karachi, – removing ability as compared with the other isolates, Pakistan. Biotechnology Letters, 22(10), 829 836. Bafana, A. (2011). Mercury resistance in Sporosarcina sp. G3. produced much more EPS, with 178.60 and Biometals, 24(2), 301–309. − 189.19 mg g 1, respectively. S. saprophyticus strain Camargo, F., Okeke, F., & Frankenberger, B. (2003). Chromate W2 (82.15 mg g−1), Lysinibacillus sp. strain W3 reduction by chromium-resistant bacteria isolated from soils −1 contaminated with dichromate. Journal of Environmental (68.03 mg g ), and Oceanobacillus sp. W4 – −1 Quality, 32(4), 1228 1233. (140.40 mg g ) had lower EPS contents which Camargo, F. A. O., Okeke, B. C., Bento, F. M., & Frankenberger, corresponded to a relatively low chromium-removing W. T. (2005). Diversity of chromium-resistant bacteria iso- ability. The positive correlation is consistent with the lated from soils contaminated with dichromate. Applied Soil – report by Ozturk and Aslim (2008). As seen in Table 2, Ecology, 29(2), 193 202. Chen, J., Tang, Y.-Q., & Wu, X.-L. (2012). Bacterial community the contents of protein and carbohydrate of bacterial shift in two sectors of a tannery plant and its Cr (VI) remov- EPS were analyzed. The EPS of P. saccharolyticum ing potential. Geomicrobiology Journal, 29(3), 226–235. strain W1 and S. saromensis W5 had the highest carbo- Desai,C.,Parikh,R.Y.,Vaishnav,T.,Shouche,Y.S.,& hydrate contents when compared with the rest of the Madamwar, D. (2009). Tracking the influence of long-term chromium pollution on soil bacterial community structures isolates. Chromium was also found in the EPS with by comparative analyses of 16S rRNA gene phylotypes. −1 contents ranged from 20.7 to 5.90 mg g , which was Research In Microbiology, 160(1), 1–9. similar to the reports of Priester et al. (2006), Freire- Dubois, M., Gilles, K. A., Hamilton, J. K., Rebers, P., & Nordi et al. (2005), and Kiran and Kaushik (2008). Smith, F. (1956). Colorimetric method for determination of sugars and related substances. Analytical Additionally, it was found that more chromium was Chemistry, 28(3), 350–356. removed when more EPS was produced, suggesting that Fein, J. B., Fowle, D. A., Cahill, J., Kemner, K., Boyanov, M., & the bacterial EPS did contribute to the Cr(VI) removal. 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. 4 Conclusion Francisco, R., Alpoim, M., & Morais, P. (2002). Diversity of chromium resistant and reducing bacteria in a chromium Five Cr(VI)-removing bacterial strains were successful- contaminated activated sludge. Journal of Applied Microbiology, 92(5), 837–843. ly isolated from the Cr(VI)-contaminated soil samples Freire-Nordi, C. S., Vieira, A. A. H., & Nascimento, O. R. (2005). by using plating method, and these strains were found to The metal binding capacity of Anabaena spiroides extracel- grow at a wide range of initial Cr(VI) concentration. lular polysaccharide: an EPR study. Process Biochemistry, – Results showed that both P. saccharolyticum strain W1 40(6), 2215 2224. Gehrke, T., Telegdi, J., Thierry, D., & Sand, W. (1998). Importance and S. saromensis W5 had high chromium-removing of extracellular polymeric substances from Thiobacillus − − rates of up to 0.21 and 0.79 mg h 1 mg 1 biomass, ferrooxidans for bioleaching. Applied And Environmental respectively. To our knowledge, this is the first report Microbiology, 64(7), 2743–2747. on chromium removal by Pseudochrobactrum sp. A He, Z., Gao, F., Sha, T., Hu, Y., & He, C. (2009). Isolation and characterization of a Cr (VI)-reduction Ochrobactrum sp. positive relationship was found between bacterial strain CSCr-3 from chromium landfill. Journal of Cr(VI) removal and EPS production, suggesting that Hazardous Materials, 163(2–3), 869–873. EPS may be important in chromium removal; however, Ilhan, S., Nurbas, M., Kiliarslan, S., & Ozdag, H. (2004). Removal the exact mechanism of how the isolated strains are of chromium, lead and copper ions from industrial waste waters by Staphylococcus saprophyticus. Turkish Electronic involved in chromium removal needs further study. 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. Acknowledgment The authors would like to thank Dr. Allyson A. Jacobs, & C. P. Avakian (Eds.), Chromium (VI) Handbook Brady at the University of Calgary for her help in improving the (pp. 1–21). Boca Raton: CRC Press. paper. This work was financially supported by the National Nat- Kampfer, P., Rossello-Mora, R., Scholz, H. C., Welinder-Olsson, ural Science Foundation of China (no. 31370053). 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, Universita degli Studi di Firenze, Florence, Italy

Correspondence: Carlo Viti, Piazzale delle Abstract Cascine 24, 50144 Florence, Italy. Tel.: +39 0553288307; Hexavalent chromium [Cr(VI)] contamination is one of the main problems of fax: +39 0553288272; environmental protection because the Cr(VI) is a hazard to human health. The e-mail: carlo.viti@unifi.it 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), cross- Received 7 May 2013; revised 13 September ing the cellular membrane via the sulfate uptake pathway, generates active 2013; accepted 28 October 2013. Final intermediates Cr(V) and/or Cr(IV), free radicals, and Cr(III) as the final prod- version published online 3 December 2013. uct. Cr(III) affects DNA replication, causes mutagenesis, and alters the struc- DOI: 10.1111/1574-6976.12051 ture and activity of enzymes, reacting with their carboxyl and thiol groups. To persist in Cr(VI)-contaminated environments, microorganisms must have effi- Editor: Bernardo Gonzalez cient systems to neutralize the negative effects of this form of chromium. The systems involve detoxification or repair strategies such as Cr(VI) efflux pumps, Keywords Cr(VI) reduction to Cr(III), and activation of enzymes involved in the ROS Cr(VI) toxicity; chromate; dichromate; detoxifying processes, repair of DNA lesions, sulfur metabolism, and iron genomics; proteomics; transcriptomics. 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.

the hexavalent [Cr(VI)] and trivalent [Cr(III)] forms Introduction (Bartlett, 1991; Zayed & Terry, 2003). These oxidation Chromium, which belongs to the group VI-B transition states have different chemical features and affect organ- metals of the periodic table, has an atomic number of 24, isms in different ways. Cr(III) is conventionally consid- is the most abundant heavy metal, together with zinc, in ered an essential micro-nutrient in the diet of animals À the lithosphere (69 lgg 1; Li, 2000) and the 21st most and humans. Nevertheless, it has been reported recently abundant element in the Earth’s crust (ranging from 100 that chromium can no longer be considered an essential À to 300 lgg 1; Cervantes et al., 2001). This metal is element because rats on a diet with low-Cr(III) suffered introduced into the environment from natural sources no adverse consequences to body composition, glucose such as volcanic eruptions, forest ﬁres, and weathering, metabolism or insulin sensitivity compared with rats on a but the largest contribution to the deposition of chro- diet with a sufﬁcient dose of Cr(III) (Di Bona et al., mium in the biosphere is the result of anthropogenic 2011). On the other hand, a high dose (supra-nutritional activities. Chromium, due to its hardness, sheen, high level) of Cr(III) in the diet improved insulin sensitivity melting point, odorlessness, and anti-corrosiveness, is uti- (Di Bona et al., 2011). Cr(III) complexes accumulating in lized in various industrial activities, including electro- the body are potentially genotoxic (Levina et al., 2003) plating, steel, and automobile manufacturing, wood and therefore their use in micro-nutrient or antidiabetic treatment, leather tanning, pigments in dyes, paints, inks, treatments should be reconsidered after the accurate

MICROBIOLOGY REVIEWS MICROBIOLOGY plastics, and military defense applications (Langard, 1980; analysis of available and/or emerging data (Levina & Lay, James, 1996; Viti & Giovannetti, 2007). 2008; Di Bona et al., 2011). Cr(III) is relatively insoluble Chromium exists in different oxidation states but its under environmental conditions (Bartlett & Kimble, 1976; two most stable oxidation forms in the environment are Sass & Rai, 1987) and is considered less toxic than Cr(VI)

The reaction mechanism of YieF is different from that grown granular bacterial biofilms, found that there was not described for ChrR of P. putida and involves an obliga- reduction of Cr(VI) to Cr(III) under nonutrient condi- tory four-electron reduction of chromate in which the tions, whereas they efficiently reduced Cr(VI) from mini- enzyme simultaneously transfers three electrons to chro- mal media in the presence of acetate. In both studies, mate to produce Cr(III) and one electron to molecular X-ray absorption near-edge structure (XANES) spectros- oxygen, generating ROS (Ackerley et al., 2004b) (Fig. 3). copy and extended X-ray adsorption fine structure This reaction mechanism generates less ROS than that (EXAFS) were used, demonstrating that these approaches described for ChrR of P. putida, based on the combina- produce useful information about the speciation and asso- tion of two- and one-electron reduction, thus YieF should ciation of the Cr immobilized on microbial biomass. be a more suitable enzyme for chromate detoxification Very little is known about the mechanisms mediating than the P. putida ChrR (Ramirez-Diaz et al., 2008). Cr(VI) reduction in fungi. However, fungi have the Although a large number of studies have demonstrated ability to reduce Cr(VI), and many studies have been per- the role of ChrR in Cr(VI) reduction, proteomics revealed formed to exploit this capability for environmental bio- that this protein in P. putida F1, possessing a ChrR with remediation. Filamentous fungi, such as Aspergillus sp., a 100% amino acid identity to that of P. putida KT2404 Penicillium sp., and Trichoderma inhamatum, reduce Cr (Barak et al., 2006), was down-regulated in response (VI) to Cr(III) by exploiting the reducing power gener- to acute chromate exposure in all conditions tested ated by carbon metabolism as mechanism of Cr (Thompson et al., 2010). On the other hand, temporal (VI) detoxification (Acevedo-Aguilar et al., 2006; Mor- genomic and proteomic studies of S. oneidensis MR-1 ales-Barrera & Cristiani-Urbina, 2008). Paecilomyces lilaci- indicated that a NADPH-dependent FMN reductase nus has demonstrated the ability to both biotransform Cr [SO3585, incorrectly annotated as putative azoreductase (VI) and accumulate it in the biomass, exerting the maxi- (Mugerfeld et al., 2009)], sharing approximately 28% mum reduction activity during the log phase of growth, of identity with ChrR of P. putida, was significantly up- when cellular metabolic activity is maximized, and maxi- regulated in Cr(VI)-exposed cells (Brown et al., 2006; mum accumulation during the stationary phase (Sharma Thompson et al., 2007), especially at the highest chro- & Adholeya, 2011). Aspergillus niger strains have been mate doses used (Thompson et al., 2007). The deletion of described as coping with chromium mainly via the bio- the so3585 gene was not critical for cell survival in the sorption of the metal into the cells, rather than via the presence of chromate, and only an initial decrease of Cr use of reducing activity (Sandana Mala et al., 2006). The (VI) reduction rate was observed (Mugerfeld et al., 2009) Ed8 strain of Aspergillus tubingensis, included in the and therefore more studies are need to understand the A. niger species complex, demonstrated the ability to role of so3585 in chromate resistance. decrease Cr(VI) concentration in the medium via a Microbial respiration with Cr(VI) as the terminal elec- reduction mechanism stimulated by carboxylic acids and tron acceptor has never been rigorously shown (Richter metal-chelating agents (Coreno-Alonso et al., 2009). et al., 2012). Nevertheless, the global transcriptomic Extracellular reduction of Cr(VI) to Cr(III) was analysis of S. oneidensis MR-1, treated with 100 lMCr observed during the growth of Candida utilis by mecha- (VI) as the sole electron acceptor, revealed the up-regula- nisms independent from the intensity of culture growth tion of genes encoding MtrA, MtrB, MtrC, and OmcA or initial chromium concentration (Muter et al., 2001). (Bencheikh-Latmani et al., 2005), which are involved in On the basis of their results Muter et al. (2001) hypothe- the dissimilatory extracellular reduction of solid ferric sized that Cr(VI) reduction in C. utilis could be partly iron [Fe(III)] (hydr)oxides, uranium [U(VI)] and techne- dependent on pH changes of broth during the exponen- tium [Tc(VII)] (Belchik et al., 2011). The cytochromes tial phase or on exo-enzymatic activities during stationary MtrC and OmcA of S. oneidensis MR-1 were deeply char- phase. Candida maltosa, isolated from tanning liquors acterized to understand their role in Cr(VI) reduction. from a leather factory and characterized by a high toler- The data obtained supported the idea that MtrC and ance level of chromate in comparison with the yeast labo- OmcA are the terminal reductases of Cr(VI) in S. oneid- ratory strains C. albicans, S. cerevisiae, and Yarrowia ensis MR-1 (Belchik et al., 2011). lipolytica, demonstrated the ability to reduce Cr(VI) both Chromate reduction has been also associated with bio- in the presence of viable intact cells and in cell-free sorption. Fein et al. (2002) showed nonmetabolic reduc- extracts (Ramirez-Ramirez et al., 2004). This ability was tion of Cr(VI) to Cr(III) by bacterial surfaces under related to NADH-dependent chromate reductase activity nonutrient conditions as probable results of the oxidation associated with soluble proteins and, to a lesser extent, of organic molecules within the cell wall that serve as elec- with the membrane fraction (Ramirez-Ramirez et al., tron donors for Cr(VI) reduction to Cr(III). Nancharaiah 2004). Recently, the reduction of Cr(VI) to Cr(III) et al. (2010), studying Cr(VI) reduction by aerobically through an enzymatic mechanism has been observed in

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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 α polycrystalline -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.

I. INTRODUCTION barriers,29 which is often a less expensive and more effective 30−33 Chromium is a common contaminant in groundwater and is alternative to pumping treatments or bioremediation; released into the environment primarily through industrial however, those processes are currently limited by surface − 34 activities,1 6 such as through its use in wood preservatives, passivation of the iron-bearing solid phase. Reactive barriers used for remediating chromium contain iron-bearing materials refractory bricks, leather tanning processes, chromate plating, − such as magnetite,35 37 goethite,38 and iron sulfides,20 as well as and steel manufacturing. Chromium exists in the environment 20,39,40 7 zerovalent iron that are placed in the path of flowing in two stable oxidation states, namely Cr(III) and Cr(VI). 41−43 Hexavalent chromium, CrO 2−, is a highly toxic heavy metal ion groundwater. Contaminants are then removed at the 4 − species and a known carcinogen1,2,7 10 which is highly interface between the reactive barrier material and the − mobile11 14 in the environment. On the other hand, Cr(III), chromium-containing aqueous phase by in situ transformations, fi the reduced form of Cr(VI), is much less mobile in the including sorption and redox chemistry. Speci cally, the environment due to its propensity to form insoluble oxy- reduction of Cr(VI) by Fe(II) solutions, iron-bearing minerals, hydroxides and adsorb to mineral surfaces.7,15,16 Cr(III) is also or zerovalent iron allows Cr(VI) to be removed from an essential nutrient for humans and animals.1,2 Given that groundwater as the less mobile, less toxic Cr(III). Once generated, aqueous Cr(III) commonly adsorbs to iron oxides redox chemistry in the environment allows these two oxidation 15,44 states to easily interconvert, they are both considered and oxyhydroxides, forming strongly bound inner-sphere 45 ff potentially carcinogenic,1,17,18 placing chromium on the short complexes and/or precipitates that e ectively remove Cr(III) list of EPA priority pollutants.19 from the aqueous phase. While chromium remediation via fi Because Cr(III) is less mobile and less toxic than Cr(VI), permeable reactive barriers is widely used, a signi cant problem many Cr(VI) remediation techniques involve its reduction to Cr(III) through the use of readily available, inexpensive, and Received: December 13, 2012 − strong reducing agents, such as iron.17,20 28 One way to Revised: February 14, 2013 remediate redox-active contaminants is the use of reactive Published: February 18, 2013

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5171 dx.doi.org/10.1021/jp3122819 | J. Phys. Chem. C 2013, 117, 5164−5171

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

Contents lists available at SciVerse ScienceDirect

Journal of Hazardous Materials

jou rnal homepage: www.elsevier.com/locate/jhazmat

Cr(VI) reduction by a potent novel alkaliphilic halotolerant strain

Pseudochrobactrum saccharolyticum LY10

a a,∗ b a,a a

Dongyan Long , Xianjin Tang , Kuan Cai , Guangcun Chen , Linggui Chen ,

a c a

Dechao Duan , Jun Zhu , Yingxu Chen

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

Southern Research Outreach Center, University of Minnesota, 35838 120th Street, Waseca, MN 56093, USA

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 ver-

satilities 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 a b s t r a c t

Article history: A novel Cr(VI)-reducing strain, Pseudochrobactrum saccharolyticum LY10, was isolated and character-

Received 19 November 2012

ized for its high Cr(VI)-reducing ability. Strain LY10 had typical characteristics of alkali-tolerance and

Received in revised form 21 March 2013

halotolerance. Kinetic analysis indicated that the maximum reduction rate was achieved under opti-

Accepted 15 April 2013 −1 −1 9 −1

mum conditions with initial pH 8.3, 20 g L NaCl, 55 mg L Cr(VI), and 1.47 × 10 cells mL of cell

Available online 20 April 2013

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

Keywords:

the cells, with cell walls acting as the major binding sites for chromium. X-ray absorption near-edge

Bioremediation

structure (XANES) analysis further conﬁrmed that the chromium immobilized by the cells was in the

Cr(VI) reduction

Cr(III) state. In the present study, Pseudochrobactrum saccharolyticum was, for the ﬁrst time, reported to

Pseudochrobactrum saccharolyticum

Alkaliphilic be a Cr(VI)-reducing bacteria. Results from this research would provide a potential candidate for biore-

Halotolerant mediation of Cr(VI)-contaminated environments, especially alkaline and saline milieus with Cr(VI) at

1. Introduction

Chromium, a priority pollutant in the United States and many

other countries, has caused great public concern in recent years

Abbreviations: TEM, Transmission electron microscopy; EDS, Energy disper-

because of its wide usage, extensive distribution, and hazardous

sive X-ray spectroscopy; XANES, X-ray absorption near-edge structure; XAS, X-ray

potential [1,2]. However, the toxicity, solubility and bioavailabil-

absorption spectroscopy.

∗ ity of Cr, depend primarily on its chemical form [3]. Although

Corresponding author. Tel.: +86 571 8898 2013; fax: +86 571 8898 2010.

E-mail address: [email protected] (X. Tang). chromium can exist in a variety of valence states, Cr(VI) and Cr(III)

D. Long et al. / Journal of Hazardous Materials 256–257 (2013) 24–32 29

cell envelopes of strain LY10 treated with a low concentration of

−1

Cr(VI) (110 mg L ) (Fig. 7D).

The EDS spectrum showed an obvious signal of chromium pre-

cipitation (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 elec-

tron 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

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

Fig. 6. The Cr(VI) reduction percentage (A) and reduction rate (B) under optimal −1

−1 −1 to 220 mg L Cr(VI). As Cr(VI) has a characteristic sharp pre-edge

conditions. Experiments were conducted with initial pH 8.3, 20 g L NaCl, 55 mg L

9 −1 feature, which is absent in the Cr(III) spectrum, the deconvolution

Cr(VI) and 1.47 × 10 cells mL of cell concentration.

of sample XANES spectra with known Cr(VI) and Cr(III) standards

is regarded as a relatively straightforward approach for determin-

ing the Cr speciation [27]. As shown in Fig. 8, for the Cr(VI)-treated

−

1 cells, the typical strong pre-edge absorbance of Cr(VI) was absent

B). However, when exposed to Cr(VI) from 110 mg L (Fig. 7C

−

1 and a small peak was observed at 6009 eV, which was consistent

and D) to 220 mg L (Fig. 7E and F), the shapes of cells became

with results observed for model Cr(OH) compounds. These results

more and more irregular and a few of them even lost their shapes. 3

conﬁrmed that the majority of the Cr(VI) was reduced by P. saccha-

Moreover, a cluster of round globules encrusted on the surface of

−

1 rolyticum LY10, and the chromium immobilized by the cells was in

cells treated with 220 mg L Cr(VI) (red circle in Fig. 7F), while

the Cr(III) state.

no precipitation was observed in the control (Fig. 7B), nor on the

−1 −1

Fig. 7. TEM images of P. saccharolyticum LY10 cells not treated with Cr(VI) (A, B), incubated with 110 mg L Cr(VI) (C, D) and with 220 mg L Cr(VI) (E, F) for 96 h. Corresponding

−1

EDS spectra of untreated bacteria (a) and cells treated with 220 mg L Cr(VI) (b, c) for 96 h.

30 D. Long et al. / Journal of Hazardous Materials 256–257 (2013) 24–32

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

−1 −1

bioreduction by strain LY10 were pH 8.3, 20 g L of NaCl, 55 mg L

9 −1

of initial Cr(VI), and 1.47 × 10 cells mL of initial cell concen-

tration (Table S2). Complete Cr(VI) reduction and the maximum

−2 −1

reduction rate of (4.45 ± 0.11) × 10 h was achieved under

optimum conditions. The reduction rate of Cr(VI) removal by P.

−2 −1

saccharolyticum LY10 is high compared to (2.52 ± 0.33) × 10 h

−3 −1

reported for Sphaerotilus natans [29] and 3.50 × 10 h 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

Fig. 8. The Cr K-edge XANES spectra of P. saccharolyticum LY10 treated with

values is cumbersome, because kinetic models for Cr(VI) reduc-

−1

220 mg L Cr(VI) for 96 h and the reference compounds, K2Cr2O7 and Cr(OH)3.

tion by different bacteria are not identical, and reduction rates

are calculated in varied methods [14,15,29]. Moreover, there are

4. Discussion differences in experiment conditions (e.g. pH, Cr(VI) concentra-

tion, cell density, and electron-donors/electron-acceptors) [40].

Bioreduction of toxic Cr(VI) to stable Cr(III) by microorganisms Nonetheless, the speciﬁc reduction rates provided here are valuable

has been regarded as a promising approach for the remediation results. Such rate information will help in optimizing the oper-

of chromium contamination [2,30–32]. Since the ﬁrst microbial ation conditions for Cr(VI) bioremediation by P. saccharolyticum

Cr(VI) reducer was discovered by Romanenko and Korenkov [33], LY10.

the search for Cr(VI)-reducing bacteria has been enthusiastically To verify that P. saccharolyticum LY10 has a strong Cr(VI) reduc-

pursued, and various species have been isolated [34,35]. However, tive capacity rather than adsorption ability, the mechanism of

in this study, it is the ﬁrst time that P. saccharolyticum has been Cr(VI) reduction was further studied. Apart from the deleterious

clearly identiﬁed to be a Cr(VI) reducing bacteria. effects of toxic Cr(VI) on bacterial cells, which were also reported for

P. saccharolyticum LY10 demonstrated a typical characteristic A. haemolyticus [41], an obvious precipitation of chromium in and

of alkali-tolerance. It could efﬁciently reduce Cr(VI) under neutral- around the P. saccharolyticum LY10 cells was observed by TEM-EDS

alkaline conditions (pH 7.0–10.7). As Cr(VI) is known to desorb from analysis. Although chromium accumulation on the cell surface has

soil more rapidly at elevated pH levels [36], and Cr(III) is immobile been reported in many Cr(VI) reducing bacteria [18,22,41], quite

under alkaline pH conditions [35], the Cr(VI) reduction under alka- a few have been reported to have chromium precipitation both

line conditions would help to reduce the mobility and availability within and surrounding cells [17]. For P. saccharolyticum LY10, it

of the Cr ions in the environment [9]. The wide pH adaptability could accumulate chromium in the cells as well as on their sur-

and efﬁcient Cr(VI) reducing ability under neutral-alkaline condi- faces, with the cell wall acting as the main binding site. Since most

tion suggested that P. saccharolyticum LY10 could play an important Cr(VI) compounds are highly soluble, and the majority of Cr(III)

role in the bioremediation of alkaline Cr-polluted sites. compounds are relatively insoluble, it is tempting to speculate that

In addition, strain LY10 was able to adapt to high NaCl con- the chromium in and around the cells is in a reduced form [most

−1

centrations (1–20 g L ). According to the deﬁnition of Margesin likely Cr(III)] resulting from the Cr(VI) reducing process [17,42].

et al. [37], microorganisms requiring salt for growth are referred To further conﬁrm the bioreduction activity of P. saccharolyticum

as halophiles, and those able to grow in the absence as well as in LY10, speciation of cell-associated chromium was assessed with

the presence of salt, are designated as halotolerant. In the present XANES analysis. The bioreduction of Cr(VI) involves electron trans-

−1

study, strain LY10 could survive high salinity (60 g L NaCl), and fer processes, which can result in a direct metal speciation change

the presence of NaCl appeared to be a requirement for its effective [35,43]. In contrast, biosorption of metals is known to be controlled

Cr(VI) reducing activity, indicating that strain LY10 had a halo- by certain forces, such as the electrostatic force [44] and van der

tolerant nature [10,11,37]. The high salt tolerance of strain LY10 Waals [45]. Therefore, this process has no effect on the speciation

would enhance its actual performance in bioremediation applica- of metals [19]. In our study, the absence of the characteristic sharp

tions under salty conditions. pre-edge absorbance for Cr(VI) and the appearance of a small peak

Kinetic studies demonstrated that the initial stage (t < 72 h) of at 6009 eV conﬁrmed that the speciation of chromium had changed

Cr(VI) bioreduction by strain LY10 could be well described by the from the initial Cr(VI) to Cr(III) [2,46]. The results clearly indicated

ﬁrst-order kinetic model (Fig 2B, Fig 4B, Fig 5 and Fig 6). The that bioreduction process was involved in the Cr(VI) removal activ-

effects of different factors on Cr(VI) bioreduction were analyzed ity of P. saccharolyticum LY10.

by comparing the ﬁrst-order rate constants observed under varied Furthermore, recent studies on biosorption have demonstrated

conditions (Table S2). Results indicated that Cr(VI) reduction rates that the adsorption of Cr(VI) decreased with the increased pH,

varied with experiment conditions, including initial pH, salinity, especially for pH levels above 4.0 [44]. However, in this study,

biomass and Cr(VI) concentration. In addition, a close relationship the amount of Cr(VI) removed dramatically increased when pH

between microbial growth and Cr(VI) reduction rate was observed. increased from 5.5 to 8.3. On the contrary, great inhibition of Cr(VI)

For pH values ranging from 7.0 to 10.7 and NaCl concentrations reduction was observed when enzyme activity was suppressed

−1 −1

between 1 g L and 20 g L , strain LY10 grew well and rapid Cr(VI) under acidic conditions (pH 5.5). The preference for alkaline envi-

reduction rate were observed. However, when the cell growth was ronment rather than acidic condition further substantiated that the

−1

inhibited at pH 5.5 and the NaCl concentration of 0 g L , the Cr(VI) Cr(VI) removal by P. saccharolyticum LY10 was mainly achieved

reduction rate and reduction amount declined signiﬁcantly. It is by an enzyme-mediated bioreduction process, rather than by a

well known that the pH has a direct impact on the structure of surface-related bioadsorption activity [47].

D. Long et al. / Journal of Hazardous Materials 256–257 (2013) 24–32 31

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

Article history: The neptunyl ion is very toxic and has the potential to be highly mobile in the environment. In an effort to Received 10 September 2012 understand how its interactions with biological surfaces may affect its movement in the environment, we Received in revised form 2 January 2013 investigated neptunyl interactions with Bacillus subtilis endospores and their exudates. The exudates Accepted 6 January 2013 were dominated by dipicolinic acid. Spectrophotometric investigations of the chemical form of neptunyl Available online 20 January 2013 in exudate solutions are consistent with the formation of 1:1 neptunyl–dipicolinate complexes. Using – – Editor: Carla M. Koretsky neptunyl endospore adsorption data and spectrophotometric measurements of neptunyl dipicolinate complexes, we determined thermodynamic stability constants for both species. Neptunyl adsorption onto Keywords: the endospore surface decreased with an increasing pH, which corresponds to increasing aqueous complex- Neptunium ation of neptunyl by dipicolinate. Adsorption was also highly ionic strength dependent with adsorption Endospores increasing as ionic strength decreased. With stability constants determined in this work, we compared con- Bacillus subtilis trols on neptunyl partitioning in a simulated system with B. subtilis endospores, vegetative cells, and generic Surface complexation natural organic matter. Neptunyl complexation by B. subtilis endospore exudates exerted the greatest biolog- Dipicolinic acid ical control in the simulated systems. © 2013 Elsevier B.V. All rights reserved.

1. Introduction One instance of non-metabolizing bacterial reduction of Np(V) to Np(IV) has also been reported under low pH-high ionic strength con- Neptunium is a manmade element produced as a byproduct of ditions (Gorman-Lewis et al., 2005b). In addition to metabolic trans- nuclear fission in nuclear fuel. Np can be found as dissolved Np4+, formations of Np(V), direct Np(V)-vegetative bacterial cell surface + 2+ NpO2 , or NpO2 ions under aqueous conditions relevant to natural interactions have been investigated by batch adsorption, X-ray ab- + systems, though neptunyl(V), NpO2 , is predicted to be the most com- sorption spectroscopy, and surface complexation modeling (Sasaki mon species in natural environments (Fahey, 1986). As a monovalent et al., 2001; Songkasiri et al., 2002; Gorman-Lewis et al., 2005b). cation in aqueous solution, Np(V) typically forms less stable com- Previous Np(V)-microbial research focused on direct Np(V)–cell plexes than ions with higher charge (Keller, 1971). Due to its interactions with little attention to Np(V) reactions occurring with monovalency, Np(V) also is less likely to adsorb onto environmental microbial exudates. It is common for microbes to release exudates surfaces than Np(IV) and is predicted to be more mobile in the envi- with substantial complexation capabilities (Tourney et al., 2008; ronment (Lemire et al., 2001). However, bacteria can influence the Deo et al., 2010): extracellular polysaccharides (EPS) are one such mobility of Np in the environment (Law et al., 2010; Anderson et example. Microbial exudates are widely known to interact with min- al., 2011); thus, understanding the fate and transport of Np in the eral surfaces and dissolved species thus influencing the partitioning subsurface necessitates investigations of its interactions with biolog- of metals (Omoike and Chorover, 2004; Guibaud et al., 2005; Comte ical interfaces and their exudates. et al., 2006; Omoike and Chorover, 2006). While it has been observed Np-related geomicrobiological research has primarily focused on many times for other metal ions, the only study to investigate Np(V) interactions with microbes resulting in the reduction of Np(V) to interactions with microbial exudates, by Deo et al. (2010), found that Np(IV) or surficial interactions with vegetative microbial cells. Many Np(V) had a similar adsorption affinity for Shewanella alga EPS as for workers reported the occurrence of biologic reduction of Np(V) by S. alga whole cells and cell walls. Consequently, in that system S. alga metabolizing pure cultures and indigenous microbial communities exudates exert controls on Np(V) partitioning if the concentration of (Banaszak et al., 1998, 1999; Lloyd et al., 2000; Soderholm et al., the exudate (EPS) is comparable or greater than the effective concen- 2000; Rittmann et al., 2003; Icopini et al., 2007; Law et al., 2010). tration of the surface EPS/cell walls. The work of Deo et al. highlights the importance of considering not only the direct interactions of ⁎ Corresponding author. Np(V) with microbial cells but also interactions with microbial E-mail address: [email protected] (D. Gorman-Lewis). exudates.

0009-2541/$ – see front matter © 2013 Elsevier B.V. All rights reserved. http://dx.doi.org/10.1016/j.chemgeo.2013.01.004 78 D. Gorman-Lewis et al. / Chemical Geology 341 (2013) 75–83 ab 60 60 Adsorption Adsorption Reversibility Reversibility 50 Reversibility 50 Reversibility Model Model

40 40

30 30 % Np Adorbed

20 % Np Adorbed 20

10 10

0 0 2.5 3 3.5 4 4.5 5 5.5 6 6.5 7 1 2 3 4 5 6 7 pH pH

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 Np adsorption onto endospores is very different from adsorption exhibits similar behavior to low ionic strength systems. Initial adsorp- onto vegetative B. subtilis cells. Cation adsorption onto vegetative tion at pH 6.7 was 6% and subsequently lowering pH to 5.7, 4.5, and cells typically increases with increasing pH due to proton active func- 3.0 altered adsorption to 3, 7, and 10%, respectively. Experiments with tional groups on the bacterial surface deprotonating, which creates a + an initial pH of 2.5 and 15% adsorption experience a decrease in the more negatively charged surface. NpO2 exhibits this behavior in the amount of Np adsorbed down to 9, 4, and 3% by raising the pH to 3.3, presence of vegetative cells at low ionic strength (I=0.001 M) and 4.4, and 5.0, respectively. high ionic strength (I=0.1 M) above pH 4.5 (Gorman-Lewis et al., + Adsorption as a function of NpO2 concentration with a 10 g/L endo- 2005b). Np in contact with endospores exhibited the opposite behavior; + spore suspension is depicted in Fig. 2.TheNpO2 concentration to endo- adsorption decreased with increased pH at both low and high ionic + spore ratio was varied from 1.5 to 14. As NpO2 concentration increases, strength. As shown below, the observed Np–endospore adsorption + adsorption increases until a plateau appears at a NpO2 concentration of edge is due to competition between the endospore-surface complex − 43 μM(Fig. 2). and NpO2(dip) aqueous complex. Competition between complexation Ionic strength dependence of monovalent cation adsorption onto of cations by bacterial surfaces and aqueous organic ligands has also B. subtilis vegetative cells is a known phenomenon. Alessi et al. (2010) been observed by Song et al. (2009) who found Cd adsorption onto measured Rb+ and Li+ adsorption onto B. subtilis vegetative cells and Comamonas spp. inhibited by pthalic acid. found both ions to be weakly adsorbing and highly ionic strength Another substantial difference between B. subtilis endospores and + dependent, similar to NpO2 adsorption. However, a major difference vegetative cells in aqueous solutions is the ability of vegetative cells + + + + between monovalent cations Rb and Li and NpO2 adsorption is to promote non-metabolic reduction of NpO2 in low pH and high the shape of the adsorption edge as a function pH. Rb+ and Li+ adsorp- ionic strength solutions as reported by Gorman-Lewis et al. (2005b). tion increased as pH increased up to ca. 7 while we observed the oppo- B. subtilis vegetative cells also appear to be capable of non-metabolic + site behavior for NpO2 adsorption. 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 irre- 80 Data versible adsorption and an increase in adsorption with time under the Model conditions that promoted reduction. Additionally, Fein et al. (2002) 70 performed X-ray adsorption near edge spectroscopy to conﬁrm the 60 oxidation state of Cr on the bacterial surface. While the present work lacks spectroscopic conﬁrmation of the oxidation state of Np on the 50 endospore surface, the adsorption behavior (reversibility and rapid achievement of steady-state adsorption) exhibits none of the indicators 40 of reduction found in the previous work described above. Furthermore, 30 the optical spectroscopy of the supernate shows no evidence of Np(IV)

% Np Adorbed under the conditions investigated. The apparent inability of B. subtilis 20 endospores to promote reduction of Np(V) suggests some fundamental difference between endospore and vegetative cell surface reactivity 10 + toward NpO2 . 0 0 20 40 60 80 100 120 140 3.2. NpO+-exudate interactions Total Np (µM) 2

Fig. 2. Concentration isotherm at ca. pH 6 (exact pH values in Table S1) with a 10 g/L Endospores released dipicolinic acid in Np-bearing solutions endospore suspension. forming complexes that we investigated with optical spectroscopy. 82 D. Gorman-Lewis et al. / Chemical Geology 341 (2013) 75–83

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WIFAK BAHAFID, NEZHA TAHRI JOUTEY, HANANE SAYEL, MOHAMED IRAQUI-HOUSSAINI, and NAIMA¨ EL GHACHTOULI∗ Microbial Biotechnology Laboratory, Faculty of Sciences and Technology, Sidi Mohammed Ben Abdellah University, Fez, Morocco

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 the elimination of this metal from water and wastewaters is important to protect public health. The use of microbial biomass Heavy metal ions are extremely toxic and harmful even at low for the removal of toxic heavy metal ions from wastewaters concentrations, which can seriously affect plants and animals has emerged as an alternative to the existing methods which and have been involved in causing a large number of afflic- include chemical precipitation, ion exchange, membrane sepa- tions (Bulut and Tez 2007; Chua et al. 1999; Martins et al. ration, reverse osmosis, evaporation and electrochemical treat- 2006). Chromium heavy metal, is widely used in many impor- ment as a result of the search of lowcost, innovative methods tant industrial applications, such as steel production, electro- (Kapoor and Viraraghavan 1998; Rengeraj et al. 2001). plating, leather tanning, textile industries, wood preservation, Biosorption is energy independent binding of metal to the anodizing of aluminum, water-cooling and chromate prepa- cell wall of organism such as algae, fungi and bacteria, for ration (Garg et al. 2007), but is also one of the most toxic removal of metal ions (Gupta et al. 2001; Nourbakhsh et al. heavy metals (Silva et al. 2009), with high water solubility and 2002; Sag and Kutsal 2000). Particularly, fungal biomass can Downloaded by [University of Notre Dame] at 12:51 24 August 2014 mobility. be cheaply and easily procured in rather substantial quantities, It is generally accepted that Cr(III) species are not highly as a byproduct from established industrial fermentation pro- toxic and due to their limited solubility, they can be easily re- cesses. Furthermore, since such abundant dead fungal biomass moved by chemical and microbial approaches. It has also been is of little use, it has been identified as a potential source of suggested that this ion is involved in the tertiary structure of biomaterial for the removal of chromium from wastewaters. proteins and the conformation of cell RNA and DNA (Gulan Live or dead fungal cells can be used as an adsorbent ma- et al. 2001; Zayed and Terry 2003). In contrast, Cr(VI) is al- terial for the removal of toxic metal ions from aqueous so- ways toxic and exhibits mutagenic and carcinogenic effects on lutions (Sanghi et al. 2009), but non-living biomass appears biological systems (Codd et al. 2001; Costa 2003). Therefore, 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 The authors thankfully acknowledge the financial and scien- to metal toxicity, nutrient supply is not necessary and the tific support of Microbial Biotechnology Laboratory, Faculty of biosorbed metal ions can be easily desorbed and biomass can Sciences and Technology and of Regional Center of Interface be reused. It has also been reported that cell walls, consist- (CURI), SMBA University, Fez, Morocco. ing mainly of polysaccharides, proteins and lipids, offer many ∗Address correspondence to Na¨ıma El Ghachtouli, Microbial functional groups that can bind metal ions. In addition to Biotechnology Laboratory, SMBA University, Faculty of Sci- these functional binding groups, polysaccharides often have ences and Technology, Route Immouzer, P. O. Box 2202, Fez, ion exchange properties (Sag and Kutsal 1996; Veglio et al. Morocco; Email: [email protected] 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 Uptake (mg/g) % Inhibition

C. fabianii Live 0.48 ± 0.07 — Dead 0.37 ± 0.13 22% W. anomalus Live 0.47 ± 0.09 — Dead 0.3 ± 0.10 37% C. tropicalis Live 0.5 ± 0.06 — Dead 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 Downloaded by [University of Notre Dame] at 12:51 24 August 2014 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 Cardenas- Gonzalez and Acosta-Rodriguez (2010), who demonstrated Fig. 2. Effect of pH on percent removal of Cr(VI) by C. fabi- that the maximum uptake was observed at pH 4.0 (96% at anii HE650139 (a), W. anomalus HE648168 (b) and C. tropicalis 7 days) using respectively Paecilomyces sp. and Helmintospo- HE650140 (c). rium 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. 428 Bahafid et al.

as indicated by high maximum biosorption capacity qmax Cardenas-Gonzalez JF, Acosta-RodrıguezI. 2010. Hexavalent chromium (29.1 mg/g, 28.14 mg/g and 18.9 mg/g) by C. tropicalis removal by a Paecilomyces sp. fungal strain isolated from environ- HE650140, W. anomalus HE648168 and C. fabianii ment. Bioinorg Chem Appl 2010:1–6. HE650139, respectively. Chojnacka K. 2010. Biosorption and bioaccumulation–The prospects for practical applications. Environ Inter 36:299–307. The variation of the adsorption intensity (RL) with the ini- Chua WLO, Lam HKH, Bi SP. 1999. A comperative investigation on the tial concentration of the solution (C0 mg/L) was determined biosorption of lead by filamentous fungal biomass. Chemosphere (Table 3).From the result it appears that the RL value ap- 39:2723–2736. proaches zero with increase in the C0 value, which confirmed Cieslak-Golonka M. 1991. Chem Inform Abstract: Spectroscopy of that the three yeasts are a suitable biosorbent for adsorption chromium(VI) species. Coord Chem Rev 109:223–249. of chromium from wastewater under the conditions used in Codd R, Dillon CT, Levina A, Lay PA. 2001. Studies on the genotoxi- this study. The three yeasts were compared with other adsor- city of chromium: from the test tube to the cell. Coord Chem Rev (216–217):537–582. bents based on their maximum adsorption capacity for Cr(VI) Costa M. 2003. Potential hazards of hexavalent chromate in our drinking and shown in Table 4. It can be observed that the three yeasts water. Toxicol Appl Pharmacol 188:1–5. compares well with the other adsorbents listed in Table 4. 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 Conclusion clypeatus. Coll Surf B Biointer 60:46–54. Das SK, Guha AK. 2009. 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J Geomicrobiol 19:369–382. calis HE650140, the uptake of Cr(VI) was maximum at pH 3. Freundlich H. 1907. Ueber die adsorption in Loesungen. Z Phys Chem Linear Langmuir and Freundlich isotherm models were 57:385–470. used to represent the experimental data. Adsorption data fit- Garg UK, Kaur MP, Garg VK, Sud D. 2007. Removal of hexavalent ted well with the Langmuir and Freundlich models. How- chromium from aqueous solution by agricultural waste biomass. J ever, Freundlich isotherm displayed a better fitting model than Hazard Mater 140:60–68. Langmuir isotherm for W. anomalus HE648168 and C. tropi- Gulan ZV, Stehlik TV, Grba S, Lutilsky LD. 2001. Chromium uptake by Saccharomyces cerevisiae and isolation of glucose tolerance factor calis HE650140 because of the higher correlation coefficient, from yeast biomass. J Biosci 26:217–223. thus, indicating the applicability of monolayer coverage of the Gupta VK, Shrivastava AK, Jain N. 2001. Biosorption of chromium(VI) Cr(VI) ion on the surface of adsorbent. Based on the results from aqueous solutions by green algae Spirogyra species. Water Res of this research, the three yeasts can be considered as an effec- 35:4079–4085. tive, available, naturals and excellent adsorbents for removing Hall K, Eagleton L, Acrivos A, Vermeulen T. 1966. Pore and solid dif- chromium. fusion 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 As- pergillus niger: Effect of pretreatment. Biores Technol 63:109–113. References Khambhaty Y, Mody K, Basha S, Jha B. 2009. Biosorption of inor- Downloaded by [University of Notre Dame] at 12:51 24 August 2014 ganic mercury onto dead biomass of marine Aspergillus niger:Ki- Acosta RI, Rodriguez X, Gutiorrez C, Moctezuma MG. 2004. Biosorp- netic, equilibrium, and thermodynamic studies. 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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

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.

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

0 adsorption at low pH suggests that HgCl2 , 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

0 0 HgCl(OH) , and is limited by site availability and transformation to Hg(OH)2 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 proton- active 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

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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 1Institute of Biogeochemistry and Pollutant Dynamics, ETH Zurich, Zurich, Switzerland 2Dipartimento di Biotecnologie Agrarie, Sez. Microbiologia, Universita` degli Studi di Firenze, Florence, Italy

hexavalent chromium, Cr(VI), predominates (Katz and Salem − 2− Pseudomonas corrugata 28 represents a microorganism that can 1994). The anions HCrO4 and CrO4 are the prevailing potentially be applied for in situ bioremediation of Cr(VI) con- aqueous species of Cr(VI) at pH values below and above 6.5, re- taminated sites. This strain combines a high resistance toward spectively (Martell et al. 2004). In suboxic and anoxic systems, toxic Cr(VI) with the ability to reduce Cr(VI) to Cr(III) under Cr(III) is the dominant form of chromium. The cation Cr3+ oxic conditions. In this study, the aerobic reduction of Cr(VI) by Pseudomonas corrugata 28 was examined under different carbon hydrolyzes strongly and tends to form sparingly soluble oxides 3+ and sulfur supply conditions to assess the influence of microbial and hydroxides as e.g., Cr(OH)3 which limits Cr activities to carbon and sulfur metabolism on Cr(VI) reduction. The fate of very low values (≤10−9) at neutral pH (Baes and Mesmer 1976; reduced chromium was elucidated by investigating the speciation Rai et al. 1987). In addition, Cr3+ strongly sorbs to mineral of chromium in solution as well as the interaction of chromium phases such as e.g., iron (hydr-)oxides (Fischer et al. 2007) and with bacterial surfaces. Reduction of Cr(VI) was found to be a metabolic process resulting mainly in the formation of dissolved can substitute for Fe(III) in (hydr-)oxides (Frommer et al. 2010; organic Cr(III)-complexes. Small amounts of reduced chromium Frommer et al. 2009). Therefore, dissolved Cr(III) concentra- were weakly associated with bacterial surfaces. The formation of tions are typically very low in aquatic systems and soils. As inorganic Cr(III)-precipitates was not indicated. for the environmental concern of chromium, large differences between Cr(III) and Cr(VI) exist. Cr(III) represents an essential Keywords hexavalent chromium, microbial reduction, aerobic con- micronutrient. Cr(VI), however, is highly toxic and known to be ditions, organic metal complexes, scanning transmission carcinogenic and mutagenic (Katz and Salem 1994). As Cr(VI), X-ray microscopy (STXM), Pseudomonas corrugata 28 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 INTRODUCTION environment. Downloaded by [University of Notre Dame] at 12:53 24 August 2014 The inorganic speciation of chromium is strongly determined For the remediation of Cr(VI) contaminated sites, techni- by its redox properties. Under strongly oxidizing conditions, cal 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 Received 26 July 2010; accepted 1 November 2010. Cr(VI) by aerobic bacteria has been proposed as an alternative The X-ray microscopy studies were conducted at the STXM beam- remediation strategy for oxic environments (Cheng et al. 2010). lines X07D (PolLux), SLS, Paul Scherrer Institute, Villigen, Switzer- The capability of reducing Cr(VI) to Cr(III) has been docu- land and X-1A, NSLS, Brookhaven National Laboratory, Upton/NY, USA. The development of the X-ray microscopes at SLS and NSLS was mented for a large variety of microbial species (Ackerley et al. financially supported by BMBF, Germany under project 05 KS4 WE1/6 2004; Al Hasin et al. 2010; Bopp and Ehrlich 1988; Daulton as well as the Office of Biological and Environmental Research, U.S. et al. 2007; Fein et al. 2002; Fendorf et al. 2000; Lovley 1993; DoE under contract DE-FG02-89ER60858 and the NSF under grant Lovley and Phillips 1994; Opperman et al. 2008; Park et al. DBI-9605045, respectively. We are deeply grateful to George Tzvetkov 2000; Silver 1997; Suzuki et al. 1992; Viti et al. 2006; Viti et al. and Sue Wirick for their help at the beamlines. Address correspondence to Iso Christl, Institute of Biogeochemistry 2003). For an effective remediation of a highly contaminated and Pollutant Dynamics, ETH Zurich, CHN, 8092 Zurich,¨ Switzerland. site, however, bacteria applied to reduce Cr(VI) must exhibit E-mail: [email protected] a high resistance toward the toxicity of Cr(VI). As for many 173 174 I. CHRISTL ET AL.

other metals, bacterial resistance to Cr(VI) is plasmid-based in scriptional unit in P.corrugata 28 (Viti et al. 2009). Gene expres- most cases (Silver 1997). The ability to reduce Cr(VI) and to sion and Phenotype MicroArray analysis indicated that sulfur resist high Cr(VI) concentrations were found to be independent uptake was modulated when P. corrugata 28 was exposed to properties of bacteria (Bopp and Ehrlich 1988; Silver 1997). Cr(VI). However, it is still unknown to which extent changes in Bacteria may take up Cr(VI) primarily via the sulfate binding sulfate uptake of P. corrugata 28 affect its capability of reduc- protein (SBP) of their membranes because of the similarity of ing Cr(VI). In case of metabolic Cr(VI) reduction, a decrease both chromate and sulfate anions (Jacobson et al. 1991). In ad- of sulfate uptake may slow or stop metabolic processes, which dition to anion efflux-mediated resistance, the discrimination in turn might also decrease and slow down Cr(VI) reduction. between chromate and sulfate anions leading to a selective up- The ability to utilize sulfur sources other than sulfate may be take of sulfate as well as an active regulation of the sulfate a key feature to maintain metabolic processes in the presence uptake mechanism in the presence of chromate were discussed of high Cr(VI) concentrations. P. corrugata 28 is able to uti- as microbial responses to chromate exposure being linked to lize a large variety of sulfur sources including organosulfur chromate resistance (Nies 2003; Viti et al. 2009). Inside the compounds (Viti et al. 2009). However, the effect of utilizing cell, Cr(VI) may readily react with the DNA causing adverse organosulfur compounds for sulfur nutrition on Cr(VI) reduc- effects and with reduced compounds being associated with the tion has not been investigated. In addition, the fate of chromium membrane and present in the cytoplasm. after reduction by P. corrugata 28 has not been studied. Infor- Microbial Cr(VI) reduction counteracting adverse effects mation on the speciation of chromium after reduction is cru- has been reported to follow various mechanisms. Reduction cial to assess the environmental fate of the microbially reduced of Cr(VI) by a compound dissolved in the cytosol was shown chromium. for Bacillus sphaericus AND303 (Pal and Paul 2004). Pseu- The objectives of this study were (i) to investigate how vari- domonas fluorescens LB300 was found to reduce Cr(VI) en- ations of carbon and sulfur supply affect growth of P. cor- zymatically in the cell (Bopp and Ehrlich 1988). The bacterial rugata 28 and Cr(VI) reduction and (ii) to study the fate of species Bacillus subtilis, Sporosarcina ureae, and Shewanella chromium after aerobic microbial reduction. In the first part, we putrefaciens were shown to reduce Cr(VI) under aerobic condi- aimed at elucidating whether Cr(VI) reduction by P. corrugata tions at the outer surface of the cells in a nonmetabolic process 28 was controlled by metabolic or nonmetabolic processes. A (Fein et al. 2002). However, chemical oxidation of bacterial Tris minimal medium (TMM) containing varying concentra- surfaces by Cr(VI) leads most likely to an irreversible damage tions of gluconate (carbon source) was used for the experi- of the integrity and functionality of the cell membranes. Al- ments. Furthermore, sulfate and ethanesulfonate were used as ternative microbial Cr(VI) reduction mechanisms may include the sulfur sources to find out if the presence of Cr(VI) can electron transfer to Cr(VI) in the periplasm and the excretion induce sulfur starvation conditions. In the second part of this of Cr(VI) reducing agents. The reduction of Cr(VI) may not study, the formation of Cr(III) solid phases and Cr(III) ad- only represent a detoxification mechanism. Cr(VI) can also be sorption to bacterial surfaces were investigated. Complemen- used as a terminal electron acceptor as shown for Shewanella tary X-ray absorption microscopy analysis of bacterial cells oneidensis under anaerobic conditions (Daulton et al. 2007). was conducted at the C K-edge and the Cr L2,3-edge to fur- In addition, microbial Cr(VI) reduction is supposedly linked to ther elucidate the fate of microbially reduced chromium. For further microbial processes. For instance, reduction of Cr(VI) chromium speciation and mass balance, dissolved Cr(VI) and in the cytoplasm and the periplasm may condition a mechanism total dissolved chromium, which also includes Cr(III), were to export the sparingly soluble Cr(III) to avoid a detrimental measured. accumulation of Cr(III) in the cell (Bencheikh-Latmani et al. Downloaded by [University of Notre Dame] at 12:53 24 August 2014 2007). Among bacterial strains able to reduce Cr(VI), the strain MATERIALS AND METHODS Pseudomonas corrugata 28 was found to combine both a high Cr(VI) resistance and a high Cr(VI) reduction capability (Viti Materials et al. 2006) making strain 28 a valuable candidate for remedi- All chemicals mentioned in the following were analytical ation purposes. The minimum inhibitory concentration (MIC), grade unless stated otherwise. Solutions and standards were which indicates the lowest concentration of a substance that in- prepared with high purity water (>18 M cm, Milli-Q, Milli- hibits bacterial growth, was reported to be as high as 40 mM pore). All glass and plastic labware used was thoroughly washed Cr(VI) for P. corrugata 28. Growth of P. corrugata 28 and with ∼3% HCl and rinsed with high purity water prior to use. reduction of Cr(VI) can be decoupled as indicated by different growth and reduction rates when varying the carbon/energy Bacterial Strain and Cultivation source (Viti et al. 2007). Exposure of P. corrugata 28 to Cr(VI) The bacterial strain Pseudomonas corrugata 28 was used in was found to result in a strong and fast expression of the genes this study. P. corrugata 28 was isolated from a soil which was oscA (organosulfur compounds) and sbp (encoding sulfate ABC artificially polluted with up to 1000 mg Cr(VI) kg−1 soil (Viti transporter periplasmic sulfate binding protein) forming a tran- et al. 2006). The Gram-negative wild-type strain 28 was shown 182 I. CHRISTL ET AL.

would be transferred to chromium in case of complete Cr(VI) ical similarity of sulfate anions and chromate anions, chromate reduction. is transported into the cell via the sulfate transport system which The reduction of Cr(VI) was strongly dependent on both may not be able to discriminate both ions (Jacobson et al. 1991). carbon and sulfur supply (Figure 3). Small amounts of Cr(VI) Therefore, it is expected that the concomitant presence of sul- corresponding to less than 5% of total Cr(VI) added were re- fate and Cr(VI) would lead to sulfur starvation conditions and duced in the absence of gluconate and sulfur supply within increased uptake and reduction of Cr(VI). Our experimental 24 h implying that P. corrugata 28 may possibly slowly re- results, however, showed that P. corrugata 28 reduced a sim- duce Cr(VI) in a nonmetabolic way. Abiotic reduction of Cr(VI) ilar amount of Cr(VI) in case of sulfate and ethanesulfonate was quantitatively similar to reduction in experiments contain- supply when cells were exposed to 0.2 mM Cr(VI). A plau- ing bacteria but no gluconate or sulfur source. Consequently, sible explanation for our experimental results is that Cr(VI) the small decrease of Cr(VI) in the absence of carbon and reduction by P. corrugata 28 is primarily linked to the expres- sulfur supply most likely includes slow abiotic Cr(VI) re- sion of oscA involved in organosulfur cell homeostasis. Re- duction occurring in the experiments in addition to micro- duced organosulfur compounds can act as reductants for Cr(VI). bial reduction. Aerobic nonmetabolic Cr(VI) reduction at the Chromate is a strong oxidant having a high standard reduction outer surface of bacterial cells was shown for Bacillus sub- potential E0 of +0.55 V at pH 7, whereas reduced organosul- tilis, Sporosarcina ureae, and Shewanella putrefaciens (Fein fur compounds exhibit low standard reduction potentials, e.g., et al. 2002). Nonmetabolic reduction was faster for Sporosarcina E0 = –0.39 V for cysteine at pH 7 (Schwarzenbach et al. 2003). ureae and Shewanella putrefaciens as compared to Bacillus Immediate abiotic reduction of Cr(VI) by cysteine in standard subtilis and was found to strongly decrease with increas- TMM was also veriﬁed experimentally (not shown). We con- ing pH as shown for Bacillus subtilis (Fein et al. 2002). clude from our growth and reduction experiments that Cr(VI) At pH 7, only a small fraction of Cr(VI) was reduced in reduction of P. corrugata 28 is regulated by an effective intra- a nonmetabolic way despite high cell concentrations (∼7% cellular reduction mechanism rather than by a modulation of of 0.1 mM Cr(VI) within 4 h with 12 g bacteria L−1 and Cr(VI) transport coupled with extracellular or surface-mediated ∼1.5% of 0.06 mM Cr(VI) within 24 h with 1.2 g bacteria reduction. L−1). In our nutrient-absent experiments, the concentration of Reduction of Cr(VI) to Cr(III) strongly affects the aqueous P. corrugata 28 biomass was only ∼40 mg L−1, but cells were speciation of chromium as chromium is transformed from an an- 2− 3+ exposed to 0.2 mM Cr(VI) at pH 7. Assuming nonmetabolic ionic species (CrO4 ) into a cationic species (Cr ). Trivalent Cr(VI) reduction kinetics of P. corrugata 28 surfaces at pH 7 chromium strongly hydrolyzes and in contrast to Cr(VI), Cr(III) to be of similar magnitude as reported by Fein et al. (2002), forms sparingly soluble (hydr-)oxides such as e.g., Cr(OH)3 the extent of nonmetabolic Cr(VI) reduction is negligible in (Baes and Mesmer 1976; Martell et al. 2004; Rai et al. 1987). our growth and reduction experiments. Comparison of bacterial Therefore, the reduction of Cr(VI) to Cr(III) may potentially growth and Cr(VI) reduction results shows that Cr(VI) reduc- lead to a formation of Cr(III)-phases either in the cytoplasm or tion was clearly growth-phase dependent (Figure 4). This result in the extracellular space. The formation of biogenic minerals is consistent with previous investigations (Viti et al. 2007) and in the vicinity of bacterial surfaces and even inside cells has demonstrates that Cr(VI) reduction of P.corrugata 28 at pH 7 in been reported for various metal-microbe systems (Bazylinski TMM was clearly dominated by metabolic reduction. Further- and Moskowitz 1997; Fortin et al. 1997; Hunter et al. 2008). more, the marginal abiotic reduction of Cr(VI) in spent TMM Based on the hydrolysis of Cr(III), low concentrations of total containing excreted compounds (Figure 3d) and the similar ex- dissolved Cr(III) are expected in noncomplexing media at pH cretion of carbon compounds by P. corrugata 28 in the presence 7. Soluble Cr(III) was found to amount to 0.7 ± 0.5 10−6 Min Downloaded by [University of Notre Dame] at 12:53 24 August 2014 and absence of Cr(VI) (Figure 5b) suggest that excretion of or- 50 mM sodium nitrate solutions after 24 days (Figure 6) indi- ganic compounds in order to reduce Cr(VI) in the extracellular cating that nitrate was complexing Cr(III) only weakly. Sim- space may play –if any– a minor role for microbial reduction of ilarly, soluble Cr(III) concentrations are limited to 10−6–10−7 Cr(VI) by P. corrugata 28 in our experiments. M at pH 7 in other noncomplexing media such as perchlorate Comparison of microbial Cr(VI) reduction using different and chloride solutions (Martell et al. 2004; Rai et al. 1987). sulfur sources reveals that Cr(VI) reduction by P. corrugata 28 For TMM, the solubility of Cr(III) cannot be predicted reliably was similar, irrespectively whether sulfur was supplied as sulfate as stability constants for Cr(III)-complexation are not avail- or ethanesulfonate (Figures 3 and 4). It was shown previously able for all potential ligands initially present in the medium. that exposure of P. corrugata 28 to Cr(VI) caused a strong and In addition, the composition of the medium changed during fast expression of the genes oscA (organosulfur compounds) the bacterial growth as indicated by the decrease of gluconate and sbp forming a transcriptional unit (Viti et al. 2009). The (Figure 5a). Furthermore, the chemical structure and the con- expression of sbp may lead to a modulation of sulfate up- centration of excreted organic compounds, which also com- take since sbp encodes the periplasmic sulfate binding protein plex Cr(III), are unknown. The results of our solubility ex- (Sekowska et al. 2000; Viti et al. 2009). Due to the physicochem- periments carried out at pH 7 with TMM showed that the AEROBIC REDUCTION OF CHROMIUM(VI) 185

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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

metals from waste water through metabolically mediated The present study evaluates the biosorption of Cr(VI) 12 by Pseudomonas aeruginosa from synthetic solution or physico-chemical pathways of uptake . It is based on mechanisms such as complexation, ion exchange, adsorp- and tannery effluents. The absorption was studied 10 under different initial Cr(VI) concentrations at differ- tion, chelation and micro precipitation . Removal of ent pH values and in the presence of other metals. The heavy metals using different biosorbents has been found Cr(VI) concentration in the effluent, sludge and soil to be highly selective depending on the typical binding of tannery industries was measured. A maximum profile of the biosorbents13. Successful application of absorption was found at 30 mg/l of Cr(VI) at pH 8, biosorption depends on the parameters like initial metal which decreased in the presence of cadmium. Cyclic concentration and contact time14. voltammogram confirmed the reduction of Cr(VI). In this present study a Gram-negative, ubiquitous, FTIR analysis showed that the carboxyl and amino aerobic rod, Pseudomonas aeruginosa, has been used to groups on the bacterial surface bind chromium. SEM assess the removal of chromium with a view to provide and EDX revealed that Cr(VI) is reduced to Cr(III). environment-friendly methods of removal of toxic chromium Cr(VI) from industrial effluents. The cell wall of Keywords: Bioremediation, metal adsorption, Pseudo- this bacterium is composed of peptidoglycan, teichoic ac- monas aeruginosa, tannery effluents. ids along with carboxyl, phosphoryl, hydroxyl and amino functional groups at the surface15. Fourier transform CHROMIUM compounds are being widely used in leather infrared (FTIR) analysis was carried out to determine the tanning, steel production, as metal corrosion inhibitor, involvement of the type of functional groups in metal alloy formation, in paints as pigments and various other adsorption. Fein et al.16 reported that some components 1 applications . Chromium thus is a contaminant in the soil, of the cell wall serve as electron donors for the reduction sediment, surface and groundwater in the trivalent and reaction while metal binding16. 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 Materials and methods states of chromium, Cr(III) and Cr(VI) occur most commonly. Cr(VI) induces toxicity as it causes mutation4 and All the chemicals were either AR or GR grade. All-glass cancer5 in animals and mutation in bacteria. It is toxic triple-distilled water (TDW) was used throughout the even at a concentration of 20 μg/l (ref. 6). Heavy metal study. removal by chemical precipitation7, ion exchange, reverse osmosis and solvent extraction has disadvantages due to Biosorption studies high cost and energy of complex processes8. On the other hand, bioremediation has advantages like the possibility The parameters responsible for removal such as time of of metal recovery, easy waste disposal of the incineration contact, initial metal concentration, pH of culture media process and low cost. Metal uptake by microorganisms is and other interfering metal ions like cadmium and iron17 9 an environment-friendly alternative of heavy metal which are normally present in tannery effluents have been remediation. Both living and dead microbial mass are studied. This study has been performed both by supple- 10 capable of taking up metal ions from aqueous solution . menting synthetic solution of Cr(VI) and also treating Microorganisms take up metal ions either actively (bioac- Cr(VI) from the effluents of tannery industries in and 11 cumulation) and/or passively (biosorption) . Biosorption around Kolkata, India. is the ability of biological materials to accumulate heavy 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 agita- *For correspondence. (e-mail: [email protected]) tion at 37°C in the presence of Cr(VI) in LB broth

CURRENT SCIENCE, VOL. 101, NO. 5, 10 SEPTEMBER 2011 645 RESEARCH ARTICLES

13. Knauer, K., Behra, R. and Sigg, L., Adsorption and uptake of cop- 27. Kang, S. Y., Lee, J. U. and Kim, K. W., Biosorption of Cr(III) and per by the green alga Scenedesmus subspicatus (Chlorophyta). Cr(VI) on to the cell surface of Pseudomonas aeruginosa. Bio- J. Phycol., 1997, 33, 596–601. chem. Eng. J., 2007, 36, 54–58. 14. Ahalya, N., Kanamadi, R. D. and Ramachandra, T. V., Biosorp- 28. Park, D., Yun, Y. S. and Park, J. M., Studies on hexavalent chro- tion of chromium (VI) from aqueous solutions by the husk of mium biosorption by chemically treated biomass of Ecklonia sp. Bengal gram (Cicer arientinum). Electron. J. Biotechnol., 2005, 8, Chemosphere, 2005, 60, 1356–1364. 258–264. 29. Gonzalez, C. F., Ackerley, D. F., Park, C. H. and Matin, A., A 15. Beveridge, T. J., Role of cellular design in bacterial metal accu- soluble flavoprotein contributes to chromate reduction and toler- mulation and mineralization. Annu. Rev. Microbiol., 1989, 43, ance by Pseudomonas putida. Acta Biotechnol., 2003, 23, 233– 147–171. 239. 16. Fein, J. B., Fowle, D. A., Cahill, J., Kemner, K., Boyanov, M. and 30. Chai, L. Y., Huang, S. H., Yang, Z. H., Peng, B., Huang, Y. and Bunker, B., Nonmetabolic reduction of chromium (VI) by bacte- Chen, Y. H., Hexavalent chromium reduction by Pannonibacter rial surface under nutrient absent conditions. Geomicrobiol. J., phragmitetus BB isolated from soil under chromium-containing 2002, 19, 369–382. slag heap. J. Environ. Sci. Health A, 2009, 44, 615–622. 17. Deepali, G. K. K., Metals concentration in textile and tannery 31. Zakaria, Z. A., Zakaria, Z., Surif, S. and Ahmad, W. A., Hexava- effluents, associated soils and groundwater. N.Y. Sci. J., 2010, 3, lent chromium reduction by Acinetobacter haemolyticus isolated 82–89. from heavy metal contaminated waste water. J. Hazard. Mater., 18. Clesceri, L. S., Greenberg, A. E. and Eaton, A. D., Standard 2007, 146, 30–38. Methods for the Examination of Water and Wastewater, American 32. McLean, J. and Beveridge, T. J., Chromate reduction by a pseu- Public Health Association, Washington DC, 1998, 20th edn. domonad isolated from a site contaminated with chromate copper 19. Tunali, S., Kiran, I. and Akar, T., Chromium (VI) biosorption arsenate. Appl. Environ. Microbiol., 2001, 67, 1076–1084. characteristics of Neurospora crassa fungal biomass. Miner. Eng., 33. Chattopadhyay, B., Chatterjee, A. and Mukhopadhyay, S. K., Bio- 2005, 18, 681–689. accumulation of metals in the East Calcutta wetland ecosystem. 20. Talapatra, S. N. and Banerjee, S. K., Detection of micronucleus Aquat. Ecosyst. Health Manage., 2002, 5, 191–203. and abnormal nucleus in erythrocytes from the gill and kidney of 34. Bidwell, A. M. and Dowdy, R. H., Cadmium and zinc availability Labeo bata cultivated in sewage-fed fish farms. Food Chem. Toxi- to corn following termination of sewage sludge application. col., 2007, 45, 210–215. J. Environ. Qual., 1987, 16, 438–442. 21. Hamil, H. W., Williams, R. R. and Mackay, C., Principles of 35. DeFilippis, L. F. and Pallaghy, C. K., Heavy metals: sources and Physical Chemistry, Prentice Hall, New Jersey, 1966, 2nd edn. biological effects. In Advances in Limnology Series: Algae and 22. Smutok, O., Broda, D., Smutok, H., Dmytruk, K. and Gonchar, Water Pollution (eds. Rai, L. C. and Soeder, C. J.), E. Scheiz- M., Chromate-reducing activity of Hansenula polymorpha recom- bartsche Press, Stuttgart, 1994, pp. 31–77. binant cells over-producing flavocytochrome b2. Chemosphere, 36. Gowd, S. S. and Govil, P. K., Distribution of heavy metals in 2011, 83, 449–454. surface water of Ranipet industrial area in Tamil Nadu, India. 23. Deng, S. and Ting, Y. P., Characterization of PEI-modified bio- Environ. Monit. Assess., 2008, 136, 197–207. mass 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) ACKNOWLEDGEMENTS. We thank the Labonya Prova Bose Trust, biosorption by chemically modified biomass of Rhizopus nigri- Kolkata for providing a fellowship to S.C. I.G. thanks Netaji Subhas cans. Water. Res., 2002, 36, 1224–1236. Engineering College, Kolkata, for permission to continue his research 25. Kapoor, A. and Viraraghvan, T., Heavy metal biosorption sites in (Ph D) at Jadavpur University, Kolkata. Partial financial assistance Aspergillus niger. Bioresour. Technol., 1997, 61, 221–227. from the Centre of Advanced Studies (CAS), PURSE, DST, Depart- 26. Yee, N., Benning, L. G., Phoenix, V. R. and Ferris, F. G., Charac- ment of Chemistry, Jadavpur University is acknowledged. terization of metal–Cyanobacteria sorption reactions: a combined macroscopic and infrared spectroscopic investigation. Environ. Sci. Technol., 2004, 38, 775–782. Received 28 March 2011; revised accepted 10 August 2011

652 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

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

Article history: The adsorption of lead on Aspergillus versicolor biomass (AVB) has been investigated in aqueous solu- Received 16 July 2010 tion with special reference to binding mechanism in order to explore the possibilities of the biomass to Received in revised form address environmental pollution. AVB, being the most potent of all the fungal biomasses tested, has been 11 November 2010 successfully employed for reducing the lead content of the effluents of battery industries to permissi- Accepted 16 November 2010 ble limit (1.0 mg L−1) before discharging into waterbodies. The results establish that 1.0 g of the biomass Available online 24 November 2010 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 Keywords: Aspergillus versicolor within 3 h following pseudo second order rate model. The adsorption process can better be described by Lead Redlich–Peterson isotherm model compared to other ones tested. Scanning electron micrograph demon- Adsorption strates conspicuous changes in the surface morphology of the biomass as a result of lead adsorption. Binding mechanism Zeta potential values, chemical modification of the functional groups and Fourier transform infrared Chemical modification 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 In recent years considerable attention has been focused on adsorption technology to remove metal ions from wastewater from Lead is a highly toxic metal and its exposure even at low con- the standpoint of eco-friendly, effective and economic considera- centration leads to adverse effects on human health particularly in tions. Of the different adsorbents like saw dust, rice hulls, palm the nervous and reproductive system as well as in kidneys, liver and kernel husk, coconut husk, banana and orange peels, modified brain [1,2]. The metal enters into environment as a result of various lignin, de-oiled allspice husk and different agricultural by-products industrial activities such as electroplating, battery manufactur- [7,8]; activated carbon, is the best. However, its use is much ing, pigment and dye industries, lead smelting and using leaded restricted due to high cost in many countries including India. petroleum engine fuels [3,4]. Because of toxic nature, the concen- This leads to search for efficient adsorbents preferably biological tration of lead in the industrial effluents must be brought down materials covering microbial biomass to remove metal ions from to permissible limit (1.0 mg L−1) before discharging into public wastewater known as biosorption or bioaccumulation [9]. This sewers as per instruction of Environmental Regulatory Author- method has certain advantages over conventional ones, e.g., non ity, India. The available methodologies that are employed for the generation of toxic sludge, effective in reducing the concentration treatment of lead containing wastewater include precipitation with of metal ions below the permissible limit and the possibility of lime, ultrafiltration, reverse osmosis or ion exchange process [5,6]. recovery of the biomaterials by regeneration. Thus it provides an However, these methods suffer from limitations like poor cost economic means for the treatment of wastewater [7]. The uptake of effectivity, incomplete precipitation, etc. In addition they gener- heavy metals by microbial biomass is a two step process involving ate huge amount of toxic metal bearing sludge difficult to dispose initial cell surface binding followed by intracellular accumulation of. Thus research on the development of effective and inexpen- which takes place only in living cells [10]. Adsorption of metal ions sive technology to treat toxic metal bearing wastewater is very on the biomass occurs through electrostatic, physical and chemi- important. cal 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 ∗ Corresponding author at: Dept. of Biological Chemistry, Indian Association for are available on fungal systems [15,16]. Fungal biomass has cer- the Cultivation of Science, 2A & B, Raja S.C. Mullick Road, Jadavpur, Kolkata, West tain advantages over bacterial biomass in respect of processing and Bengal 700032, India. Tel.: +91 33 2473 4971/5904x502; fax: +91 33 2473 2805. E-mail addresses: [email protected], [email protected], handling of the biomass. The fungal genera of Rhizopus and Peni- [email protected], [email protected] (A.K. Guha). cillium [17,18] and the use of raw and pretreated Aspergillus niger

Table 3 [2] V.K. Gupta, A. Rastogi, Biosorption of lead from aqueous solutions by green Removal of Pb (II) from industrial effluent by AVB. algae Spirogyra species: kinetics and equilibrium studies, J. Hazard. Mater. 152 (2008) 407–414. Concentration of Pb (II) in Concentration of Pb (II) in [3] C. Raji, T.S. Anirudhan, Chromium (VI) adsorption by sawdust carbon: kinetics industrial effluent before industrial effluent after and equilibrium, Indian J. Chem. Technol. 4 (1997) 228–236. −1 −1 treatment (mg L ) treatment (mg L ) [4] A. Ornek, M. Ozacar, I.A. Sengil, Adsorption of lead onto formaldehyde or sul- phuric acid treated acorn waste: equilibrium and kinetic studies, Biochem. Eng. 3.20 0.45 J. 37 (2007) 192–200. 5.10 0.72 [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. 3.6. Reusability of AVB [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. Desorption of lead from the lead adsorbed biomass to the extent [7] A. Demirbas, Heavy metal adsorption onto agro-based waste materials: a review, J. Hazard. Mater. 157 (2008) 220–229. of 85% can be achieved with 0.1 M HCl and the biomass can again [8] J. Cruz-Olivares, C. Pérez-Alonso, C. Barrera-Díaz, G. López, P. Balderas- be used for adsorption of lead. The adsorption–desorption cycle ernández, Inside the removal of lead(II) from aqueous solutions by de-oiled can be conducted for five times after which loss in activity was allspice husk in batch and continuous processes, J. Hazard. Mater. 181 (2010) 1095–1101. noted probably due to loss of integrity of the cell (data not shown). [9] F. Veglio, F. Beolchini, Removal of metals by biosorption: a review, Hydromet- Thus we believe that adsorption involving AVB is a highly effec- allurgy 44 (1997) 301–316. tive and efficient methodology for removing lead from polluted [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. waterbodies. [11] P.G.G. Burnett, C.J. Daughney, D. Peak, Cd adsorption onto Anoxybacillus flavithermus: surface complexation modeling and spectroscopic investigations, 3.7. Interaction of lead with AVB in industrial effluent 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, The efficiency of AVB to remove lead was also executed using Geomicrobiol. J. 19 (2002) 369–382. the effluent of battery industry as a feed solution. The concentra- [13] A.Y. Dursun, A comparative study on determination of the equilibrium, kinetic −1 −1 and thermodynamic parameters of biosorption of copper (II) and lead (II) ions tion of lead in the effluent varied from 3.2 mg L to 5.1 mg L . The onto pretreated Aspergillus niger, Biochem. Eng. J. 28 (2006) 187–195. adsorption experiment was carried out after adjusting the pH at 5.0. [14] B. Volesky, Biosorption and me, Water Res. 41 (2007) 4017–4029. It was observed that the percentage removal of lead from indus- [15] J.N.L. Latha, K. Rashmi, M.P. Maruthi, Cell-wall-bound metal ions are not taken trial effluent was 86% (Table 3). This value is very similar to that up in Neurospora crassa, Can. J. Microbiol. 51 (2005) 1021–1026. [16] J. Wang, C. Chen, Biosorption of heavy metals by Saccharomyces cerevisiae:a obtained from the monometallic system, suggesting that the pres- review, Biotechnol. Adv. 24 (2006) 427–451. ence of anions and cations present in the industrial effluent has no [17] Y.S. Saˇg, T. Kutsal, Determination of the biosorption heats of heavy metal inhibitory effect on lead adsorption under the studied experimental ions on Zoogloea ramigera and Rhizopus arrhizus, Biochem. Eng. J. 6 (2000) 145–151. condition. [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: 4. Conclusion 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. We demonstrate a viable option for the removal of lead from [20] M. Amini, H. Younesi, N. Bahramifar, A.A.Z. Lorestani, F. Ghorbani, A. Daneshi, contaminated water with AVB. The maximum adsorption capac- M. Sharifzadeh, Application of response surface methodology for optimization of lead biosorption in an aqueous solution by Aspergillus niger, J. Hazard. Mater. ity of AVB has been found to be 45 mg Pb (II) per gram of the 154 (2008) 694–702. dry weight of the biomass. The Redlich–Peterson isotherm model [21] A. Kapoor, T. Viraraghavan, D.R. Cullimore, Removal of heavy metals using the describes the adsorption process satisfactorily suggesting that the fungus Aspergillus niger, Bioresour. Technol. 70 (1999) 95–104. [22] S.K. Das, A.K. Guha, Biosorption of chromium by Termitomyces clypeatus, Col- adsorption mechanism is a hybrid one and does not follow ideal loids Surf. B: Biointerfaces 60 (2007) 46–54. monolayer adsorption and the possibility of multilayer adsorption. [23] S.K. Das, A.K. Guha, Biosorption of hexavalent chromium by Termitomyces Scatchard plot analysis reveals multiple and non equivalent bind- clypeatus biomass: kinetics and transmission electron microscopic study, J. Hazard. Mater. 167 (2009) 685–691. ing sites on the AVB cell surface. The adsorption process is very [24] S. Majumdar, S.K. Das, T. Saha, G.C. Panda, T. Bandyopadhyou, A.K. Guha, fast initially and more than 80% is completed within 60 min. FTIR Adsorption behavior of copper ions on Mucor rouxii biomass through micro- study and chemical modifications of biomass cell surface suggest scopic and FTIR analysis, Colloids Surf. B: Biointerfaces 63 (2008) 138–145. the major involvement of carboxyl functional groups in the adsorp- [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. tion process. Thus it may be summarized that AVB can remove lead [26] G.M. Loudon, Organic Chemistry, First ed., Reading, MA, USA, Addison-Wesley, from its aqueous solution successfully. 1984. [27] A. Markowska, J. Olejnik, J. Michalski, Selektive Alkylierung von mehrbasigen Saüren des 4-bindigen Phosphors mit Trialkylphosphit, Chem. Ber. 108 (1975) Acknowledgements 2589–2592. [28] L.F. Fieser, M. Fieser, Reagents for Organic Synthesis, vol. 1, Wiley, New York, One of the authors (Mr. H. Bairagi) is thankful to the University 1967. [29] J. Gardea-Torresdey, M.K. Becker-Hapak, J.M. Hosea, D.W. Darnall, Effect of Grants Commission, New Delhi, India for awarding the Rajiv Gandhi chemical modification of algal carboxyl groups on metal ion binding, Environ. National Fellowship. The authors also gratefully acknowledge Mr. Sci. Technol. 24 (1990) 1372–1378. D. Halder (Dept. of Food Technology and Biochemical Engineering, [30] M. Sastri, A. Ahmad, M.I. Khan, R. Kumar, Biosynthesis of metal nanoparticles using fungi and actinomycete, Curr. Sci. 85 (2003) 162–170. Jadavpur University, Kolkata), Mr. S. Majhi and Mr. S. Naskar (Dept. [31] W. Lo, H. Chua, K.-H. Lam, S.-P. Bi, A comparative investigation on the biosorp- of Material Science, IACS, Kolkata), Dr. R. Chakravarty and Mr. A. tion of lead by filamentous fungal biomass, Chemosphere 39 (1999) 2723–2736. Ghosh (Dept. of Biological Chemistry, IACS, Kolkata), Dr. S.K. Das [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. (Dublin City University, Ireland) and Dr. A.R. Das (Polymer Science [33] P. Miretzky, M.C.B.M.F. Jardim, J.C. Rocha, Factors affecting Hg(II) adsorption in Unit, IACS, Kolkata) for their kind help throughout the work. 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. References [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. [1] M. Nadeem, A. Mahmood, S.A. Shahid, S.S. Shah, A.M. Khalid, G. Mckay, Sorption [36] B. Xiao, K.M. Thomas, Adsorption of aqueous metal ions on oxygen and of lead from aqueous solution by chemically modified carbon adsorbents, J. nitrogen functionalized nanoporous activated carbons, Langmuir 21 (2005) Hazard. Mater. B 138 (2006) 604–613. 3892–3902. Chemical Engineering Journal 162 (2010) 122–126

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Chemical Engineering Journal

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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 article info abstract

Article history: The immobilized fungus Coriolus versicolor was examined in a continuous fixed bed column for long-term Received 5 February 2010 Cr(VI) reduction at its physiological pH. The effects of operating parameters like flow rate, glucose con- Received in revised form 1 May 2010 centration in the influent feed, COD, initial Cr(VI) concentration on the Cr(VI) reduction were investigated. Accepted 10 May 2010 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) concen- Keywords: tration of up to 60 mg/L and thereafter showed a gradual decline. A Fourier transform infrared spectra Thioester were employed to elucidate the possible biosorption mechanism as well. The readiness of the thiol group Chromium reduction Fungal biomass of the fungal protein to interact with the Cr(VI) ion in addition to its strong reducing ability makes it a Column particularly important entity in the metabolism of Cr(VI). The possible role of thiol in the Cr(VI) reduction Immobilization 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 hydroxyl, sulphydryl, amino and phosphate groups of the microorganisms [7]. Chromium has been widely recognized as a toxic mutagen [1] The surfaces of fungal cells appear to act as ion exchange resins and a carcinogen yet is an important metal, which is used in a [8]. From the quantitative point of view, the surface sorption usually variety of industrial applications. Chromium is a metal that can can contribute the larger proportion to total metal uptake, and thus exist in oxidation states from −2 to +6, The trivalent oxidation binding to cell walls appears to be the most significant mechanism state is the most stable form of chromium In biological systems, of sorption. Since it is energy independent, it occurs in both living chromium is naturally found in its trivalent state at very vari- and dead microbial biomass, including fungal mycelium [9]. able levels, whereas the hexavalent form is generally a derivative The aim of the present work is to assess the long-term perfor- of man’s activities. Cr(VI) tends to associate with oxygen gener- mance of thiol containing live fungus Coriolus versicolor for the 2− ating the powerful oxidants chromate (CrO4 ) and dichromate continuous reduction of Cr(VI) in upflow fixed bed columns in 2− (Cr2O7 ). The biological effects of chromium are highly depen- a growth-supportive medium at physiological pH. The effects of dent on the oxidation state. Derivatives of Cr(III) are water insoluble some operating parameters, such as inlet Cr(VI) concentration, compared to Cr(VI) derivative compounds that are highly soluble media composition, and flow rate were examined and optimized [2]. for the long-term Cr(VI)-reduction performance in the column. In Requirements of large quantity of chemicals or energy can be this study we have used white-rot fungus C. versicolor as a biore- a limitation for the application of physicochemical methods for ductant since it has a high growth rate and can grow under a variety removing Cr(VI). Removal of heavy metal ions using biosorption of environmental conditions including low pH, high pollutant con- could be a promising technology and has received more and more centration. The possible mechanism for the reduction of Cr(VI) to attention in recent years [3,4,5]. Microorganisms, which are capa- Cr(III) by the fungus is also discussed. ble of transforming metals from one oxidation state to another, facilitate detoxification and/or the removal of chromium, and have 2. Materials and methods thus received recognition [6]. In the concept of biosorption, several chemical processes may be involved, such as bioaccumulation, 2.1. Reagents bioadsorption, precipitation by H2S production, ion exchange, and covalent binding with the biosorptive sites, including carboxyl, A stock solution of Cr(VI) was prepared containing 18.6736 g K2Cr2O7 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 Corresponding author. Fax: +91 512 2597866. E-mail address: [email protected] (R. Sanghi). AR grade.

2.2. Microorganism and media Table 1 Reduction of Cr(VI) ion at different flow rates (initial metal ion concentration 30 mg/L). A white-rot fungal strain, Coriolus versicolor, was obtained from Institute of Microbial Technology, Chandigarh, India. The strain was Flow rate (mL/h) mg Cr(VI) reduced/L h ◦ maintained at 4 C on malt agar slants. The liquid growth medium 20 11.51 ± 0.4 used for inoculating the fungus, consisted of 10 g/L glucose and 5 g/L 25 11.13 ± 0.5 malt extract (S.D Fine Chemicals, Mumbai). The growth medium 30 11.55 ± 0.5 35 11.72 ± 0.7 used was always autoclaved (WidWo Cat. AVD 500 horizontal auto- 40 8.88 ± 0.5 clave) at 15 psi for 30 min and cooled to room temperature before 45 6.67 ± 0.4 use. The inlet pH of feed containing Cr(VI) was 7.2. No adjust- 50 5.41 ± 0.2 ments 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, 2.6. Glucose analysis KH2PO4 2 g, MgSO4 1.023 g, CaCl2 0.1325 g and MnCl2 0.099 g. The determination of reducing sugar was done using 3,5- dinitrosalicylic acid method [11]. The effect of glucose concentra- 2.3. Immobilization of fungus in column tion on % Cr(VI) reduction was studied by varying the concentration from 0 to 5 g/L. Once the glucose concentration was optimized to The columns were made of borosilicate glass with 2.2 cm ID, 2 g/L, further experiments were carried out at this concentration 31 cm height, and 22.5 cm bed length. The glass column was packed only. The glucose consumed with time was also monitored and with ceramic beads on which the fungus was immobilized. To avoid equated with the Cr(VI) reduction with time. channel effects the Cr(VI) solutions (10–80 mg/L) were continuously pumped upward through the column by a peristaltic pump 3. Results and discussion (Mclin). The flow rates of the solutions were varied from 20 to 50 mL/h. The hydraulic residence time for the column reactor was 3.1. Monitoring of column conditions 54.8 min. Initially the columns were conditioned by recycling fungus- The columns were run in duplicates. The results revealed the containing growth media for a week followed by the feeding of presence of both Cr(VI) and Cr(III) in the effluent solution after sorp- fresh growth media until the COD reduction reached a steady tion of Cr(VI) on the biomass whereas on the fungal biomass, only state. Thereafter the influent growth medium was supplied with Cr(III) was found to be present. A decrease in pH from 7.2 to 3.2 known concentration of metal solution. The conditions for maxi- was observed during each Cr(VI) reduction cycle. In the first hour mum Cr(VI) reduction like influent Cr(VI) concentration, flow rate itself the pH reduced to 4.6 and thereafter the pH decrease was and glucose concentration in growth medium, were investigated slow. Reduction in pH of medium is possibly due to the accumula- and then subsequent analyses were performed under these opti- tion of organic acid metabolites [12]. With the decrease in solution mized condition. Column effluent samples were collected at regular pH, protonation of amine sites (NH ) of fungus increased favouring time intervals and COD, pH, Cr(total), Cr(VI), Cr(III), concentration 2 more electrostatic attraction of negative HCrO − ion yielding high were monitored. The COD measurements were made by closed 4 removal of Cr(VI). Perusals of the literature had also reported the reflux method according to the standard APHA procedure [12]. All reduction of Cr(VI) to Cr(III) at acidic pH along with protons been experiments were performed in duplicates at room temperature. consumed supported by Eq. (1) [13]: − + + + − → 3+ + E0 = . 2.4. Instrumental analysis HCrO4 7H 3e Cr 4H2O 1 33 V (1) Variation of flow rate though the column (shown in Table 1) Infrared (IR) spectra were recorded on a BRUCKER, VERTEX-70, revealed that maximum reduction occurs up to the range of and Infrared spectrophotometer making KBr pellets in reflectance 35 mL/h, above this flow rate chromate reduction decreases grad- mode. The pH of solutions was measured using a Digital pH-meter ually due to lesser contact time between the metal ions and the (MK VI Systronic). Spectrophotometric analysis was carried out biomass. on a Perkin Elmer, lambda-40 UV-VIS spectrophotometer with a The reduction rate of Cr(VI) was very fast initially; about 65% of 1 cm path length. For the SEM studies, samples of fungal biomass the starting Cr(VI) (30 mg/L) was reduced within the first 2.8 h of were coated under vacuum with a thin layer of gold and examined the reaction. However, the residual concentration of Cr(VI) reached by scanning electron microscopy [FEI (QANTA 200)] at 10–17.5 kV ◦ its minimum in 24 h. This rapid rate of Cr uptake by the immobilized with a tilt angle 45 . fungus has a significant practical importance for applications in small reactor volumes, thus ensuring efficiency as well as economy. 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 Glucose concentration was varied from 0 to 5 g/L and it was forms a red-violet complex selectively with Cr(VI). The color was observed that the reduction rate significantly increased from 0 to fully developed after 15 min and the sample solutions were anal- 2 g/L but became almost constant after 2 g/L. In the absence of ysed using the UV-VIS spectrophotometer and the absorbance glucose only sorption process takes place which results in poor of the color was measured at 540 nm. The total chromium con- reduction (Table 2). A glucose concentration of 2 g/L and above in centration was determined by oxidizing any trivalent chromium media resulted in the best Cr(VI) reduction. These results suggested with potassium permanganate, followed by analysis as hex- that glucose plays a vital role of carbon source in the reduction pro- avalent chromium. Cr(III) was determined from the difference cess. A comparison of quantitative uptake of glucose reveals that between total chromium and Cr(VI) concentrations. The instru- the rate and extent of metal uptake is significantly enhanced by the ment response was periodically checked with metal ion standard presence of glucose (Fig. 1). This enhanced metal removal capability solutions. may be related to an increase in the availability of energy and cel- 126 R. Sanghi, A. Srivastava / Chemical Engineering Journal 162 (2010) 122–126

4. Conclusions [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. Use of C. versicolor for the very long-term Cr(VI) reduction ability [5] A.Y. Dursun, C. Uslu, Y. Cuci, Z. Aksu, Bioaccumulation of copper(II), lead(II) in continuous upflow column represents a very potential successful and chromium(VI) by growing Aspergillus niger, Process Biochem. 38 (2003) strategy to bioremediate Cr(VI) toxic wastewaters in their natural 1647–1651. [6] B. Volesky, Z.R. Holan, Biosorption of heavy metal, Biotechnol. Prog. 11 (1995) habitat with extremes of weather conditions. Due to the surface 235–250. immobilization of fungal hyphal biomass, the solute could easily [7] B. Krantz-Rulcker, B. Allard, J. Schunrer, Interactions between a soil fungus, Tri- pass through the highly porous matrix of ceramic beads for reach- choderma harzianum, and IIb metals—adsorption to mycelium and production ing the functional groups of the biomass. The main purpose of this of complexing metabolites, Biometals 6 (1993) 223–230. [8] B. Krantz-Rulcker, E. Frandberg, J. Schunrer, Metal loading and enzymatic paper is to establish the optimization parameters for the reduction degradation of fungal cell walls and chitin, Biometals 8 (1995) 12–18. of Cr(VI) with C. versicolor such that the process is sustainable for a [9] N. Saglam, R. Say, A. Denizli, S. Patir, M.Y. Arica, Biosorption of inorganic mer- very long term. The long-term potential of the column packed with cury and alkylmercury species on to Phanerochaete chrysosporium mycelia, Process Biochem. 34 (1999) 725–730. fungal biomass for Cr(VI) detoxification was demonstrated very [10] American Public Health Association, Standard Methods for the Analysis of effectively. The fungus clearly is able to thrive, grow and success- Water and Wastewater, American Public Health Association, Washington, DC, fully reduce toxic Cr(VI) to its less toxic form even under repeated 1998. [11] G.L. Miller, Use of dinitrosalicylic acid reagent for determination of reducing exposure of Cr(VI) for more than a year. The results indicated that sugar, Anal. Chem. 31 (1959) 426–428. the packed columns with the immobilized living cells is more con- [12] J.T. Spadaro, M.H. Gold, V. Renganathan, Degradation of azo dyes by the lignin- venient in operation and economic in treatment compared with degrading fungus Phanerochaete chrysosporium, Appl. Environ. Microbiol. 58 (1992) 2397–2401. traditional methods as it could resolve the problem of blockage as [13] J.B. Fein, D.A. Fowle, J. Cahill, K. Kemner, M. Boyanov, B. Bunker, Nonmetabolic well as recovery of the metal ions from wastewater. 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 chemi- Acknowledgments cally modified biomass of Rhizopus nigricans, Water Res. 36 (2002) 1224–1236. [15] D.W.J. Kwong, D.E. Pennington, Stoichiometry, kinetics and mechanism of The authors are thankful to International Foundation of Science chromium (VI) oxidation of l-cysteine at neutral pH, Inorg. Chem. 23 (1984) 2528–2532. (Sweden) for the financial support to carry out this work. [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. References [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 [1] J.E. Gruber, K.W. Jennette, Metabolism of the carcinogen chromate by rat liver biotechnology in industry, agriculture and medicine, in: M. Moo-Young (Ed.), microsomes, Biochem. Biophys. Res. Commun. 82 (1978) 700–770. Comprehensive Biotechnology, Pergamon Press, 1983, p. 235. [2] E.W. Cary, Chromium in air, soil and natural waters, in: S. Langard (Ed.), Bio- [19] P.H. Connett, K.E. Wetterhahn, Metabolism of the carcinogen chromate by cel- logical and Environmental Aspects of Chromium, Elsevier, Amsterdam, 1982, lular constituents, Struct. Bonding 54 (1983) 93–125. pp. 49–64. [20] W. Mazijrek, P.J. Nichols, B.O. West, Chromium (VI)-thioester formation in N,N- [3] J. Wang, Biosorption of copper (II) by chemically modified biomass of Saccha- dimethylformamide, Polyhedron 10 (1991) 753–762. romyces cerevisiae, Process Biochem. 3 (2002) 847–850. APPLIED AND ENVIRONMENTAL MICROBIOLOGY, Apr. 2010, p. 2433–2438 Vol. 76, No. 8 0099-2240/10/$12.00 doi:10.1128/AEM.02792-09 Copyright © 2010, American Society for Microbiology. All Rights Reserved.

Immobilization of Cr(VI) and Its Reduction to Cr(III) Phosphate by Granular Bioﬁlms Comprising a Mixture of Microbesᰔ 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

Received 18 November 2009/Accepted 8 February 2010 Downloaded from

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 day؊1 g؊1, with reduction of 0.17 mM ,day؊1 g؊1 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 http://aem.asm.org/ 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.

Chromium is a common industrial chemical used in tanning aqueous environments. Bioflocs, the active biomass of acti- leather, plating chrome, and manufacturing steel. The two vated sludge-process systems are transformed into dense gran- stable environmental forms are hexavalent chromium [Cr(VI)] ular biofilms in sequencing batch reactors (SBRs). As granular and trivalent chromium [Cr(III)] (20). The former is highly biofilms settle extremely well, the treated effluent is separated on August 24, 2014 by UNIV OF NOTRE DAME soluble and toxic to microorganisms, plants, and animals, en- quickly from the granular biomass by sedimentation (9, 24). tailing mutagenic and carcinogenic effects (6, 22, 33), while the Previous work demonstrated that aerobic granular biofilms latter is considered to be less soluble and less toxic. Therefore, possess tremendous ability for biosorption, removing zinc, cop- the reduction of Cr(VI) to Cr(III) constitutes a potential de- per, nickel, cadmium, and uranium (19, 26, 31, 32, 40). How- toxification process that might be achieved chemically or bio- ever, no study has investigated the role of cellular metabolism logically. Microbial reduction of Cr(VI) seemingly is ubiqui- of aerobically grown granular biofilms in metal removal exper- tous; Cr(VI)-reducing bacteria have been isolated from both iments. Despite vast knowledge about biotransformation by Cr(VI)-contaminated and -uncontaminated environments (6, pure cultures, very little is known about reduction and immo- 7, 23, 38, 39). Many archaeal/eubacterial genera, common to bilization by mixed bacterial consortia (8, 12, 13, 16, 20, 31, 36). different environments, reduce a wide range of metals, includ- Our research explored, for the first time, the metabolically ing Cr(VI) (6, 16, 21). Some bacterial enzymes generate Cr(V) driven removal of Cr(VI) by microbial granules. by mediating one-electron transfer to Cr(VI) (1, 4), while The main aim of this study was to investigate Cr(VI) reduc- many other chromate reductases convert Cr(VI) to Cr(III) in tion and immobilization by mixed bacterial consortia, viz., aer- a single step. obically grown granular biofilms. Such biofilm-based systems Biological treatment of Cr(VI)-contaminated wastewater are promising for developing compact bioreactors for the rapid may be difficult because the metal’s toxicity potentially can kill biodegradation of environmental contaminants (17, 24, 29). the bacteria. Accordingly, to protect the cells, cell immobiliza- Accordingly, we investigated the microbial reduction of Cr(VI) tion techniques were employed (31). Cells in a biofilm exhibit by aerobically grown biofilms in batch and fed-batch experi- enhanced resistance and tolerance to toxic metals compared ments and analyzed the oxidation state and association of the with free-living ones (15). Therefore, biofilm-based reduction chromium immobilized on the biofilms by X-ray absorption of Cr(VI) and its subsequent immobilization might be a satis- near edge spectroscopy (XANES) and extended X-ray absorp- factory method of bioremediation because (i) the biofilm- tion fine structure (EXAFS). bound cells can tolerate higher concentrations of Cr(VI) than planktonic cells, and (ii) they allow easy separation of the MATERIALS AND METHODS treated liquid from the biomass. Ferris et al. (11) described Cultivation of aerobic granular sludge. Aerobic granular biofilms were grown microbial biofilms as natural metal-immobilizing matrices in 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. * Corresponding author. Mailing address: Brookhaven National The reactor was operated at room temperature (30 Ϯ 2°C) at a volumetric Laboratory, Environmental Sciences Department, Building 490A, Up- exchange ratio of 66% and a 6-h cycle, comprising 60 min of anaerobic static fill, ton, NY 11973. Phone: (631) 344-4534. Fax: (631) 344-7303. E-mail: 282 min of aeration, 3 min of settling, 10 min of effluent decantation, and 5 min [email protected]. of being idle. The SBR was fed with acetate-containing synthetic wastewater as ᰔ Published ahead of print on 19 February 2010. discussed by Nancharaiah et al. (27). Granules, collected 2 months after the

2433 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 NR(Å) ␴2 ⌬E0 F Cr(III) phosphate Cr-O 5.7 Ϯ 0.7 1.97 Ϯ 0.01 0.002 Ϯ 0.001 0.5 Ϯ 1.0 0.038 Cr-P 4.0 Ϯ 1.3 3.11 Ϯ 0.05 0.007 Ϯ 0.002 10.5 Ϯ 1.4

Bacterial granules Cr-O 6.3 Ϯ 1.5 1.98 Ϯ 0.04 0.001 Ϯ 0.001 1.2 Ϯ 0.8 0.017 Cr-P 4.0 Ϯ 1.3 3.11 Ϯ 0.03 0.007 Ϯ 0.002 13.3 Ϯ 2.5

a N, coordination number (number of atoms); R, interatomic distances; ␴2, disorder parameter; ⌬E0, energy shift; F, goodness-of-fit parameter. Downloaded from mass. The microbial species composition of the granular sludge Office of Science, U.S. Department of Energy under contract no. was not identified in the present study; nonetheless, the com- DE-AC02-98CH10886. Y.V.N. gratefully acknowledges the American Society for Microbiology for the Indo-US Visiting Research Profes- mencement of reduction of chromium immediately after expo- sorship Award. sure to Cr(VI) suggests that bacteria able to reduce chromium We thank Avril D. Woodhead for editorial help. already were present in the granules (without prior enrichment); the lack of a delay demonstrates that the necessary REFERENCES enzymes are constitutively expressed. Seemingly, previous ex- 1. Ackerley, D. F., C. F. Gonzalez, M. Keyhan, R. Blake II, and A. Matin. 2004. http://aem.asm.org/ posure to chromium and subsequent microbial enrichment are Mechanism of chromate reduction by the Escherichia coli protein, NfsA, and not prerequisites for successful bioreduction. This could be the role of different chromate reductases in minimizing oxidative stress during chromate reduction. Environ. Microbiol. 6:851–860. mainly due to the involvement of constitutive chromate reduc- 2. Al Hasin, A., S. J. Gurman, L. M. Murphy, A. Perry, T. J. Smith, and P. H. E. tases, thus corroborating the earlier observation of the rapid Gardiner. 2010. Remediation of chromium(VI) by a methane-oxidising bacterium. Environ. Sci. Technol. 44:400–405. reduction of Cr(VI) by Pseudomonas putida unsaturated bio- 3. APHA. 1995. Standard methods for the examination of water and wastewa- films (32). ter, 19th ed. American Public Health Association, Washington, DC. Aerobic granular sludge cultivated in an SBR using acetate 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 reduc- and lacking prior exposure to chromium efficiently reduced tases and generation of an improved enzyme by directed evolution. Appl. on August 24, 2014 by UNIV OF NOTRE DAME Cr(VI) from minimal media. Passive biosorption by the gran- Environ. Microbiol. 72:7074–7082. ular biomass was ruled out because Cr(VI) removal was neg- 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 ligible in the absence of a carbon source and by lyophilized reactor. Water Res. 33:2283–2290. granules. Analysis of chromium speciation by XANES further 6. Cervantes, C., J. Campos-García, S. Devars, F. Gutie´rrez-Corona, H. Loza- Tavera, J. C. Torres-Guzmań, and R. Moreno-Sańchez. 2001. Interactions of confirmed the bioreduction of Cr(VI) to Cr(III), thereby chromium with microorganisms and plants. FEMS Microbiol. Rev. 25:335– pointing to the involvement of cell metabolism. Nonmetabolic 347. reduction of Cr(VI) to Cr(III) by bacterial surfaces under 7. Chardin, B., M.-T. Giudici-Orticoni, G. De Luca, B. Guigliarelli, and M. Bruschi. 2003. Hydrogenases in sulfate-reducing bacteria function as chro- nonnutrient conditions has been reported by Fein et al. (10). In mium reductase. Appl. Microbiol. Biotechnol. 63:315–321. this study, no such reduction of Cr(VI) to Cr(III) was observed 8. Chung, J., R. Nerenberg, and B. E. Rittmann. 2006. Bioreduction of soluble under nonnutrient conditions. EXAFS analyses revealed that chromate using a hydrogen-based membrane biofilm reactor. Water Res. 40:1634–1642. the granular biofilm-bound Cr(III) occurs as Cr(III) phos- 9. de Kreuk, M. K., J. J. Heijnen, and M. C. M. van Loosdrecht. 2005. Simul- phate. Earlier, Neal et al. (28) reported that only Cr(III) was taneous COD, nitrogen, and phosphate by aerobic granular sludge. 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A recent study showed reduction of Cr(VI) to Cr(III) by meth- 13. Francis, A. J. 2007. Microbial mobilization and immobilization of plutonium. ane-oxidizing bacteria, a ubiquitous group of environmental J. Alloys Compd. 444–445:500–505. bacteria (2). EXAFS analysis showed that Methylococcus cap- 14. Ganguli, A., and A. K. Tripathi. 2002. Bioremediation of toxic chromium from electroplating effluent by chromate-reducing Pseudomonas aeruginosa sulatus-associated chromium predominantly existed as Cr(III) A2Chr in two bioreactors. Appl. Microbiol. Biotechnol. 58:416–420. and most likely associated with phosphate groups. EXAFS 15. Harrison, J. J., H. Ceri, and R. J. Turner. 2007. Multimetal resistance and spectra of our Cr(III)-laden granular biomass revealed the 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 presence of Cr(III)-phosphate after Cr(VI) reduction. Overall, 2006, posting date. Low temperature reduction of hexavalent chromium by a our findings suggest the potential use of mixed microbial gran- microbial enrichment consortium and a novel strain of Arthrobacter aurescens. BMC Microbiol. 6:5. doi:10.1186/1471-2180-6-5. ules to bioremediate Cr(VI)-containing wastewater or indus- 17. Inizan, M., A. Freval, J. Cigana, and J. Meinhold. 2005. Aerobic granulation trial effluents. 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. Sholto- ACKNOWLEDGMENTS Douglas, 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 This research was supported by the Department of Atomic Energy, analysis. Science 306:686–687. Government of India, and in part by the Environmental Remediation 19. Liu, Y., S. F. Yang, S.-F. Tan, Y.-M. Lin, and J.-H. Tay. 2002. 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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 . Chromiumreduction . Electrondonors . Growthkinetics

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, Achro- mobacter 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],andreductionusingironparticles [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

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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 article info abstract

Article history: Biosorption of Cr+6 by Termitomyces clypeatus has been investigated involving kinetics, transmission elec- Received 1 August 2008 tron microscopy (TEM) and Fourier transform infrared spectroscopic (FTIR) studies. Kinetics experiments Received in revised form 8 January 2009 reveal that the uptake of chromium by live cell involves initial rapid surface binding followed by relatively Accepted 8 January 2009 slow intracellular accumulation. Of the different chromate analogues tested, only sulfate ion reduces the Available online 19 January 2009 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 Keywords: and sodium azide inhibit chromate accumulation by ∼30% in live cell. This indicates that accumulation of Termitomyces clypeatus Biosorption chromium into the cytoplasm occurs through the active transport system. TEM-EDXA analysis reveals that Cr(VI) the chromium localizes in the cell wall and also in the cytoplasm. Reduction of chromate ions takes place Intracellular accumulation by chromate reductase activity of cell-free extracts of T. clypeatus. FTIR study indicates that chromate ions Transmission electron microscopy (TEM) accumulate into the cytoplasm and then reduced to less toxic Cr+3 compounds. © 2009 Elsevier B.V. All rights reserved.

1. Introduction into the cytoplasm through specific carrier system. The transport process in prokaryotic organisms has been studied in some details Chromium, a toxic heavy metal, dissipates into the environment [18–22]. The state of art in the field of biosorption of heavy metals as a result of various industrial activities [1,2]. In view of toxicity has recently been reviewed by Volesky [23].However,onlyafew and related environmental hazards [3], it is essential that the con- reports are available on fungal systems [24,25]. Fungal biomass has centration of chromium in the effluent must be brought down to certain advantage over bacterial biomass in this natural ‘ecofriendly permissible limit [4] before discharging into water bodies. Among green technological process’ in respect of processing and handling different available technologies [5,6] the removal of metal ions of the biomass. Further, in comparison to bacteria, fungi are known from wastewater by adsorption on biological materials specially to secret much higher amount of exopolymers, thereby significantly microbial biomass known as biosorption/bioaccumulation [7–10] increasing the productivity of biosorption/bioremediation process has recently gained much importance. This method does not gen- [26]. In this manuscript we describe the biosorption/or bioaccumu- erate toxic sludge, capable of reducing the concentration of metal lation mechanism of chromium on Termitomyces clypeatus biomass ions below the permissible limit and the possibility of regeneration (TCB) from kinetics study in presence of different co-ions and of the materials and thus provide an effective and economic means metabolic inhibitors with support from Fourier transform infrared for the remediation of heavy metal polluted wastewater [11–14]. spectroscopy and transmission electron microscopic investigations. The uptake of heavy metals by microbial biomass is essentially a biphasic process consisting of metabolism independent initial 2. Materials and methods cell surface binding that can occur either in living or inactivated organisms, followed by energy dependent intracellular accumula- 2.1. Chemicals tion which takes place only in the living cells [15]. The cell wall materials are involved in the initial surface binding of metal ions Dehydrated microbiological media and ingredients were pro- though electrostatic, physical and/or chemical interaction [16,17]. cured from Himedia, India. All other reagents were of analytical In living cells besides surface adsorption, metal ions may enter 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], A stock solution of chromium (100 mg/l) was prepared by dis- [email protected] (A.K. Guha). solving potassium dichromate (K2Cr2O7) in double distilled water

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(5). Nonetheless, large-scale industrial use of hexavalent Remediation of Chromium(VI) by a chromium continues around the world and continues to be Methane-Oxidizing Bacterium an environmental threat (6, 7). Chromate(VI) reductase activity, capable of reducing chromium from the +6tothe+3 oxidation state, has † ABUBAKR AL HASIN, previously been characterized in aerobic bacteria and ‡ STEPHEN J. GURMAN, facultative anaerobes, such as Escherichia coli, Pseudomonas § † LORETTA M. MURPHY, ASHLEE PERRY, putida (8-11), Paracoccus denitrificans (12), and Bacillus THOMAS J. SMITH,*,† AND † subtilis (13), as well as in anaerobic sulfate-reducing bacteria PHILIP H. E. GARDINER (14, 15). Here, we have investigated the chromium(VI) Biomedical Research Centre, Sheffield Hallam University, reduction in methane-oxidizing bacteria, a ubiquitous group Howard Street, Sheffield S1 1WB, United Kingdom, of environmental bacteria (16), in which to our knowledge Department of Physics and Astronomy, University of this reaction has not previously been investigated. Leicester, University Road, Leicester LE1 7RH, Methane-oxidizing bacteria (or methanotrophs) are de- United Kingdom, and School of Chemistry, Bangor fined by their ability to use methane as their sole carbon and University, Bangor LL57 2UW, United Kingdom energy source. The best characterized examples belong the R and γ subdivisions of the proteobacteria, although recent Received June 11, 2009. Revised manuscript received results have indicated that aerobic methanotrophs are November 13, 2009. Accepted November 20, 2009. 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); methan- Methane-oxidizing bacteria are ubiquitous in the environment otrophs of the R-proteobacteria (type II methanotrophs) and are globally important in oxidizing the potent greenhouse include Methylosinus trichosporium OB3b (16). Methan- gas methane. It is also well recognized that they have wide otrophs are well recognized as environmentally significant potential for bioremediation of organic and chlorinated organisms. Mole-for-mole environmental methane (pro- organic pollutants, thanks to the wide substrate ranges of the duced by anaerobic breakdown of organic matter) is 21 times methane monooxygenase enzymes that they produce. Here more powerful as a greenhouse gas than carbon dioxide (22), we have demonstrated that the well characterized model and so oxidation of methane by methanotrophs is important methanotroph Methylococcus capsulatus (Bath) is able to in controlling global warming (23). In addition to their ability bioremediate chromium(VI) pollution over a wide range of to oxidize their growth substrate methane, the methane concentrations (1.4-1000 mg L-1 of Cr6+), thus extending the monooxygenase enzymes produced by methanotrophic bacteria can also co-oxidize diverse hydrocarbons and bioremediation potential of this major group of microorganisms halogenated organic compounds, including aromatics and to include an important heavy-metal pollutant. The chromium(VI) the priority pollutant trichloroethylene, and so application reduction reaction was dependent on the availability of of methanotrophs in bioremediation of such compounds reducing equivalents from the growth substrate methane and has been widely investigated (24-26). Here we have inves- was partially inhibited by the metabolic poison sodium tigated the interaction of chromate (VI) with well character- azide. X-ray spectroscopy showed that the cell-associated ized representatives of the two major groups of methane- chromium was predominantly in the +3 oxidation state and oxidizing bacteria and have identified a strain that is able to associated with cell- or medium-derived moieties that were most bioremediate this major heavy-metal pollutant. likely phosphate groups. The genome sequence of Mc. capsulatus (Bath) suggests at least five candidate genes for Materials and Methods the chromium(VI) reductase activity in this organism. Bacterial Strains and Growth Conditions. The methanotrophs Ms. trichosporium OB3b and Mc. capsulatus (Bath) were obtained from the culture collection of H. Dalton and Introduction J. C. Murrell (University of Warwick, U.K.) and were grown and propagated aerobically in nitrate mineral salts (NMS) Microbiologically catalyzed reduction reactions offer a pos- - medium or on NMS agar (27) containing 1 mg L 1 of sible solution to environmental pollution with chromate(VI), CuSO · 5H O using methane (1:4 v/v in air) as the source of which is a highly oxidizing, soluble, mutagenic, and toxic 4 2 carbon and energy. All cultures and chromate-reduction form of the metal that is produced as an effluent from metal experiments were incubated at the optimal growth temper- plating, tanning, paper making, and other industries (1, 2). ature of the organism concerned, 30 and 45 °C, respectively, Reduction of chromium(VI) to the +3 oxidation state for Ms. trichosporium OB3b and Mc. capsulatus (Bath). Except produces a form of the metal that is less toxic, less bioavailable where stated otherwise, chromate reduction experiments and more able to adsorb to negatively charged biopolymers were performed in 50 mL liquid cultures in 250-mL conical and soil particles (1-3). Governments have sought to restrict Quickfit flasks sealed with a Subaseal (Fisher) to prevent loss its use and release; for instance, the European Union has of methane, while allowing addition of liquids and the taking restricted the use of chromium in electrical equipment (4) of samples using hypodermic syringes. Cultures were allowed and exposure limits have been implemented to protect to grow to an OD of 0.3-0.8 before addition of potassium individuals from harm due to chromate (VI) contamination 600 chromate (VI) or potassium dichromate (VI) to give the concentration of hexavalent chromium stated for each + + * Corresponding author phone 44 114 225 3042; fax 44 114 225 experiment. Fermentor cultivation of Mc. capsulatus (Bath) 3066; e-mail [email protected]. † Sheffield Hallam University. was performed using methane as the source of carbon and ‡ University of Leicester. energy according to the published method (28), in a Bioflo § Bangor University. 110 fermentor (New Brunswick Scientific, vessel capacity 4.5

400 9 ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 44, NO. 1, 2010 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 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 FIGURE 4. EXAFS spectroscopy of the particulate fraction from a consistent with the formation of insoluble chromium(III) Mc. capsulatus culture treated with 1000 mg L-1 of hexavalent phosphate in the phosphate-containing growth medium. chromium for 96 h: (a) isolated EXAFS oscillations and (b) Fourier Indeed, the electron-dense particles seen in via transmission transform EXAFS. Experimental data are shown as open diamonds electron microscopy of cells from cultures exposed to 500 - joined with dotted lines and fits as solid lines. mg L 1 of chromium(VI) may be particles of such chromium(III) phosphate. Phosphorus/oxygen coordination en- precipitated from NMS medium when chromium(III) was vironments would also be produced from association of (a added to a concentration of 1000 mg L-1, so we suggest that proportion of) the chromium(III) with phosphate-containing this sample contains a suspension of chromium phosphate. cellular components such as nucleotide coenzymes (11) and There was no sign of a diminution of the chromium DNA, where phosphorus/oxygen coordination of chro- fluorescence signal during the six hours of data taking, mium(III) may be important in the mutagenic properties of suggesting that no settling is occurring: the suspension must chromate (VI) (37). The apparently uniform staining of the therefore be very finely divided. cells with electron dense material after exposure to the lower - The environment of chromium in the cells is identical to concentration of chromate (VI) of 100 mg L 1 suggests that in the solution containing phosphate. Again the inclusion association of the chromium with some form of cellular of a second shell significantly lowers the fit index. The results material. imply that chromium in the cells is in the Cr(III) state (as The effect of the metabolic inhibitor sodium azide on the shown by the Cr-O distance) and probably in a suspension chromate (VI) reduction reaction, and also the fact that the (as shown by the second shell contribution). This suspension reaction is not shown by autoclaved (dead) cells, supports is most probably insoluble Cr(PO4). To within our rather large the conclusion that the reaction is dependent on active uncertainties there is no change in the chromium environ- cellular metabolism rather than merely a reaction between ment with time of exposure of cells to chromium(VI) between the chromate and cellular constituents as described by Fein 24 and 96 h nor with the temperature of EXAFS data et al. (38). The effect of azide is presumably an indirect one acquisition (Supporting Information Table 1). This suggests since inhibition of the reduction of dioxygen to water would that chromium in these cells is reduced to Cr(III) and not per se prevent channelling of reducing equivalents into precipitated out within 24 h. reduction of chromate. The absolute dependence of chromate Genome of Mc. capsulatus Contains Candidate Chro- reductase activity at 1.4 mg L-1 of chromium(VI) on the mate Reductase Genes. To identify genes that could encode presence of methane very strongly suggests that the growth proteins involved in reduction of chromate by Mc. capsulatus, substrate methane is required to supply electrons for the database of translated open reading frames from the reduction of chromate. While there is currently no clear complete genome sequence (i.e., all the known and potential evidence to show which cellular components are responsible proteins of the organism) was searched for homologues to for reduction of chromium(VI), sequence similarity searches the known classes of chromate reductases from other using known chromate-reducing enzyme sequences indicate bacteria. BLAST searches indicated the presence of three at least five possible candidates for the chromium-reducing homologues of the E. coli Fre chromate reductase (11; enzyme in Mc. capsulatus. One of these is the well charac- accession no. M74448) in the Mc. capsulatus (Bath) genome terized reductase component of soluble methane monooxy- all of which are known or likely flavin/Fe2S2 oxidoreductases: genase, a component of one of the enzyme systems that is (1) a protein annotated as a putative oxygenase (accession involved in oxidation of methane to methanol (26, 39). It is no. YP_114919, E ) 5 × 10-12), (2) the reductase component unlikely that this particular candidate is involved in reduction of sMMO (MmoC, E ) 4 × 10-10), and (3) a protein annotated of chromate in these experiments because the cells were as a putative Na+-translocating NADH-quinone reductase cultivated under conditions of high copper-to-biomass ratio, subunit (accession no. YP_114800, E ) 3 × 10-9). The E. coli when the soluble methane monooxygenase is not expressed nitroreductase NfsA, which also reduces chromate (8), has (40). Future availability of the genome sequence for Ms.

VOL. 44, NO. 1, 2010 / ENVIRONMENTAL SCIENCE & TECHNOLOGY 9 403 (29) McLean, J.; Beveridge, T. J. Chromate reduction by a (37) Zhitkovich, A.; Song, Y.; Quievryn, G.; Voitkun, V. Non-oxidative pseudomonad isolated from a site contaminated with chro- mechanisms are responsible for the induction of mutagenesis mated copper arsenate. Appl. Environ. Microbiol. 2001, 67, by reduction of Cr(VI) with cysteine: role of ternary DNA adducts 1076–1084. in Cr(III)-dependent mutagenesis. Biochemistry 2001, 40, 549– (30) Gurman, S. J.; Binsted, N.; Ross, I. A rapid, exact curved-wave theory 560. for EXAFS calculations. J. Phys. (Paris) 1984, C17, 143–151. (38) Fein, J. B.; Fowle, D. A.; Cahill, J.; Kemner, K.; Boyanov, M.; (31) Joyner, R. W.; Martin, K. J.; Meehan, P. Some applications of Bunker, B. Nonmetabolic reduction of Cr(VI) by bacterial statistical tests in analysis of EXAFS and SEXAFS data. J. Phys. surfaces under nutrient-absent conditions. Geomicrobiol. J. (Paris) 1987, C20, 4005–4012. 2002, 19, 369–382. (32) Bajt, S.; Clark, S. B.; Sutton, S. R.; Rivers, M. L.; Smith, J. V. (39) Lund, J.; Woodland, M. P.; Dalton, H. Electron transfer reactions Synchrotron X-ray microprobe determination of chromate in the soluble methane monooxygenase of Methylococcus content using X-ray-absorption near-edge structure. Anal. Chem. capsulatus (Bath). Eur. J. Biochem. 1985, 147, 297–305. 1993, 65, 1800–1804. (40) Stanley, S. H.; Prior, S. D.; Leak, D. J.; Dalton, H. Copper stress (33) M. V. Aldrich, Gardea-Torresdey, J. L.; Peralta-Videa, J. R.; underlies the fundamental change in intracellular location of Parsons, J. G. Uptake and reduction of Cr(VI) to Cr(III) by methane monooxygenase in methane-oxidizing organismss mesquite (Prosopis spp.): chromate-plant interaction in hy- Studies in batch and continuous cultures. Biotechnol. Lett. 1983, droponics and solid media studied using XAS. Environ. Sci. 5, 487–492. Technol. 2003, 37, 1859–1864. (41) Kim, H. J.; Graham, D. W.; DiSpirito, A. A.; Alterman, M. A.; (34) Opperman, D. J.; Piater, L. A.; van Heerden, E. A novel chromate Galeva, N.; Larive, C. K.; Asunskis, D.; Sherwood, P. M. A. reductase from Thermus scotoductus SA-01 related to Old Methanobactin, A copper-acquisition compound from methane- Yellow Enzyme. J. Bacteriol. 2008, 190, 3076–3082. oxidizing bacteria. Science 2004, 305, 1612–1615. (35) Cervantes, C.; Ohtake, H.; Chu, L.; Misra, T. K.; Silver, S. Cloning, (42) Choi, D. W.; Do, Y. S.; Zea, C. J.; McEllistrem, M. T.; Lee, nucleotide sequence, and expression of the chromate resistance S.-W.; Semrau, J. D.; Pohl, N. L.; Kisting, C. J.; Scardino, L. L.; determinant of Pseudomonas aeruginosa plasmid pUM505. J. Hartsel, S. C.; Boyd, E. S.; Geesey, G. G.; Riedel, T. P.; Shafe, Bacteriol. 1990, 172, 287–291. P. H.; Kranski, K. A.; Tritsch, J. R.; Antholine, W. E.; DiSpirito, (36) Ward, N.; Larsen, O.; Sakwa, J.; Bruseth, L.; Khouri, H.; Durkin, A. A. Spectral and thermodynamic properties of Ag(I), Au(III), A. S.; Dimitrov, D.; Jiang, L.; Scanlan, D.; Kang, K. H.; Lewis, M.; Cd(II), Co(II), Fe(III), Hg(II), Mn(II), Ni(II), Pb(II), U(IV), and Nelson, K. E.; Methe´, B.; Wu, M.; Heidelberg, J. F.; Paulsen, I. T.; Zn(II) binding by methanobactin from Methylosinus tricho- Fouts, D.; Ravel, J.; Tettelin, H.; Ren, Q.; Read, T.; DeBoy, R. T.; sporium OB3b. J. Inorg. Biochem. 2006, 100, 2150–2161. Seshadri, R.; Salzberg, S. L.; Jensen, H. B.; Birkeland, N. K.; (43) Jenkins, M. B.; Chen, J.-H.; Kadner, D. J.; Lion, L. W. Nelson, W. C.; Dodson, R.; Grindhaug, S. H.; Holt, I.; Eidhammer, Methanotrophic bacteria and facilitated transport of pol- I.; Jonasen, I.; Vanaken, S.; Utterback, T.; Feldblyum, T. V.; Fraser, lutants in aquifer material. Appl. Environ. Microbiol. 1994, 60, C. M.; Lillehaug, J. R.; Eisen, J. A. Genomic insights into 3491–3498. methanotrophy: The complete genome sequence of Methylo- coccus capsulatus (Bath). PLoS Biol. 2004, 2, 1616–1628. ES901723C

VOL. 44, NO. 1, 2010 / ENVIRONMENTAL SCIENCE & TECHNOLOGY 9 405 Journal of Basic Microbiology 2008, 48, 135 – 139 135

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 Department of Environmental Sciences, Chatrapati Shahu Ji Maharaj University, Kalyanpur – Kanpur – 208024 UP, India 2 Department of Microbiology, Gayatri College of Biomedical Sciences, Dehradun (UK), India 3 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. –1 Untreated tannery effluent contains 2.673 ± 0.32 to 3.268 ± 0.73 mg l 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. –1 Investigations include the exposure of higher concentrations of Cr(VI) 1.0 to 4.0 mg l 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* human exposure is occupational [2]. Cr(VI) is reactive and a potent carcinogenic species [3]. Wastewater con- Heavy metal contamination of the rivers is a world taining chromium must be treated before being dis- wide environmental problem and its removal is a great charged into the environment. The most commonly challenge. The di-chromate compounds are used as used method to remove Chromium from liquid efflu- oxidizing agents in quantitative analysis of various ents is alkaline precipitation, but the method is expen- water quality parameters such as chemical oxygen sive, therefore cheaper and effective bioremediation demand (COD) and in tanning processes. Chromium techniques using bacteria [4, 5], soils [6], algae [7] and compounds are used in the textile and aircraft industry plants [8] are being studied all over the world. Microor- as mordents and anodizing agents respectively. The fate ganisms can physically remove heavy metals from solu- of chromium in the environment is strongly dependent tion through either bioaccumulation or biosorption. In on its valence state [1]. However, chromium levels in bioaccumulation, metals are transported from the out- air, water, and food are generally very low; the major side of the microbial cell, through the cellular membrane, and into the cell cytoplasm, where the metal is

Correspondence: Dr. Jatin K. Srivastava, Department of Environmental sequestered. Earlier reports of Wong and So [9] sug- Sciences, Chatrapati Shahu Ji Maharaj University, Kanpur – 208024 UP, gested the accumulation of Cu(II) ions by the isolated India E-mail: [email protected] Pseudomonas pudida II-11, from electroplating effluent.

© 2008 WILEY-VCH Verlag GmbH & Co. KGaA, Weinheim www.jbm-journal.com 138 J. Srivastava et al. Journal of Basic Microbiology 2008, 48, 135 – 139 crease in 24 h followed by 48 h however; almost no study least cysteine residues were found in the mi- change was observed in the mass collected after 72 h of crobes (Fig. 2). exposure.

Conclusion Discussion Significant reduction in the metal concentration was Metals play an integral role in the life processes of observed in all the sets with a significant accumulation micro-organisms [18]. Some metals such as Ca, CO, Cr, in the cells. However; the bacterial cells could not sur- Cu, K, Mg, Zn and Na are required nutrients and are vive for longer period in the broth exhibiting the toxic essential for the growth of microbes however; the response of the metal on the bacterial mass. The find- higher concentrations of these metals are toxic to every ings support the well established fact that living organ- living cell. Micro-organisms are highly effective in se- isms naturally selected and can survive the harsh con- questering heavy metals. In the present study bacterial ditions is naturally selected and can be used for the mass was isolated from tannery effluent which is a mitigation of pollution from the water and soil. 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 References of 9 different types of microbes the bacterium (in the present study) was chosen on the basis of growth per- [1] Jeremy, F., Joshua, C., David, F., Ken, K., Bruce, B. and formance in laboratory condition and convenient in Maxim, B., 2000. Non-metabolic reduction of Cr(VI) by bacterial surfaces under nutrient absent conditions. Jour- handling. The bacterial colony was exposed to different nal of Conference & Abstracts, 5, 396. concentration of Cr metal which showed significant [2] Chou, I. N., 1989. Distinct cytoskeleton injuries induced reduction along with accumulation in the cells. The by As, Cd, Co, Cr, and Ni compounds. Biomedical Envi- study showed that bacterial biomass in broth could ronmental Sciences, 2, 358–365. adsorb the metal on the outer surface initially and died [3] Gibbs, H.J., Lees, P.S., Pinsky, P.F. and Rooney, B.C., 2000. as soon as it gets inside the cell of bacterium. The bio- Lung Cancer among workers in chromium chemical pro- sorption of metals does not consume cellular energy. duction. Amer. J. Indust. Medicines, 38, 115–126. Positively charged metal ions are sequestered primarily [4] Ohtake, H, Fujii, E. and Toda, K., 1990. Reduction of toxic through the adsorption of metals to the negative ionic chromate reducing strain of Enterobacter cloacae. Envi- ronm. Technol., 11, 663–668. groups on cell surfaces, the polysaccharide coating [5] Coleman, R.N. and Paran, J.H., 1991. Biofilm concentra- found on most forms of bacteria, or other extra-cellular tion of chromium. Environm. Technol., 12, 1079–1093. structures such as capsules or slime layers. Binding [6] Losi, M.E., Amrhein, C. and Frankenberger, W.T., Jr., sites on microbial cell surfaces usually are carboxyl 1994. Bioremediation of chromate contaminated ground- residues, phosphate residues, SH groups, or hydroxyl water by reduction and precipitation in surface soils. J. groups. Non-essential metals bind with greater affinity Environm. Quality, 23, 1141–1150. to SH group [19]. Bacterial cells which are capable of [7] Brady, D., Letebele, B., Duncan, J.R. and, Rose, P.D., 1994. forming an extra-cellular polysaccharide coating e.g., Bioaccumulation of metals by Scenedesmus, Selenastrum and Pseudomonas sp. bio-adsorbs (biosorp) metal ions and can Chlorella algae. Water SA, 20, 213–218. prevents them from interacting with vital cellular [8] Wolverton, B.C. and Mc Donald, R.C., 1979. The water hyacinth: from prolific pest to potential provider. Ambio, components [20]. The amount of metal biosorbed to the 8, 2–9. exterior of bacterial cells often exceeds the amount [9] Wong, P.K. and So, C.M., 1993. Copper accumulation by a predicted using information about the charge density strain of Pseudomonas putida. Microbios, 73, 113–121. of the cell surface. In the present study, longer survival [10] Silver, S., 1991. Bacterial heavy metal resistance systems of the bacterial cells were found in lower concentration and possibility of bioremediation. In: Biotechnology, Brid- of Cr ions and less in higher concentration. It indicated ging Research and Applications, pp. 265–287. Kluwer that Cr metal might have entered into the bacterial cell Academic Publishers, London. either by ligand interaction or by active membrane [11] Cole, F.A. and Clausen, C.A., 1997. Bacterial biodegrada- uptake. Rouch et al. [21] demonstrated the presence tion of CCA treated waste wood. In: Proceedings, For- est Products Society Conference on Use of Recycled of the cysteine rich proteins in bacterial cells such as Wood and Paper in Building Applications, pp. 201–204 Pseudomonas sp. and Synechococcus spp. which provide (September 9 1996, Madison, WI, USA). Forest Products resistance to the bacterial cells. However; in the present Society.

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 is known to cause allergic skin reactions and cancer [7].Asa result, the total chromium level in effluent is strictly regulated Toxic heavy metals are frequently contained in wastewaters in many countries. In the USA, the concentration of chromium produced by many industrial processes, such as those employed in drinking water has been regulated with a maximum level of in the electroplating, metal finishing, metallurgical, tannery, 0.1 mg/l for total chromium [8]. chemical manufacturing, mining, and battery manufacturing The removal of heavy metals from aqueous solutions has industries [1,2]. The existence of heavy metals in the environ- therefore received considerable attention in recent years. How- ment represents a very significant and long-term environmental ever, the practical application of physicochemical technology hazard. Even at low concentrations these metals can be toxic to such as chemical precipitation, membrane filtration and ion organisms, including humans. In particular, chromium is a con- exchange is sometimes restricted due to technical or economical taminant that is a known mutagen, teratogen and carcinogen [3]. constraints. For example, the ion exchange process is very effec- Chromium is generally found in electroplating and metal finish- tive but requires expensive adsorbent materials [9,10]. The use ing industrial effluents, as well as sewage and wastewater treat- of low-cost waste materials as adsorbents of dissolved metal ions ment plant discharges [4]. Among the several oxidation states provides economic solutions to this global problem and can be (di, tri, penta and hexa), trivalent chromium, Cr(III), together considered an eco-friendly complementary [11,12]. At present, with the hexavalent state, Cr(VI), can be the main forms present emphasis is given to the utilization of biological adsorbents for 2− in aquatic environments [5]. Chromate (CrO4 ) is the prevalent the removal and recovery of heavy metal contaminants. species of Cr(VI) in natural aqueous environments, and is the Biomass involving pure microbial strains has shown high major pollutant from chromium-related industries [6]. Although capacities for the selective uptake of metals from dilute metal- Cr(III) is less toxic than Cr(VI), long-term exposure to Cr(III) bearing solutions. Several investigations have reported that Pseudomonas aeruginosa displays efficiency for metal uptake [13–15]. Chang and Hong [16] found that the amount of mer- ∗ Corresponding author. Tel.: +82 62 970 2442; fax: +82 62 970 2434. cury adsorbed by a P.aeruginosa biomass sample (180 mg Hg/g E-mail address: [email protected] (K.-W. Kim). dry cells) was higher than that bound to a cation exchange

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: Fig. 5. FT-IR spectra of P. aeruginosa prepared in KBr disks: (a) pristine; (b) 2− − + 3+ Cr2O7 + 16e + 14H ↔ 2Cr + 7H2O (4) 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 carried out using chromium-loaded P. aeruginosa. The absorp- metabolic activity [27,28], and most of these focused on the tion spectrum of chromium-loaded biomass (at pH 5) was requirement of external electron donors for reduction to occur compared with that of pristine biomass. The chromium-loaded [29,30]. However, Fein et al. [31] suggested that the Cr(VI) biomass was washed, dried and powdered after biosorption of reduction is not dependent on cell metabolism and that some chromium ions under the same conditions used in the preparation component of the cell wall serves as the electron donor for the of pristine biomass. A change of absorption bands can be seen reduction reaction. We have also investigated the reduction of when comparing the FT-IR spectra of pristine and chromium- Cr(VI) to Cr(III) in the absence of externally supplied electron loaded biomass (Fig. 5). donors. Fig. 5(b) shows the changes in the spectrum of the biomass In the second process, chromium ions are removed from after sorption of Cr(III) by P. aeruginosa. An interesting phe- wastewater using the adsorptive functional groups of P. aerug- nomenon was the sharp decrease in the band intensity at inosa. The adsorptive property is due to the electrostatic inter- 1414 cm−1 corresponding to C O stretching after metal bind- action between the charged surfaces of bacteria and chromium ing. On the basis of the change of the band, it was reasonable ions. The experimental sorption isotherms of Cr(VI) are rep- to assume that the peak value suggested the chelating (biden- resented by the Freundlich sorption isotherm in Fig. 4. The tate) character of the Cr(III) biosorption onto carboxyl groups linearized form of Freundlich is represented by the following [32]. The structure of the metal bound to carboxyl ligands on equation [26]: the bacteria is likely to take the following form [33]: 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) biosorp- In the case of Cr(VI)-loaded bacteria, the spectral analysis tion to P. aeruginosa. The difference of concentration between of P. aeruginosa before and after metal binding indicated that total and hexavalent chromium was taken as the concentration of –NH is involved in Cr(VI) biosorption (Fig. 5(c)). There is a trivalent chromium. These results show that the bacterial func- substantial decrease in the absorption intensity of –NH bands tional groups of P.aeruginosa can act as reductive and adsorptive at 1660 and 1551 cm−1. The broad overlapping range for N–H sites in metal biosorption. 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 3.3. FT-IR spectra of chromium-loaded P. aeruginosa the shift. These amino groups are protonated at pH 3 [34] and the negatively charged chromate ions become electrostatically To conﬁrm the difference between functional groups in rela- attracted to the positively charged amines of the biomass cell tion to biosorption of Cr(III) and Cr(VI), the FT-IR study was wall. Similar to Cr(III)-loaded bacteria, the characteristic peak 58 S.-Y. Kang et al. / Biochemical Engineering Journal 36 (2007) 54–58

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Biological removal of hexavalent chromium in trickling filters operating with different filter media types

E. Dermoua, A. Velissarioub, D. Xenosa, D.V. Vayenasa* aDepartment 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] bHellenic 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.

1. Introduction pounds through oxidation, reduction or methyl- iation into more volatile, less toxic or readily pre- Hexavalent chromium contamination in the cipitating form [11]. environment is a result of the extensive use of Microorganisms interact with toxic metals and chromate and dichromate in numerous industries mediate their removal through different processes including stainless-steel production, leather tan- like bioaccumulation, biosorption and enzymatic ning, electroplating, pigment fabrication, wood reduction [12]. Metabolic enzymatic Cr(VI) re- preserving, power plants and nuclear facilities [1]. duction has been observed both in the presence

Chromium can occur at several different oxida- and absence of O2 and Cr(VI) reduction rates are tion states ranging from –2 to 6. However, only strongly dependent on pH, temperature, electron Cr(III) and Cr(VI) are the stable forms in the natu- donor and Cr(VI) concentrations. The optimum ral environments. Cr(VI) is rarely naturally oc- pH conditions for metabolic enzymatic Cr(VI) curring, is relatively soluble in aqueous systems reduction by bacteria appears to be close to pH 7 and is readily transformed in groundwater [2]. [13]. Biochemical studies of enzymatic Cr(VI) Exposure to Cr(VI) poses an acute health risk be- reduction reveal that Cr(VI)-reducing mechanisms cause it is highly toxic and chronic exposure can are likely associated with bacterial electron trans- lead to mutagenesis and carcinogenesis [3]. On port systems [12]. Therefore, several researchers the contrary Cr(III) is naturally occurring, is much have proposed that Cr(VI) reduction is mediated less toxic and even essential to human glucidic by enzymes that are not substrate specific for metabolism, contributing to the glucose tolerance Cr(VI) and that “chromate reductases” may be factor necessary for insulin-regulated metabolism serendipitous contributors to Cr(VI) reduction [4]. while engaged in other primary physiological The discharge of Cr(VI) to surface water is functions [1]. regulated to below 0.05 mg/l by the US EPA [5] In literature, most of the previous studies on and the European Union [6], while total Cr, biological reduction of Cr(VI) were conducted Cr(III), Cr(VI) and its other forms, is regulated to under sterilized conditions (pure cultures of mi- below 2 mg/l. The most commonly used technol- croorganisms) and resulted in very small removal ogy for treatment of heavy metals in wastewaters rates [8,14,15]. There are many microorganisms is chemical precipitation, which involves reduc- which can reduce Cr(VI) and have been identi- tion of Cr(VI) to Cr(III) by a reducing agent un- fied; examples include Shewanella putrefaciens der low-pH conditions and subsequent adjustment MR-1, Pseudomonas fluorescens LB300, which of solution pH to near neutral ranges to precipi- reduces Cr(VI) under aerobic and anaerobic con- tate Cr(III) as hydroxides. However, this method ditions, Agrobacterium radiobacter EPS-916, results in high costs and secondary waste genera- which is capable of reducing Cr(VI) aerobically, tion due to various reagents used [7]. Enterobacter clocae strain HO1, which only re- Nowadays, microbiological detoxification of duces Cr(VI) under anaerobic conditions using a hexavalent chromium has gained importance due limited selection of electron donors. Also, a mem- to the emphasis placed on protection of the envi- ber of the obligatory anaerobic sulfate bacteria ronment [8]. Several bacterial species are capable (D. vulgaris NCIMB 8303) has a well-established of transforming Cr(VI), into the much less toxic ability to reduce metals, including Cr(VI), if a and less mobile Cr(III), including both Gram-posi- complexing agent is incorporated to chelate Cr(III) tive and Gram-negative species [9,10]. Bacteria [16,17]. However, there are a few recent studies may protect themselves from toxic substances in investigating Cr(VI) reduction using mixed cul- the environment by transforming toxic com- tures of microorganisms [18–20]. 158 E. Dermou et al. / Desalination 211 (2007) 156–163

Batch reactors, continuous-flow and fixed film 2.3. Analytical methods bioreactors were also used for biological reduc- During all experiments, hexavalent chromium tion of Cr(VI) [8,13,21–23]. In a recent study [11] concentration, pH, temperature, dissolved oxygen a pilot-scale trickling filter was constructed and concentration and TOC measurements were made tested for biological chromium (VI) removal from on a daily basis. Samples were filtered through industrial wastewater. Three different operating 0.45 µm–Millipore filters (GN-6 Metricel Grid 47 modes were used to investigate the optimal per- mm, Pall Corporation). Hexavalent chromium formance and efficiency of the filter, i.e. batch, concentration was determined by the 3500-Cr D continuous and SBR with recirculation. The lat- Colorimetric method according to Standard Meth- ter one was found to achieve removal rates up to ods for the Examination of Water and Wastewa- 3.5 g Cr(VI)/d while aeration was taking place ter [24]. Total Organic Carbon measurements naturally without the use of any external mechani- (TOC) were conducted in order to determine the cal means. In this study, an attempt was made to feed sodium acetate concentration both in the estimate the effect of the support media on Cr(VI) batch reactor and the bulk liquid of the bioreactor, reduction rate. Mixed culture of microorganisms, following the methods described in Standard originating from an industrial sludge, were used Methods for the Examination of Water and Waste- in pilot-scale trickling filters using two different water [24] by using, Total Organic Carbon Ana- filter media types, i.e. calcitic gravel and hollow lyzer (TOC-V , SHIMAZDU Corporation, JA- plastic tubes, under SBR operation with recircu- CSH PAN). The cell density of liquid culture was de- lation, suggesting the most efficient support me- termined as optical density at 600 nm on a Jasco dia type. V-530, spectrophotometer.

2. Materials and methods 2.4. Isolation and enrichment of indigenous bacteria 2.1. Media Samples of industrial sludge were taken from The influent feed to the bioreactor was pre- the Hellenic Aerospace Industry S.A. In order to pared by dissolving 1g NH Cl, 0.2 g MgSO 7H O, 4 4 2 grow bacterial strains able to reduce hexavalent 0.001 g FeSO 7H O, 0.001 g CaCl 2H O, 2.5 g 4 2 2 2 chromium, a sludge sample of 10 grams was added CH COONa3H O and 0.5 g K HPO in 1.0 l of 3 2 2 4 in a 2 l Erlenmeyer flask and was diluted in an tap water. acetate-minimal medium and concentrated chromium solution (in the form of K Cr O ) resulting 2.2. Reagents 2 2 7 in a final hexavalent chromium concentration of Stock Cr(VI) solution (500 mg/l) was prepared 50 mg/l. The final volume of the solution was 1 l. by dissolving 141.4 mg of 99.5% K2Cr2O7, previ- Acetate-minimal medium (AMM) was compris- ously dried at 103°C for 2 h, in Milli-Q water and ing (per litre) 1g NH4Cl, 0.2 g MgSO47H2O, 0.001 diluting to 100 ml. Diphenyl carbazide solution g FeSO47H2O, 0.001 g CaCl22H2O, 2.5 g was prepared by dissolving 250 mg of 1,5-diphe- CH3COONa3H2O, 0.5 g yeast extract and 0.5 g nylcarbazide in 50 ml of HPLC-grade acetone and K2HPO4 in 1.0 l of tap water (micronutrients were storing in a brown bottle, for Cr(VI) concentra- supplied using tap water as diluent), while the fi- tion measurements. 1,5-diphenylcarbazide was nal pH of the nutrient solution was adjusted to 7. purchased from Fluka Chemical, potassium The solution was kept under oxic conditions dichromate was purchased from Sigma Chemical through aeration and mixing while nutrients and Co. hexavalent chromium were added according to the E. Dermou et al. / Desalination 211 (2007) 156–163 163

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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 insoluble Cr(OH)3 at neutral pH, and becomes almost immobile in the environment [5,6].Cr3+ is also an essential micronutrient Chromium, a toxic heavy metal, dissipates into environ- compared with the toxic, mutagenic and carcinogenic hexava- ment as a result of various industrial activities [1,2] such as lent chromium [7] in addition to its reduced toxicity due to its steel manufacturing, metal plating, mining, leather tanning, tex- low bioavailability. In view of toxicity and related environmen- tile dying, cement industries etc. Of all the different oxidation tal hazards the level of chromium in wastewater must be reduced states trivalent and hexavalent chromium exist as stable species. to a permissible limit (5.0 mg/L and 0.5 mg/L for trivalent The hexavalent chromium species exists in aqueous solution and hexavalent chromium, respectively) [8] before discharg- −2 as oxyanionic entities like chromate (CrO4 ), bichromate ing into water bodies. The removal of chromium employing − −2 (HCrO4 ) and dichromate (Cr2O7 ), the relative distribution conventional methodologies [9,10] like ion exchange, chemi- of which depends on the solution pH [3,4]. Two other forms cal precipitation or reverse osmosis suffer from limitations like −2 −2 of chromium, e.g. Cr3O10 , and Cr4O13 have also been high operating cost, incomplete precipitation, sludge genera- detected in highly acidic medium [4]. The oxyanionic entities tion, etc. On the other hand biosorption is receiving increasing of Cr6+ do not bind to the negatively charged mineral surfaces, attention as an emerging technology for the removal of heavy e.g. silica or clay, become highly mobile in the environment metals from contaminated effluents. The process is based on and soluble in a solution of neutral pH. In comparison, Cr3+ the adsorption behavior of certain biological materials towards (3−n)+ forms stable hydroxo complexes [e. g. Cr(OH)n ] and the organic or inorganic substances from their solution. Different cationic Cr3+ having strong affinity for particle surfaces yields types of adsorbents [8,11–17] including activated carbon, sawdust, cactus leaves, lignin, spent grain, chitin, chitosan, jacobsite (MnFe O ), etc. used for the removal of chromium results in low ∗ 2 4 Corresponding author. Tel.: +91 33 2473 4971/5904x502; removal requiring prolonged equilibrium time. Recently many fax: +91 33 2473 2805. E-mail addresses: sujoy [email protected] (S.K. Das), efforts have been directed towards the development of specific [email protected], [email protected] (A.K. Guha). biosorbents, the performance of which depends on the biomass 1 Tel.: +91 33 2473 4971/5904x502; fax: +91 33 2473 2805. characteristics and microenvironment of target metal ion solu-

Fig. 1. Effect of pH on chromium adsorption by Termitomyces clypeatus biomass at 30 ◦C (A), (᭹)Cr6+ uptake, ()Cr3+,() Zeta potential; adsorption isotherm of Cr6+ on T. clypeatus biomass (B), (᭹) 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 optimum pH of chromate reductase is around pH 7.0 and activ- 3+ value 6.0 (Fig. 1A). Precipitation of Cr as Cr(OH)3 at pH ity decreases significantly at low or high pH values [37]. So, value > 6.0 restricted the experiment for solution having pH val- enzymatic reduction of Cr6+ to Cr3+ can be ruled out at low ues beyond 6.0. Zeta potential measurements indicate that the pH. Accordingly we have measured the amount of both hexava- overall surface charge remained positive at low pH values, which lent and trivalent chromium in the solution after adsorption with facilitated the attraction of negatively charged oxyanionic chro- TCB at pH 2.0, 3.0, and 4.0. The results presented in Table 2 mate ions. On the other hand, decreased adsorption of Cr3+ at show the presence of both hexavalent and trivalent chromium pH 2.0–3.0 is attributed to the competition of the binding sites in the aqueous solution after sorption of chromate with TCB. for H+ ions. From the experimental data, it is noted that at pH This indicates that chromate being a strong oxidizing agent oxi- 2.0 adsorption of Cr6+ was less than 20% compared with that dizes some functional groups of biomass and in the process gets at pH 3.0. Similar behavior of hexavalent chromium adsorption reduced to trivalent state. It was also observed that proportion on a variety of biosorbents including fungal biomasses was also of trivalent chromium to hexavalent chromium decreased with reported earlier [33–35]. However, the rationale behind such influence of pH on the sorption process has not yet been ade- Table 2 quately discussed. A number of factors are to be considered to Influence of pH on reduction of Cr6+ understand the role of pH in the sorption behavior of chromium 6+ 3+ 3+ 6+ on TCB. Hexavalent chromium can be reduced to trivalent state pH of the Uptake Total Cr Cr mg/L Cr mg/L Cr /Cr solution (mg/g) (mg/L) both chemically [36] or enzymatically [37]. The cell wall of TCB contains different electron donors and their close proximity to 2 8.77 25.69 3.25 22.44 6.9 aqueous chromate ions result in the formation of Cr3+ species. 3 11.15 7.25 1.51 5.74 3.8 4 7.03 39.61 24.36 15.25 0.63 This reduction is facilitated at low pH values. On the other hand, 50 S.K. Das, A.K. Guha / Colloids and Surfaces B: Biointerfaces 60 (2007) 46–54 increasing in pH of the solution. Additionally, to confirm the model (Fig. 1, panels B and D) exhibiting deviation from linear- presence of both the forms of chromium on the biomass, we have ity over the entire concentration range. 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 3.5. Model analysis chromium in the eluent. The results show (data not shown) the presence of both trivalent and hexavalent chromium but the per- Linear coefficients of determination, r2, and non-linear χ2- 3+ 6+ centage of Cr being much higher than that of Cr in the HCl square test, χ2 may be used to evaluate the ‘goodness of fit’ of eluent while in NaOH eluent opposite results were observed. curves to the data they summarize and may be used to estimate This further confirms our hypothesis that chromate ions after the probability of obtaining any series of deviations of observed 3+ initial binding to the cell surface were reduced to Cr . Thus, values from predicted values [43]. The value of linear coefficient the optimum pH value (3.0) can be explained as; at low pH of determination, r2, represents the percentage of variability in value positively charged functional groups adsorbed chromate the dependent variable that has been explained by the regression ions through anion exchange mechanism, however, as soon as line and may vary from 0 to 1. If there is no relationship between the pH was lowered below the optimum pH value, the chromate the predicted values and actual values, the coefficient of deter- 3+ 3+ oxidize the biomass and produced Cr ions [36,38]. These Cr mination is zero or very low, a perfect fit gives a coefficient of ions then competed for the binding sites with the protons via the 1.0. On the other hand χ2-square test, χ2 is basically the sum cation exchange reaction resulting in low adsorption. The des- of the squares of the differences between the experimental data 3+ orption of Cr from the biomass at that low pH value also lead and data obtained from the models, with each square difference to low uptake of chromium [38]. Consequently, the sorption of divided by the corresponding data obtained by calculation from 6+ Cr from its aqueous solution by TCB was maximized at a pH the models. This can be represented mathematically as value which was high enough for anion exchange as well as q − q 2 redox reaction to proceed simultaneously. Thus, we presumed χ2 = ( e e,m) that the influence of hydrogen ion concentration on the present qe,m sorption process is complex in nature and the noted optimum pH value is the resultant of all these factors. However, 25% of where qe,m is equilibrium uptake obtained by calculation from total adsorption also took place at a higher pH (6.0–7.0) value, the model (mg/g) and qe is the experimental data of the equi- χ2 χ2 although cell surface contained negative charges. Based on this librium uptake (mg/g). -square, will be small number if experimental data we conclude that in addition to electrostatic the experimental data are close to that obtained from model forces of attraction, other factors such as reduction, precipitation, and will be bigger if they differ. Therefore, it is necessary to chemical interaction and physical forces such as hydrogen bond- analyze the data sets using both linear coefficient of determi- χ2 ing and or ion–dipole interaction also involved in the sorption nation, and non-linear -square test to establish the best fit process. isotherm model for the adsorption system. The coefficient of determination, r2, and Chi square test, χ2, were determined 3.4. Sorption isotherm in the range of the whole metal ion concentration. The linear regression coefficient, r2, values were 0.998 and 0.965, respec- Sorption isotherm is a prerequisite to describe the stoi- tively for Langmuir and Freundlich models. On the other hand χ2 χ2 chiometric solute–solid interaction. A few parameters, such the -square, , values of 5.03 and 18.53 for Langmuir and as maximum sorption capacity, are important for optimizing Freundlich, isotherm, respectively with 9 degrees of freedom φ the design of sorption system, and analyzed using linearized ( ) corresponds to the significance level between 75–90% and forms of the isotherm models of Langmuir [39] and Glas- 2.5–5% according Fischer and Yates chart [44]. Thus, Lang- tone [40]. The result shows that the present sorption process muir isotherm model appears to be the best fit for the present of chromium on the live TCB followed the Langmuir model adsorption process. (Fig. 1, panels B and C) indicating chemisorption and monolayer X-ray elemental mapping of the biomass after chromium coverage of sorbate on the sorbent. The theoretical monolayer sorption (Fig. 2) depicts a uniform distribution of chromium over the entire surface area which further supports the observed saturation capacity (Qmax) of the sorbate on the sorbent calculated (54.05 mg/g) from the slope of the linearized curve of Langmurian behavior of sorbate on the sorbent surface. Langmuir model (Fig. 1C) was very close to that obtained experimentally from isotherm study (53.95 mg/g) (Fig. 1B), which 3.6. Chemical characterization of metal ion adsorption is higher than those reported for other types of biosorbents [12–14,16,41]. Sorption capacity of live TCB towards chromium In living cells the sorption mechanisms include both was found to be higher than that reported for activated carbon metabolism dependent and independent processes. Metabolism (15.47 mg/g) [18], though it covers much higher BET surface independent uptake process essentially involves cell surface area (500–3000 m2/g) [42] compared to that of TCB (15.4 m2/g). binding through ionic and chemical interaction, while dependent Therefore, we summarized that besides BET surface area other process deals with the binding of both the surfaces followed by properties of the sorbent such as functional groups of the biomass intracellular accumulation [21,45]. The cell wall of the fungal played important roles in the studied sorption process. The sorp- biomass generally contains large amounts of chitin, chitosan, tion phenomenon did not fit well with the Freundlich isotherm glucan and mannan as well as small amounts of glycoprotein 54 S.K. Das, A.K. Guha / Colloids and Surfaces B: Biointerfaces 60 (2007) 46–54 and Prof. A. Dasgupta (Department of Biochemistry, Calcutta [26] R.L. Smith, F.E. Strohmair, R.S. Oremland, Arch. 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BMC Microbiology BioMed Central

Research article Open Access 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 Received: 26 May 2005 Accepted: 25 January 2006 BMC Microbiology 2006, 6:5 doi:10.1186/1471-2180-6-5 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.

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 cold- Cr10°CFigure6+ reduction 2 by isolate P4 under aerobic conditions at active. Cr6+ reduction by isolate P4 under aerobic conditions at 10°C. 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 A number of studies suggest both growth-dependent and reduction in the environment is also possible. The lack of growth-independent chromium reduction [20,29,40]. In reduction in the sterile controls along with the lack of Cr6+ either case, chromium reduction does seem to be biomass adsorption to cell biomass in the three adsorption experi- dependent in our study as well as in others [21,41]. The ments suggests that the members of the enrichment com- lag at the beginning of the consortium reduction experi- munity (which included isolate P4) were responsible for ments as well as observations of increased turbidity the reduction of Cr6+. Since most aquifers contaminated throughout the experiments suggests that adequate cell with Cr6+ have levels below 60 mg/l, these experiments biomass must be produced before reduction begins in ear- would also suggest remediation of the lower levels of Cr6+ nest. Bopp and Ehrlich [28] showed that higher concen- contamination present in aquifers is possible. Bioremedi- trations (1000 mg/l) of Cr6+ produced a much longer lag ation literature suggests low levels of contamination are phase and a significantly lower final yield of biomass than very difficult to completely remediate. Lack of induction lower concentrations. The reduced biomass would also of enzyme systems at low contaminant concentrations contribute to the lack of complete reduction found at and problems with availability of contaminants bound to higher concentrations in many studies [22,25] as well as organics and sequestered in other matrices all contribute in the higher concentrations tested in our lab (data not to persistence of contaminants in the environment. It has shown). Previous studies using cellular biomass grown on also been suggested that indigenous microorganisms may uncontaminated substrates to test Cr6+ reduction greatly be more successful in reducing low contaminant concen- decreased the amount of time required to completely trations [8]. The complete reduction of Cr6+ at 10°C in reduce Cr6+ [21,39], similar to our findings with the iso- this study using an indigenous member of the Hanford late P4 reduction experiments (Figure 2). Increased tur- microbial community and past studies with indigenous bidity after only 24 hours in R2 broth at 10°C (grown mesophilic microorganisms suggest that there are envi- aerobically) and the achievement of stationary phase (as ronmental candidates for reduction of the low levels of determined by absorbance readings, 1:10 dilution in R2 contamination usually found in aquifers [23,43]. 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 Studies have shown Arthrobacter species adsorbing Fe, Cd, suggesting the isolate is a true psychrophile. Growth and Cu, but not Cr [44,45]. Chromium has, however, appeared fastest at 18°C. The ability to increase biomass been shown to adsorb to both Shewanella and Bacillus spe- in a short time given the proper nutrients suggests that P4 cies [46]. Adsorption studies performed on the Hanford

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39. Shen H, Wang YT: Modeling hexavalent chromium reduction in Escherichia coli 33456. Biotechnology and Bioengineering 1994, 43:293-300. 40. 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. 41. Phillip L, Venkobachar C, Iyengar L: Immobilized microbial reactor for the biotransformation of hexavalent chromium. Inter- national Journal of Environment and Pollution 1999, 11:202-210. 42. Chirwa EMN, Wang Y: Hexavalent Chromium reduction by Bacillus sp. in a packed-bed bioreactor. Environ Sci Technol 1997, 31:1446-1451. 43. 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. 44. Pagnanelli F, Petrangeli Papini M, Toro L, Trifoni M, Veglio F: Bio- sorption of metal ions on Arthrobacter sp.: Biomass characterization and biosorption modeling. Environ Sci Technol 2000, 34:2773-2778. 45. 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. 46. Fein JB, Fowle DA, Cahill J, Kemner K, Boyanov M, Bunker B: Non- metabolic reduction of Cr(VI) by bacterial surfaces under nutrient-absent conditions. Geomicrobiol J 2002, 19:369-382. 47. 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. 48. 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. 49. Margesin R, Schinner F: Bacterial heavy metal-tolerance- extreme resistance to nickel in Arthrobacter spp. strains. J Basic Microbiol 1996, 36:269-282. 50. 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. 51. 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. 52. 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. 53. Sheridan PP, Loveland-Curtze J, Miteva VI, Brenchley JE: Rhodoglo- bus 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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Page 8 of 8 (page number not for citation purposes) 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 with their acute toxicities and poorly understood degradation pathways. Organophosphorus compounds (OPs) are widely used as pes- Our focus in this paper is the fate of OPs in the aqueous envi- ticides, insecticides in agricultural as well as non-agricultural ronment, in which they are known to degrade spontaneously practices. They have mostly replaced organochlorine (OCs) through different pathways including hydrolysis, photolytic compounds in such applications because compared to the OCs, oxidation, microbial transformations and other biological pro- OPs are less bioaccumulative and more readily degradable in cesses. Among these processes, biodegradation is particularly the environment [1]. Currently OPs account for about one-third attractive because of its effectiveness and low cost. OPs can be of the total pesticide consumption in the world [2]. OPs are biodegraded by plants, algae, fungi and bacteria, and the key known to inhibit the activity of acetylcholinesterase (AChE), players in these processes are enzymes, such as hydrolase and with subsequent accumulation of acetylcholine at nerve endings, oxidase [4–7]. causing major acute toxic effect [3]. The widespread contam- The analysis of OPs in water samples is generally per- ination of soils, sediments and aquatic environment by OPs formed by solid-phase extraction (SPE) [8–11] followed by thus creates different set of environmental problems associated 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 ∗ Corresponding author. Tel.: +86 532 88963253; fax: +86 532 88963253. thermally labile, highly polar and non-volatile properties. Liquid E-mail address: [email protected] (F.S.C. Lee). chromatography–mass spectrometry (LC–MS) in these cases

0021-9673/$ – see front matter © 2006 Elsevier B.V. All rights reserved. doi:10.1016/j.chroma.2006.10.091 182 J. Liu et al. / J. Chromatogr. A 1137 (2006) 180–187 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 Fig. 2. HPLC/DAD chromatograms of MP: (a) standard and (b) degradation 0.1% (g/g) ammonium formate in order to improve the ioniza- without bacteria present for a week at room temperature. tion of analytes and make them amenable to MS detection. The solvent programming involved first an isocratic run from 0 to solution after adsorption by SPE cartridge, and the acetonitrile 5 min with a mixture of 10% of B in A; followed by solvent extract of the solids portion from centrifugation. Preliminary program from 10 to 50% B in A in the next 5–20 min. The flow screening of the three samples showed that most of the OP rated was constant at 0.8 mL/min, of which one half of the flow degradation products were in the first fraction, i.e., the soluble was directed into the MS ionization chamber. components extracted by SPE. Thus, later analytical effort was The MS detector was a Agilent G1969A TOFMS combined focused primarily on this product fraction. In Fig. 2, the HPLC- with electrospray (ESI) as the ion source. The drying gas tem- UV chromatograms of a fresh standard MP solution and the same perature was set at 300 ◦C and the velocity of the drying gas standard solution after 1 week standing at room temperature are was 10 L/min. The ion polarity was positive, and the capillary compared. In the absence of bacteria, the MP is shown to be sta- voltage was set to 4000 V. Two different collision induced dis- ble without the production of detectable degradation products. sociation (CID) voltages, 50 and 150 V, were used in order to The MP peak (18.33 min) in the chromatograms was identified provide more structural information by MS analysis. The m/z by both standard calibration and TOFMS spectrum. The pro- acquisition range was set at 10–1000. This TOFMS possesses cess described below exemplifies a typical peak identification high mass resolving power of more than 10,000, which means process in TOFMS analysis. The exact mass of peak 1 was mea- that the deviation in a measured mass can be controlled to be sured as m/z 264.0089 (Fig. 3a), resulting from the protonation less than 5 ppm in terms of m/m. The acquisition rate for this adduction of the molecular ion (mass 263). Data were processed TOFMS is 10000 times per second. Ions with m/z 121.0509 and using Analyst QS software from Agilent TOFMS. First, param- 922.0098 were used as reference ions in mass measurement in eters, such as elemental composition and number limit, mass order to eliminate systematic errors. error tolerance and number of charges must be pre-set based on the background knowledge. For this TOFMS, mass error tol- 3. Results and discussion erance can be set at 5 ppm, which represents the uncertainties imposed by the systematic error. For analysis of MP and its 3.1. Identification of bacteria strains degradation products, elements C, H, O, N, P, S were chosen as the possible elements. Besides mass measurement, the isotopic From the results of the degradation experiment, two bacteria mass distributions also provide information about the variety and strains showed significant MP degrading activity. They were number of elements. The software in TOFMS is capable of cal- marked as L-10 and S-2. After extraction, PCR amplification and culating all possible molecular formula satisfying the observed sequencing of their 16S rRNA genes [24], they were identified data. Compared with other possible assignments, C8H11NO5PS to belong to the genus Shewanella and Vibrio parahaemolyticus, has the least mass error. Furthermore, its theoretical isotopic respectively. Some Shewanella species have been confirmed to mass distributions match excellently with the one measured by be a kind of metal ion reducing bacteria capable of detoxicating TOFMS in both intensity and m/z position (Fig. 3b). Combin- metals from high valence to low valence state, e.g., U(VI) to ing this information with double bond equivalent (DBE) and U(III) or Cr(VI) to Cr(III) [25–27], and V. parahaemolyticus is the MS fragment pattern in GC/MS (Table 2), the analyte can be known to cause gastrointestinal illness in humans [28]. unequivocally identified as MP.The other m/z 281.0352 (Fig. 3a) also can be identified as formula C8H14N2O5PS (Fig. 3c), which 3.2. LC/ESI-TOFMS analysis of degradation products is the ammonium adduct of MP because of the presence of 0.1% ammonium formate in the mobile phase. As described earlier in Section 2, three fractions were sep- Fig. 4 shows the results of MP degraded by bacteria for 1 arated from the samples after biodegradation and they were: week. Here MP peak (peak 1) disappears almost completely. the soluble components extracted by SPE, the supernatant clear The estimated % of parent species degraded based on relative J. Liu et al. / J. Chromatogr. A 1137 (2006) 180–187 187

3.4. Mechanism pathways of OPs degradation groups in the MP parent molecule, and thus the weakening of the nucleophilicity of the P atom, which is known to be responsible The bio-transformation of MP generally proceed in two for MP toxicity. phases as same as those commonly observed in the metabolism of extrinsic matters in vivo [31]. In phase one, oxidation, Acknowledgement reduction and hydrolysis are the main reactions. Afterwards, the metabolites will combine with some intrinsic matters The authors thank Qingdao “2004 JiangCai Plan” (04-3- through hydroxyl, amino or carboxyl linkage. During the whole JJ-11) and Agilent Technologies Co., Ltd. (China) to provide metabolic process, the molecular backbone of these extrinsic HPLC/TOFMS and GC/MS system for our lab. compounds maintain their integrity, with changes occur only in the substituent functionalities. References Combining information obtained from GC/MS and LC/TOFMS analysis, the biodegradation pathways for MP can [1] M. 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Liu (Ed.), Practice Pharmacokinetics, Chinese Medical Technology Publish, Beijing, 2003, p. 69. of the thiophosphoric moiety. However, the toxicity could be [32] H. Soreq, S. Seidman, Nat. Rev. Neurosci. 2 (2001) 294. somewhat reduced because the dominant processes involved in [33] C.B. Millard, G. Koellner, A. Ordentlich, A. Shafferman, I. Silman, J.L. the degradation processes is the conversion of nitro to amino Sussman, J. Am. Chem. Soc. 121 (1999) 9883. SYNCHROTRON RADIATION INFRARED SPECTROMICROSCOPY:ANONINVASIVE CHEMICAL PROBE FOR MONITORING BIOGEOCHEMICAL PROCESSES

H.‐Y. N. Holman1,2 and M. C. Marti3 n 1Ecology Department, Earth Sciences Division, Lawrence Berkeley National Laboratory, University of California, Berkeley, California 94720 2Virtual Institute for Microbial Stress and Survival, Lawrence Berkeley National Laboratory, University of California, Berkeley, California 94720 3Advanced 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 non- invasive 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

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 SYNCHROTRON IR SPECTROMICROSCOPY 81 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 inter- ested 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. 116 H.‐Y. N. HOLMAN AND M. C. MARTIN

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 polychlori- nated 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 versati- lities 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 ϩ .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

1, 1,2 3,† 1 DENIS G. RANCOURT, *PIERRE-JEAN THIBAULT, DENIS MAVROCORDATOS, and GILLES LAMARCHE 1Lake Sediment Structure and Evolution (LSSE) Group, Department of Physics, University of Ottawa, Ottawa, ON K1N 6N5, Canada 2Department of Earth Sciences, University of Ottawa, Ottawa, ON K1N 6N5, Canada 3Particle Laboratory, Swiss Federal Institute for Environmental Science and Technology, Postfach 611, U¨ berlandstrasse 133, CH-8600 Dübendorf/Zurich, Switzerland

(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, ϳ108 cells/mL) and under otherwise identical chemical conditions, starting from Fe(II) (10Ϫ2,10Ϫ3,or10Ϫ4 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) (10Ϫ3.5–10Ϫ4.5 mol/L) was added to a fixed concentration of cells (ϳ108 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-Fe2ϩ by a proposed steric and charge transfer effect and (2) it stabilizes Fe(II) as sorbed-Fe2ϩ 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

1. INTRODUCTION al., 1999; Fowle and Fein, 1999; Fein et al., 2001; Yee and Fein, 2001; Kulczycki et al., 2002) rather than direct physical 1.1. Bacterial Metal Sorption and methods such as local coordination environment and valence Bacterium-HFO Associations state sensitive spectroscopies. This followed early direct observations of nonmetabolic metal sorption onto bacterial cell walls Bacteria are ubiquitous in aquatic and humid environments, in sediments (Degens et al., 1970; Degens and Ittekkot, 1982) where they constitute a quantitatively important metal sorbing and early laboratory investigations of the phenomenon (e.g., compartment (Beveridge and Doyle, 1989; Ledin, 2000; Warren and Haack, 2001) that is actively being studied, usually Starkey, 1945; Pringsheim, 1949; Beveridge and Murray, by wet chemical methods such as titration (e.g., Fein et al., 1980). The importance of passive (i.e., nonmetabolic) cell wall 1997; Daughney and Fein, 1998; Daughney et al., 1998; Cox et 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., * Author to whom correspondence should be addressed Cr, using XANES; Fein et al., 2002a), as we have done for the ([email protected]). ﬁrst time for Fe, using Mössbauer spectroscopy. This directly †Deceased. gives bonding environments, electron transfers, steric re- 553 Biotic precipitation of hydrous ferric oxide 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 Fe3ϩ doublet contribution (dotted line) and a cell wall sorbed Fe2ϩ doublet contribution (solid line), and (b) spectrum SB3-9N-LNT (Tables 1 and 2) showing a 2L-Fh Fe3ϩ doublet (dotted line) and no Fe2ϩ 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 (Tables 1 and 2): (1) 21 Ϯ 2% in sample SB1 after4dofwet magnetic order of the HFO at 4.2 K, (2) cannot be sorbed in aging; (2) 18 Ϯ 6% in sample SB6 after 1d of wet aging; and (3) dense patches of Fe2ϩ that would have intercation magnetic 20 Ϯ 10% in sample SB5, counting only the sorbed fraction and superexchange bonds, as would be required if they were to excluding the 80% of total Fe that is HFO, after1dofwetaging. form a template for oxide growth by re-oxidation, and (3) most Based on the larger total Fe concentration used in a fourth sample probably do not form a separate Fe2ϩ-rich solid phase because (SB2, Table 1), we expect that 16.6% of the total Fe in this sample most such phases would be magnetically ordered at 4.2 K. is sorbed-Fe, since 10Ϫ4.5 mol/L gives full sorption coverage at Since TEM observations of similar samples (Warren and Ferris, this pH (sample SB1 and Warren and Ferris, 1998). Therefore, the 1998) show either only sorbed-Fe or sorbed-Fe with HFO observed 5 Ϯ 2% of total Fe that is sorbed-Fe2ϩ in sample SB2 nanogranules, depending on pH and total Fe concentration, we after 13 d of wet aging corresponds to 30 Ϯ 12% of sorbed-Fe that conclude that some of the originally sorbed Fe3ϩ must have is sorbed-Fe2ϩ. been reduced to sorbed-Fe2ϩ, presumably by either oxidizing The observed production of sorbed-Fe2ϩ may have been the functional group or by transferring an electron from some enhanced by metabolic consumption of oxygen during wet other component of the bacterium. aging, using dead cells and ancillary organics as electron A2ϩD contribution is always detected in spectra that display donors. It appears to be a form of nonmetabolic reduction measurable amounts of sorbed-Fe3ϩ (Table 2) and the detected not unlike the reduction of Cr(VI) to Cr(III) observed by sorbed-Fe2ϩ amount always increases with increasing durations of Fein et al. (2002a). The observed sorbed-Fe2ϩ is compatible wet aging (under bottle top atmosphere, at 4°C, and with some with our a/b-HFO experiments described below (section exposure to higher temperatures) (Tables 1 and 2). There appears 5.3), where Fe(II) was employed and was stabilized up to to be a maximum amount of sorbed-Fe2ϩ of ϳ20% of total ϳ100% of available bacterial sorption sites, in the b-HFO sorbed-Fe. This saturation amount occurs in at least three samples preparations. Biotic precipitation of hydrous ferric oxide 575 ing the making of the SB samples and sharing the costs of the Möss- Cowen J. P. and Bruland K. W. (1985) Metal deposits associated with bauer measurements. We thank Dr. Nagina Parma (University of To- bacteria: Implications for Fe and Mn marine biogeochemistry. ronto) for performing all the surface complexation (SB) syntheses. We Deep-Sea Res. 32, 253–272. thank Prof. Danielle Fortin for her help and guidance in initiating the Cox J. S., Smith D. S., Warren L. 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(2004) Characterization of iron-oxides formed by oxidation of Ferris F. G., Shotyk W., and Fyfe W. S. (1989b) Mineral formation and ferrous ions in the presence of various bacterial species and inor- decomposition by microorganisms. In Metal Ions and Bacteria ganic ligands. Geomicrobiol. J. 21, 99–112. (eds. T. J. Beveridge and R. J. Doyle), pp. 412–441. John Wiley, Coleman A. W. (1980) Enhanced detection of bacteria in natural New York. environments by fluorochrome staining of DNA. Limnol. Ocean- Fleisch H. (1975) Physico chemical and biological effects of pyrophos- ogr. 25, 948–951. phate and diphosphonates. Colloques internationaux du Centre na- Cornell R. M. and Schwertmann U. (1979) Influence of organic anions tional de la recherche scientifique. Physico-chimie et cristallogra- on the crystallization of ferrihydrite. Clays Clay Miner. 27, 402– phie des apatites d’intérêt biologique. 230, 263–273. Paris, Éditions 410. du Centre national de la recherche scientifique. Cornell R. M. and Schwertmann U. (1996) The Iron Oxides—Structure, Fortin D. and Beveridge T. J. (1997) Role of the bacterium Thiobacillus Properties, Reactions, Occurrence and Uses. VCH, Weinheim, in the formation of silicates in acidic mine tailings. Chem. Geol. Germany. 141, 235–250. APPLIED AND ENVIRONMENTAL MICROBIOLOGY, Dec. 2005, p. 7679–7689 Vol. 71, No. 12 0099-2240/05/$08.00ϩ0 doi:10.1128/AEM.71.12.7679–7689.2005 Copyright © 2005, American Society for Microbiology. All Rights Reserved.

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

Received 26 November 2004/Accepted 24 July 2005 Downloaded from

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 http://aem.asm.org/ 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) on August 24, 2014 by UNIV OF NOTRE DAME remained. In the case of xylene, this was less clear.

The use and release of heavy metals to air, water, and soils unclear since no studies have addressed this point. If species has created a significant number of contaminated sites across richness is reduced in sites contaminated with complex mix- the United States and the world. Thus, the effect of metal tures, the communities may be less resilient because the prob- contamination on the microbial community has been exten- ability that an ecotype is capable of a specific required function sively studied over the past several decades. The acute effects is reduced (13). Previous research in our laboratory showed of short-term exposure to toxic heavy metals upon a broad that the microbial community structure in a long-term mixed- array of microbial processes have been well documented (9, 10, waste contaminated site might reflect both metal and aromatic 21, 40, 47). More recently, investigators have examined habi- hydrocarbon concentrations in the soil. For individual mi- tats exposed to anthropogenic or natural metal contamination crobes to persist under complex conditions, they must tolerate over an extended period of time (5, 20, 22, 24, 39, 45). Studies both local metal and hydrocarbon contaminants. Shi et al. (42) that focused on the culturable fraction of the microbial com- found a very broad distribution of metal tolerances within the munity indicated that as few as 10 to almost 100% of the microbial communities in these soils. This is consistent with a bacteria in habitats contaminated for extended periods were heterogeneous distribution of microbes in both highly contam- metal resistant. Thus, there may be substantial variability in inated and noncontaminated microsites. community responses to metal exposure between locations. As For this study, we used microcosms to segregate the effects non-cultivation-based methods have become available, re- of two metals, Pb and Cr, and also to study the impact of searchers have begun examining the impact of metal exposure aromatic hydrocarbons on microbial community structure and on the entire indigenous community (6, 24, 25, 26) and have tried to address the impact of these exposures on community activity. By using microcosms, soils could be homogenized to diversity (30) and resiliency (15). evenly distribute both the microbial populations and toxicants A confounding factor in metal-contaminated sites is the fre- and thereby reduce spatial variability. This allowed us to test 2ϩ 6ϩ quent co-occurrence of organic contaminants. These organic experimental additions of Pb or Cr and to assess their molecules may be metabolizable energy sources, toxicants, or effects on the microbial community. One of two energy sub- both. The combined effect of metals and organic carbon pol- strates (glucose or xylene) was added to provide the necessary lutants on microbial activity and community composition is 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 * Corresponding author. Mailing address: Department of Agron- omy, Purdue University, West Lafayette, IN 47907-2054. Phone: (765) changes in community activity were related to molecular anal- 496-2997. Fax: (765) 496-2629. E-mail: [email protected]. yses of community composition and functional gene levels.

7679 VOL. 71, 2005 ORGANIC CARBON AND METALS 7687 more sweeping replacements of community members in re- centrations than those required when glucose was the added sponse to metal additions than did xylene-amended ones. The energy source. As noted above, this difference may reflect the number of different organic substrates and metals tested was narrower pool of metal-tolerant species that could catabolize small; however, this result may reflect the distinction between xylene than of those that could catabolize glucose. cases where there are large numbers of organisms that might A final physiological consequence of metal addition was the be recruited (glucose catabolizers) and those (xylene) where reduction in net biomass increases at the expense of added potential responders (and their levels of metal resistance) are organic carbon. As noted above, growth dynamics were still more limited. The Cr(VI) levels were different for each sub- present to produce changes in DGGE fingerprints, culturable strate and were titrated to reduce catabolism by a predeter- bacterial numbers, and catabolic gene copies. However, the mined value; the amount added to glucose microcosms to lack of biomass increase may also reflect an energy expenditure reduce activity in the glucose microcosms by 90% [4 mg gϪ1of to implement metal tolerance, as is true in the case of ATP- Downloaded from Cr(VI)] resulted in no metabolic activity in microcosms with dependent efflux systems (43). The ratio of mineralized to xylene. assimilated organic C was found to increase in other metal- This experimental approach allowed us to identify some of contaminated soils (6), which is consistent with the increased the indigenous soil populations that responded to these con- energy expenditure under these conditions. ditions. Intense bands generally only occurred in 16S rRNA The BPH4-type biphenyl dioxygenase and PHE genes were PCR-DGGE profiles for microcosms in which carbon miner- routinely detected in control microcosms and most xylene- alization occurred. We believe these intense bands correspond amended microcosms, rather than the xylene monooxygenase to organisms that have multiplied at the expense of the added and toluate or benzoate dioxygenases usually associated with http://aem.asm.org/ organic substrate and are present as a significant fraction of the xylene metabolism (The University of Minnesota Biocatalysis/ total population. This argument is strengthened by the signif- Biodegradation Database [http://umbbd.ahc.umn.edu/]). The icantly higher copy numbers of the catabolic BPH4 and PHE presence of aromatic oxygenase genes in the control micro- genes in the microcosms with xylene addition. Furthermore, cosms was not unexpected because the soil sample was taken xylene-degrading bacteria that were isolated from the micro- from a site contaminated with aromatic hydrocarbons. BPH4 cosms contained BPH4 or PHE genes (Beasley et al., in prep- and PHE gene copies increased in xylene microcosms and aration) and corresponded to intense bands in the community corresponded to changes in CO2 evolution, and in some cases, PCR-DGGE profiles. Other approaches, such as heavy isotope biomass; this suggests that there was an enrichment of strains on August 24, 2014 by UNIV OF NOTRE DAME additions (29, 35), have been suggested as an approach to link harboring these oxygenase genes. Phenol hydroxylase catalyzes populations responsible for functions in soil communities, but the continued oxidation of hydroxylated intermediates in xy- the approach used here is much simpler and less expensive and lene and toluene catabolism (3) and has been detected in other can be used on a broader scale in the field. The combination of xylene-amended-microcosm studies (B. Baldwin, personal PCR-DGGE of 16S rRNA genes with quantitative PCR of communication). The selection of BPH4 oxygenase genes fol- functional genes does have limitations because an individual lowing the addition of xylenes may seem counterintuitive; how- band (organism) is not unequivocally linked to a specific func- ever, previous reports have noted the sequence similarity and tion and because not all gene variants for specific catabolic functional overlap of biphenyl and alkyl-benzene dioxygenases, functions are known. However, at the field scale, the correla- including toluene dioxygenase (17, 44). The BPH4 subfamily of tion between stimulation of specific rRNA gene sequences and biphenyl dioxygenase genes, in particular, is closely related to specific functional genes does provide a good initial basis to isopropylbenzene dioxygenase genes. stimulate detailed analyses of the consequences of changes in Although Cr(VI) inhibits microbes, there are biotic and abi- community structure upon community function. otic detoxification mechanisms in soil (4). One biotic mecha- The phylogenetic identity of many of the bacteria selected nism (under both aerobic and anaerobic conditions) occurs via under these conditions is consistent with in situ analyses of Cr(VI) reduction to less toxic and less mobile Cr(III) (14, 48). these (21a) and other metal-contaminated soils. The Actino- Cr(VI) reduction to nontoxic levels could have been a precon- mycetales have been reported to be important in metal-im- dition for the onset of microbial activity in the microcosms. pacted soils (18). In cases where glucose only was added or However, this was not the case in systems to which glucose was xylene plus Cr(VI) was added, organisms from several genera added; these could tolerate relatively large additions of of the Actinomycetales commonly responded to the stimuli. Cr(VI), and microbes were mineralizing glucose when most of Although the primary objective of this study was to analyze the Cr(VI) remained. In the case of xylene amendments, the changes in community composition, activities were monitored relationship is less clear. Because only low levels of Cr(VI) to provide some basis for understanding community dynamics. could be added to retain any xylene degradation at all, even a In general, activity responses were similar to what has been modest rate of biological reduction resulted in very small re- observed in a number of other microcosm studies (8, 19, 37, sidual Cr(VI) concentrations. Coupled with the lag in detect- 38). In the absence of metals, carbon additions stimulated able xylene mineralization even in the absence of Cr(VI), these increases in carbon mineralization, microbial biomass, and the data do not convincingly show that xylene catabolism occurred number of culturable bacteria. Heavy metals generally sup- while Cr(VI) was present. More sensitive analyses (for exam- pressed these community responses, as lower carbon mineral- ple, with radiotracers) would be required to resolve this issue. ization rates and longer lag phases were observed. On a molar Rarefaction analysis of a 16S rRNA gene clone library cre- basis, Cr was more toxic than Pb; this may reflect its greater ated from several soils at this site indicated that they contain a mobility and bioavailability in soil (34). In addition, heavy relatively low diversity of microbes (21a), and the provision of metals inhibited the xylene-degrading community at lower con- a single carbon source in microcosms produced a further re- 7688 NAKATSU ET AL. APPL.ENVIRON.MICROBIOL. duction in diversity. These effects are explicable in light of the siliency differences in hyporheic microbial communities in response to fluvial resource heterogeneity hypothesis (46), i.e., the site is rela- heavy-metal deposition. Appl. Environ. Microbiol. 70:4756–4765. 16. Findlay, R. H. 1996. The use of phospholipid fatty acids to determine mi- tively uniformly barren with respect to resource availability due crobial community structure, p. 1–17. In A. K. Akkermans, J. D. van Elsas, to the lack of plant vegetation. As a result, heterotrophic pro- and F. de Bruijn (ed.), Molecular microbial ecology manual. Kluwer Aca- demic Publishers, Dordrecht, The Netherlands. ductivity and diversity are low. The addition of a single C 17. Furukawa, K., J. Hirose, A. Suyama, T. Zaiki, and S. Hayashida. 1993. Gene source to microcosms results in even lower organic resource components responsible for discrete substrate specificity in the metabolism heterogeneity, and there is strong selection for a few types that of biphenyl (bph operon) and toluene (tod operon). J. Bacteriol. 175:5224– 5232. use that dominant C source. When an additional selection 18. Gremion, F., A. Chatzinotas, and H. Harms. 2003. Comparative 16S rDNA (heavy metals) factor was imposed, the community dynamics and 16S rRNA sequence analysis indicates that Actinobacteria might be a were found to be much greater for components expected to dominant part of the metabolically active bacteria in heavy metal-contaminated bulk and rhizosphere soil. Environ. Microbiol. 5:896–907. Downloaded from contain a large degree of functional redundancy (glucose-ca- 19. Gremion, F., A. Chatzinotas, K. Kaufmann, W. V. Sigler, and H. Harms. tabolizing bacteria) than a more restricted catabolic function 2004. Impacts of heavy metal contamination and phytoremediation on a microbial community during a twelve-month microcosm experiment. FEMS (xylene degradation). Consistent with this broader capacity was Microbiol. Ecol. 48:273–283. the greater robustness of glucose catabolism, that is, it recov- 20. Hutchinson, T. C., and M. S. Symington. 1997. Persistence of metal stress in ered and proceeded at much higher metal concentrations than a forested ecosystem near Sudbury, 66 years after closure of the O’Donnell roast bed. J. Geochem. Explor. 58:323–330. did xylene catabolism. 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– ACKNOWLEDGMENTS 1011. 21a.Joynt, J., M. Bischoff, R. Turco, A. Konopka, and C. H. Nakatsu. Microbial http://aem.asm.org/ This work was supported by a grant from the DOE Natural and community analysis of soils contaminated with lead, chromium and petro- Accelerated Bioremediation Research (NABIR) program (grant DE- leum hydrocarbons. Microb. Ecol., in press. FG02-98ER62681). 22. Kamaludeen, S. P. B., M. Megharaj, R. Naidu, I. Singleton, A. L. Juhasz, We thank the Indiana Department of Transportation, and in par- B. G. Hawke, and N. Sethunathan. 2003. Microbial activity and phospholipid ticular Bill Jervis, for giving us access to the site, Judy Lindell for fatty acid pattern in long-term tannery waste-contaminated soil. Ecotox. technical assistance, and Leon Toussaint and Joanne Becker for as- Environ. Safety 56:302–310. sisting in soil collection. 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. REFERENCES 24. Konopka, A., T. Zakharova, M. Bischoff, L. Oliver, C. Nakatsu, and R. F. Turco. 1. Acosta-Martıńez, V., Z. Reicher, M. Bischoff, and R. F. Turco. 1999. The role 1999. Microbial biomass and activity in lead-contaminated soil. Appl. on August 24, 2014 by UNIV OF NOTRE DAME of tree leaf mulch and nitrogen fertilizer on turfgrass soil quality. Biol. Fert. Environ. Microbiol. 65:2256–2259. Soils 29:55–61. 25. Konstantinidis, K. T., N. Isaacs, J. Fett, S. Simpson, D. T. Long, and T. L. 2. Altschul, S. F., T. L. Madden, A. A. Schaffer, J. Zhang, Z. Zhang, W. Miller, Marsh. 2003. Microbial diversity and resistance to copper in metal-contam- and D. J. Lipman. 1997. Gapped BLAST and PSI-BLAST: a new generation inated lake sediment. Microb. Ecol. 45:191–202. of protein database search programs. Nucleic Acids Res. 25:3389–3402. 26. Kozdroj, J., and J. D. van Elsas. 2001. Structural diversity of microorganisms 3. Arenghi, F. L. G., D. Berlanda, E. Galli, G. Sello, and P. Barbieri. 2001. in chemically perturbed soil assessed by molecular and cytochemical ap- Organization and regulation of meta cleavage pathway genes for toluene and proaches. J. Microbiol. Methods 43:197–212. o-xylene derivative degradation in Pseudomonas stutzeri OX1. Appl. Environ. 27. Lane, D. J. 1991. 16S/23S rRNA sequencing, p. 115–175. In E. Stackebrandt Microbiol. 67:3304–3308. and M. Goodfellow (ed.), Nucleic acid techniques in bacterial systematics. 4. Avudainayagam, S., A. Megharaj, G. Owens, R. S. Kookana, D. Chittleborough, John Wiley & Sons, New York, N.Y. and R. Naid. 2003. Chemistry of chromium in soils with emphasis on tannery 28. LaPara, T. M., C. H. Nakatsu, L. Pantea, and J. E. Alleman. 2000. Phylo- waste sites. Rev. Environ. Contam. Toxicol. 178:53–91. genetic analysis of bacterial communities in mesophilic and thermophilic 5. Baath, E. 1992. Measurement of heavy metal tolerance of soil bacteria using bioreactors treating pharmaceutical wastewater. Appl. Environ. Microbiol. thymidine incorporation into bacteria extracted after homogenization-cen- 66:3951–3959. trifugation. Soil Biol. Biochem. 24:1167–1172. 29. Manefield, M., A. S. Whiteley, R. I. Griffiths, and M. J. Bailey. 2002. RNA 6. Baath, E., M. Diaz-Ravina, A. Frostegard, and C. D. Campbell. 1998. Effect stable isotope probing, a novel means of linking microbial community func- of metal-rich sludge amendments on the soil microbial community. Appl. tion to phylogeny. Appl. Environ. Microbiol. 68:5367–5373. Environ. Microbiol. 64:238–245. 30. Moffett, B. F., F. A. Nicholson, N. C. Uwakwe, B. J. Chambers, J. A. Harris, 7. Baldwin, B. R., C. H. Nakatsu, and L. Nies. 2003. Detection and enumera- and T. C. J. Hill. 2003. Zinc contamination decreases the bacterial diversity tion of aromatic oxygenase genes by multiplex and real-time PCR. Appl. of agricultural soil. FEMS Microbiol. Ecol. 43:13–19. Environ. Microbiol. 69:3350–3358. 31. Muyzer, G., E. C. de Waal, and A. G. Uitterlinden. 1993. 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Wiley, New York, N.Y. for examination of water and wastewater, 20th ed. American Public Health 35. Padmanabhan, P., S. Padmanabhan, C. DeRito, A. Gray, D. Gannon, J. R. Association, Washington, D.C. Snape, C. S. Tsai, W. Park, C. Jeon, and E. L. Madsen. 2003. Respiration of 13. Ekschmitt, K., and B. S. Griffiths. 1998. Soil biodiversity and its implications 13C-labeled substrates added to soil in the field and subsequent 16S rRNA for ecosystem functioning in a heterogeneous and variable environment. gene analysis of 13C-labeled soil DNA. Appl. Environ. Microbiol. 69:1614– Appl. Soil Ecol. 10:201–215. 1622. 14. Fein, J. B., D. A. Fowle, J. Cahill, K. Kemner, M. Boyanov, and B. Bunker. 36. Pielou, E. C. 1969. An introduction to mathematical ecology. Wiley-Inter- 2002. Nonmetabolic reduction of Cr(VI) by bacterial surfaces under nutri- science, New York, N.Y. ent-absent conditions. Geomicrobiol. J. 19:369–382. 37. Rajapaksha, R. M. C. P., M. A. Tobor-Kaplon, and E. Baath. 2004. Metal 15. Feris, K. P., P. W. Ramsey, M. Rillig, J. N. Moore, J. E. Gannon, and W. E. toxicity affects fungal and bacterial activities in soil differently. Appl. Envi- Holbert. 2004. Determining rates of change and evaluating group-level re- ron. Microbiol. 70:2966–2973. 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

Drew Gorman-Lewis, Ph.D.

______Jeremy B. Fein, Director

Graduate Program in Civil Engineering and Geological Sciences

Notre Dame, Indiana

November 2005

Drew Gorman-Lewis

2005

THE ADSORPTION OF CHEMICAL

CONTAMINANTS ONTO GEOLOGIC SURFACES WITH SPECIAL

CONSIDERATION OF THE BACTERIAL SURFACE

Abstract

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 site- specific 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

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

n corr Figure 3.4. − ∑Q x 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

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.

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

+4 + reduction leads to enhanced adsorption of Np 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

69 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 metal- bacteria 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.

123 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 low-pH conditions and subsequent adjustment of solution pH to near neutral ranges to precipitate Cr(III) as hydroxides Chromium is one of the most toxic heavy metals dis- [5]. However, this method is not completely satisfactory charged into the environment through various industrial because of the large amount of secondary waste products due wastewaters, and has become a serious health problem. Metal to various reagents used in the above-mentioned processes. plating, tanneries and industrial processes using catalysts Biological treatments arouse great interest because of their discharge worldwide huge amounts of chromium every lower impact on the environment as opposed to chemical year. The efﬂuents of these industries contain Cr(VI) and treatments. Recent studies have shown that certain species of Cr(III) at concentrations ranging from tenths to hundreds bacteria are capable of transforming hexavalent chromium, of milligrams/liter. While Cr(VI) is highly toxic and is Cr(VI), into the much less toxic and less mobile trivalent known to be carcinogenic and mutagenic to living organisms form, Cr(III) [6,7]. Bacteria may protect themselves from [1], Cr(III) is generally only toxic to plants at very high toxic substances in the environment by transforming toxic concentrations and is less toxic or non-toxic to animals [2]. compounds through oxidation, reduction or methylation into The discharge of Cr(VI) to surface water is regulated to more volatile, less toxic or readily precipitating forms. below 0.05 mg/l by the US EPA [3] and the European Union The processes by which microorganisms interact with [4], while total Cr, including Cr(III), Cr(VI) and its other toxic metals enabling their removal/and recovery are forms, is regulated to below 2 mg/l [3]. bioaccumulation, biosorption and enzymatic reduction At present, the most commonly used technology for [8]. Microbial heavy metal accumulation often comprises treatment of heavy metals in wastewaters is chemical of two phases. An initial rapid phase involving physical precipitation. Conventional chemical treatment involves adsorption or ion exchange at cell surface and by a subse- reduction of Cr(VI) to Cr(III) by a reducing agent under quent slower phase involving active metabolism-dependent transport of metal into bacterial cells [9]. Biosorption is a ∗ Corresponding author. Tel.: +30 26410 39517; fax: +30 26410 39576. metabolism-independent process and thus can be performed E-mail address: [email protected] (D.V. Vayenas). by both living and dead microorganisms. This adsorption is

0304-3894/$ – see front matter © 2005 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2005.06.008 E. Dermou et al. / Journal of Hazardous Materials B126 (2005) 78–85 79 based on mechanisms such as complexation, ion exchange, 7H2O, 0.001 g CaCl2·2H2O, 5 g CH3COONa·3H2O and coordination, adsorption, chelation and microprecipitation, 0.5 g K2HPO4 in 1.0 l of tap water. which may be synergistically or independently involved [10]. Enzymatic reduction of Cr(VI) into Cr(III) is believed to be 2.2. Reagents one of the defense mechanisms employed by microorganisms living in Cr(VI)-contaminated environments. The reduced Stock Cr(VI) solution (500 mg/l) was prepared by dissolv- ◦ Cr(III) may precipitate as chromium hydroxide in neutral pH ing 141.4 mg of 99.5% K2Cr2O7, previously dried at 103 C range [11]. for 2 h, in Milli-Q water and diluting to 100 ml. Diphenyl Most of the previous studies on biological reduction of carbazide solution was prepared by dissolving 250 mg of Cr(VI) were conducted in batch reactors (flasks) using mainly 1,5-diphenylcarbazide in 50 ml of HPLC-grade acetone and pure cultures. For instance, Wang and Xiao [12] studied storing in a brown bottle. Potassium hydrogen phthalate stan- several factors affecting hexavalent chromium reduction in dard (KHP) was prepared by dissolving 425 mg in distilled pure cultures of bacteria in flasks. Wang and Shen [5] studied water and diluting to 1000 ml. Digestion solution was pre- the kinetics of Cr(VI) reduction by pure bacterial cultures pared by dissolving 10.216 g K2Cr2O7, previously dried at ◦ in flasks. Shakoori et al. [13] isolated a dichromate-resistant 103 C for 2 h, in 500 ml distilled water, 167 ml conc. H2SO4 Gram-positive bacterium from effluent of tanneries and used and 33.3 g HgSO4 and diluting to 1000 ml (for the determi- flasks as batch reactors. Fein et al. [14] used pure bacterial nation of the COD values). cultures in flasks to study the non-metabolic reduction of 1,5-Diphenylcarbazide was purchased from Fluka Chem- Cr(VI) by bacteria under nutrient-absent conditions. Srinath ical, potassium dichromate was purchased from Sigma et al. [8] studied Cr(VI) biosorption and bioaccumulation Chemical Co. All the others chemicals were purchased from by pure cultures of chromate resistant bacteria in flasks. Riedel-de Haen. Megharaj et al. [15] studied hexavalent chromium reduction in flasks, by pure cultures of bacteria isolated from soil 2.3. Analytical methods contaminated with tannery waste. Recently, continuous-flow and fixed-film bioreactors During all experiments, hexavalent chromium concen- were also used for biological reduction of Cr(VI). Shen and tration, pH, temperature, dissolved oxygen concentration Wang [16] demonstrated Cr(VI) reduction in a two-stage, and TOC measurements were made on a daily basis. continuous-flow suspended growth bioreactor system. Samples were filtered through 0.45 ␮m –Millipore filters Escherichia coli cells grown in the first-stage completely (GN-6 Metricel Grid 47 mm, Pall Corporation). Hexavalent mixed reactor were pumped into the second-stage plug-flow chromium concentration was determined by the 3500-Cr D reactor to reduce Cr(VI). Chirwa and Wang [11] demon- Colorimetric method according to Standard Methods for the strated the potential of fixed-film bioreactors for Cr(VI) Examination of Water and Wastewater [17]. Total organic reduction. This was the first report on Cr(VI) reduction carbon measurements (TOC) were conducted in order to through biological mechanisms in a continuous-flow determine the feed sodium acetate concentration both in laboratory-scale biofilm reactor without the need to con- the liquid culture (chemostat) and the liquid volume of the stantly resupply fresh Cr(VI)-reducing cells. Bacillus sp. was bioreactor, following the methods described in Standard used in this work for the transformation of Cr(VI) into Cr(III). Methods for the Examination of Water and Wastewater Virtually all the previous studies on biological reduction [17] by using, Total organic carbon analyzer (TOC-VCSH, of Cr(VI) were conducted in laboratory scale apparatus SHIMAZDU Corporation, Japan). Total chromium con- (reactors), using sterilized conditions and pure cultures of centration measurements were made according to Standard microorganisms. The present study is the first to report on Methods for the Examination of Water and Wastewater [17] Cr(VI) biological reduction in a pilot-scale trickling filter using an atomic absorption spectrophotometer (model AAS- using mixed culture of microorganisms, originating from 700, Perkin-elmer) (results not shown for total chromium an industrial sludge. The operation of the trickling filter as concentrations). a Sequencing Batch Reactor (SBR) with recirculation led to significantly high Cr(VI) reduction rates, thus promising 2.4. Isolation and enrichment of indigenous bacteria a feasible technological solution to a serious environmental problem. 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 2. Materials and methods was added in a 2 l Erlenmeyer flask and was diluted in an acetate-minimal medium and concentrated chromium solu- 2.1. Media tion (in the form of K2Cr2O7) resulting in a final hexavalent chromium concentration of 50 mg/l. The final volume of the The influent feed to the bioreactor was prepared by solution was 1 l. Acetate-minimal medium (AMM) was com- dissolving 1 g NH4Cl, 0.2 g MgSO4·7H2O, 0.001 g FeSO4· prising (per litre) 1 g NH4Cl, 0.2 g MgSO4·7H2O, 0.001 g E. Dermou et al. / Journal of Hazardous Materials B126 (2005) 78–85 85

[14] J.B. Fein, D.A. Fowle, J. Cahill, K. Kemner, M. Boyanov, B. [18] J.R. Marchesi, T. Sato, A.J. Weightman, T.A. Martin, J.C. Fry, Bunker, Nonmetabolic reduction of Cr(VI) by bacterial surfaces S.J. Hiom, W.G. Wade, Design and evaluation of useful bacterium- under nutrient-absent conditions, Geomicrobiol. J. 19 (2002) 369– speciﬁc PCR primers that amplify genes coding for bacterial 16S 382. rRNA, Appl. Environ. Microbiol. 64 (1998) 795–799. [15] M. Megharaj, S. Avudainayagam, R. Naidu, Toxicity of hexava- [19] S.F. Altschul, T.L. Maden, A.A. Schaffer, J. Zhang, Z. Zhang, W. lent chromium and its reduction by bacteria isolated from soil Miller, D.J. Lipman, Gapped BLAST and PSI-BLAST: a new gen- contaminated with tannery waste, Curr. Microbiol. 47 (2003) 51– eration of protein database search programs, Nucl. Acids Res. 25 54. (1997) 3389–3402. [16] H. Shen, Y. Wang, Hexavalent chromium removal in two-stage biore- [20] P. Pattanapipitpaisal, N.L. Brown, L.E. Macaskie, Chromate reduc- actor system, J. Environ. Eng. 121 (11) (1995) 798–804. tion and 16S rRNA identiﬁcation of bacteria isolated from a Cr(VI)- [17] APHA, AWWA and WPCF, Standard Methods for the Examina- contaminated site, Appl. Microbiol. Biotechnol. 57 (2001) 257–261. tion of Water and Wastewater, 17th ed., American Public Health [21] R. Francisco, M.C. Alpoim, P.V. Morais, Diversity of chromium- Association, American Water Works Association and Water Pollu- resistant and -reducing bacteria in a chromium-contaminated sludge, tion Control Federation, Washington, DC, 1989. 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 Gram- negative 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 function of different environmental variables (e.g., pH, Bacteria are ubiquitous in near-surface geological solution and substrate composition), is responsible for most systems, and are known to play important role in different surface interactions of bacteria. While the bulk chemical biogeochemical processes, including contaminant trans- composition of bacterial cell walls is often known, their port and degradation,1,2 mineral dissolution and precipi- ability to complex metals and attach to surfaces as a tation,3 and metal sorption by minerals and their redox function of different environmental conditions is not well transformations.4-9 Bacteria-water interfacial chemistry understood. is one of the critical variables that play a central role in The Gram-positive cell wall is primarily made up of mediating these bacterial reactions. In addition, bacterial peptidoglycan (ca. 40-80% of the dry weight of the wall), transport through porous media, adhesion to minerals which is a polymer of N-acetylglucosamine and N- and biological tissue, response to antibiotics, and the acetylmuramic acid, containing mainly carboxyl, amide, formation and chemistry of biofilms are also modified by and hydroxyl functional groups.15 The two other important the bacterial surface chemistry.10-14 The composition and constituents of Gram-positive cell walls are teichoic acid, structure of bacterial cell walls, and their variation as a a polymer of glycopyranosyl glycerol phosphate, and teichuronic acid, which is similar to teichoic acid, but * Corresponding author. Telephone: (609) 258-3827. E-mail: replaces the phosphate functional groups with carboxyls. [email protected]. The cell walls of Gram-negative bacteria are more complex † Princeton University. ‡ University of Guelph. due to the presence of an outer membrane in addition to § Lawrence Berkeley National Laboratory. a thin peptidoglycan layer, but do not contain teichoic or (1) Corapcioglu, M. Y.; Kims. Water Resour. Res. 1995, 31, 2639- teichuronic acids.16 Instead, the outer membrane contains 2648. phospholipids, lipoproteins, lipopolysaccharides, and pro- (2) Watanabe, K.; Hamamura, N. Curr. Opin. Biotechnol. 2003, 14, 289-295. teins. (3) Stillings, L. L.; Drever, J. I.; Brantley, S. L.; Sun, Y.; Oxburgh, Several recent investigations examined the sur- R. Chem. Geol. 1996, 132,79-90. face chemistry of intact bacterial cells and their cell (4) Berveridge, T. J.; Forsberg, C. W.; Doyle, R. J. J. Bacteriol. 1982, 150, 1438-1448. walls using both macroscopic (e.g., potentiometric titra- (5) Suzuki, T.; Miyata, H.; Kawai, K.; Takmizawa, K.; Tai, Y.; Okazaki, tion, ion adsorption) and molecular tools (microscopy and M. J. Bacteriol. 1992, 174, 5340-5345. (6) Jackson, T. A.; West, M. M.; Leppard, G. G. Environ. Sci. Technol. 1999, 33, 3795-3801. (11) Burdman, S.; Okon, Y.; Jurkevitch, E. Crit. Rev. Microbiol. 2000, (7) Holman, H. N.; Perry, D. L.; Martin, M. C.; Lamble, G. M.; 26,91-110. McKinney, W. R.; Hunter-Cevera, J. C. Geomicrobiol. J. 1999, 16, 307- (12) Nelson, Y. M.; Lion, L. W.; Shuler, M. L.; Ghiorse, W. C. Environ. 324. Sci. Technol. 1996, 30, 2027-2035. (8) Klaus-Joerger, T.; Joerger, R.; Olsson, E.; Granqvist, C. G. Trends (13) Decho, A. W. Continental Shelf Res. 2000, 20, 1257-1273. Biotechnol. 2001, 19,15-20. (14) Suci, P. A.; Geesey, G. G.; Tyler, B. J. J. Microbiol. Methods (9) Fein, J. B.; Fowle, D. A.; Cahill, J.; Kemner, K.; Boyanov, M.; 2001, 46, 193-208. Bunker, B. Geomicrobiol. J. 2002, 19, 369-382. (15) Berveridge, T. J. Int. Rev. Cytol. 1981, 72, 229-317. (10) McWhirter, M. J.; Bremer, P. J.; Lamont, I. L.; McQuillan, A. (16) Perry, J. J.; Staley, J. T.; Lory, S. Microbial Life; Sinauer J. Langmuir 2003, 19, 3575-3577. Associates, Inc.: Sunderland, MA, 2002; pp 61-100.

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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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Treatment Technologies for Chromium(VI) 279

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 con-

tained 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 inﬂu- ence 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, rhizodegrada- tion (also called phytostimulation or plant-assisted bioremediation), and

L1608_C08.fm Page 284 Friday, July 23, 2004 5:46 PM

284 Chromium(VI) Handbook

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, 3– a nonspeciﬁc method of ion transport for anions such as phosphate (PO4 ) 2– 2– and SO4 (Cervantes et al., 2001). Cr is toxic to yeast because it inhibits SO4 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 L1608_C08.fm Page 303 Friday, July 23, 2004 5:46 PM

Treatment Technologies for Chromium(VI) 303

remediation strategies combine multiple technologies and mechanisms. For example, toxicity reduction occurs when Cr(VI) is reduced to Cr(III). How- ever, 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 elec- trodes 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

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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, Colo- rado, 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, Tech- nology 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 Process- es, 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 nanofiltra- tion, 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 1Department of Civil and Environmental Engineering, Stanford University, Stanford, California 943051; telephone: (650) 723-9032; fax: (650) 725-3164; e-mail: [email protected]. 2Marine Biology Research Division and Center for Marine Biotechnology and Biomedicine, Scripps Institution of Oceanography, University of California—San Diego, La Jolla, California 92093-02022 3National 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 denitrifying cells exposed to Cr(VI) showed reduced chromium remediation, but the effects of competing chromium precipitates both extracellularly on the cell electron acceptors are still poorly understood. We inves- surface and, for the first time, as electron-dense round tigated chromate (Cr(VI)) reduction in batch cultures of globules inside cells. © 2003 Wiley Periodicals, Inc. Biotech- Shewanella oneidensis MR-1 under aerobic and denitri- nol Bioeng 83: 627–637, 2003. fying conditions and in the absence of an additional elec- Keywords: Shewanella oneidensis MR-1; chromate; re- tron acceptor. Growth and Cr(VI) removal patterns sug- duction; denitrification; inhibition gested 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 INTRODUCTION 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 Hexavalent chromium [Cr(VI)] is an Environmental Protec- µmoles) Cr(VI) reduced и [mg protein]−1 for high levels of tion Agency (EPA) priority pollutant and a known carcino- Cr(VI) added as a single spike. For low levels of Cr(VI) gen. Due to its widespread industrial use, it is often found in added sequentially, Tc increased to 3.33 mg (64 µmoles) contaminated groundwater. In order to achieve levels of Cr(VI) reduced и [mg protein]−1, indicating that it is limited by toxicity at higher concentrations. During denitri- chromium below the EPA maximum contaminant level (100 fication and aerobic growth, MR-1 reduced Cr(VI), with ␮g/L), remediation strategies focus on the reduction of much faster rates under denitrifying conditions. Cr(VI) Cr(VI) to insoluble trivalent forms, which are relatively had no effect on nitrate reduction at 6 µM, was strongly stable and nontoxic. The only compounds able to oxidize inhibitory at 45 µM, and stopped nitrate reduction above trivalent chromium at any appreciable rate are manganese 200 µM. Cr(VI) had no effect on aerobic growth at 60 µM, but severely inhibited growth above 150 µM. A factor oxides (Eary and Rai, 1987). Bioremediation may be effec- that likely plays a role in Cr(VI) toxicity is intracellular tive for the removal of Cr(VI) from groundwater, as many reduced chromium. Transmission electron microscopy aerobic and anaerobic microorganisms reduce Cr(VI) to (TEM) and electron energy loss spectroscopy (EELS) of 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 Correspondence to: Craig S. Criddle Contract grant sponsor: the National Institute of Environmental Health few studies that compare the kinetics of Cr(VI) reduction by Sciences (NIEHS) Superfund Basic Research Program bacteria under different electron-accepting conditions or Contract grant numbers: ES04911 (to the Institute for Environmental that study the effects of Cr(VI) on the reduction of other Toxicology at Michigan State University), ES10337 to the University of electron acceptors. California San Diego) Shewanella oneidensis MR-1 (formerly Shewanella pu- Contract grant sponsor: the National Science Foundation Graduate Fel- lowship Program (to SM) trefaciens MR-1) is a facultative Gram-negative bacterium Contract grant sponsor: the Swiss National Science Foundation Post whose respiratory versatility has prompted interest in its use Graduate Fellowship (to RBL). in bioremediation. A nonfermenting bacterium, MR-1 can

© 2003 Wiley Periodicals, Inc. RESULTS AND DISCUSSION reduction will precipitate as insoluble Cr(III) hydroxides (Jardine et al., 1999). However, Cr(III) is known to complex Cr(VI) Reduction Without Additional Electron with organic ligands (Nieboer and Jusys, 1988). Soluble Acceptors Present Cr(III) is commonly computed as the difference between total soluble Cr and total soluble Cr(VI). In medium con- The capacity of MR-1 to reduce Cr(VI) in the absence of taining lactate as electron donor, 13–14% of the reduced other electron acceptors was investigated. All experiments Cr(III) remained soluble, presumably as a complex to lac- were performed in the minimal medium defined in the pre- tate or components of yeast extract or bactopeptone. When vious section. MR-1 reduced Cr(VI) with lactate as the elec- 100 mM HEPES was included in the medium as a buffer, tron donor (Fig. 1) but not formate (data not shown). Con- the percentage of soluble Cr(III) increased to 47%. This trols prepared without cells, without lactate, and with auto- observation is important for bioremediation of Cr(VI) be- claved cells showed little decrease in Cr(VI) concentration cause the EPA maximum contaminant level (MCL) is mea- ␮ (approximately 20, 36, and 20 M, respectively, over 10 sured in terms of total soluble Cr, including both soluble days). Abiotic reduction of Cr(VI) could have resulted from Cr(VI) and Cr(III). 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 do- Cr(VI) Reduction During Denitrification

nor. Cr(VI) reduction in the absence of electron donor has MR-1 can reduce nitrate to nitrite to N2O coupled to growth also been described by Fein et al. (2002). As expected for an (Krause and Nealson, 1997). Batch studies were performed 2− anion such as CrO4 , adsorption was not observed at this in minimal medium to study simultaneous Cr(VI) and ni- pH. Cr(VI) reduction in autoclaved controls did not differ trate reduction. Cultures were spiked with 5 mM lactate and from abiotic controls. ∼2 mM nitrate and allowed to grow for several hours to MR-1 cultures were unable to grow during the reduction ensure the onset of nitrate reduction. Cr(VI) was then added of Cr(VI), as indicated by viable cell count (Fig. 1) and total at ∼6 and 45 ␮M to a subset of the bottles. As shown in cellular protein (data not shown). Colony-forming units did Figure 2, in the absence of Cr(VI) lactate was oxidized to not increase after the addition of almost 175 ␮M Cr(VI). In acetate, nitrate was reduced to nitrite, and nitrite was almost

addition, the specific activity of Cr(VI) reduction decreased completely reduced (presumably to N2O). When Cr(VI) was with each addition of Cr(VI). If Cr(VI) reduction supported added at 6 ␮M, Cr(VI) was reduced during nitrate and ni- growth, the rate of reaction would increase. These results trite reduction. Patterns of lactate oxidation, nitrate and ni- suggest secondary utilization or cometabolism as the trite reduction, and growth were nearly identical to controls mechanism for Cr(VI) reduction. 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 Cr(III) Solubilization the MCL (2 ␮M). Controls without lactate showed no ni- Total soluble chromium was measured at the end of experi- trate reduction (data not shown). No Cr (VI) reduction was ments in order to probe for the presence of soluble Cr(III). observed in abiotic controls containing nitrate, nitrite, and At the pH of most groundwater, Cr(III) formed from Cr(VI) 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 ␮MCr(VI)wasaddedatday2and5.(ࡗ), MR-1 + lactate + Cr(VI); (᭝), MR-1 + Cr(VI); (᭹), Cr(VI) + lactate; (᭺), autoclaved MR-1 + lactate + Cr(VI); and (*), viable cell count for MR-1 + lactate + Cr(VI).

630 BIOTECHNOLOGY AND BIOENGINEERING, VOL. 83, NO. 6, SEPTEMBER 20, 2003 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; (᭹), Cr outside ࡗ ᭝ ᮀ the cells; ( ), Cr(VI) standard; ( ), Cr(III) oxide standard; ( ), Cr(III)Cl3 standard. tions. Bacteria often preferentially utilize more energeti- precipitates both within and surrounding cells, including cally favorable electron acceptors. For example, in the pres- intracellular reduced chromium globules, that may contrib- ence of nitrate, reduction of Mn(IV), thiosulfate, and Fe(III) ute to the inhibitory effects of chromium. in Shewanella putrefaciens 200 was inhibited, indicating that nitrate was the preferred electron acceptor for anaerobic We thank James Bouwer from the National Center for Micros- respiration (DiChristina, 1992). Although oxygen and ni- copy and Imaging Research for help with the Gatan Imaging trate have half-reaction reduction potentials that are higher Filter. We also thank Jizhong Zhou for supplying the MR-1 strain used in this study, and Alfred Spormann and Weimin Wu for or comparable (respectively) to Cr(VI), MR-1 is able to helpful suggestions on the manuscript. 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 References inhibited by oxygen or nitrate. However, it is clear that kinetic parameters vary significantly between oxic and an- Alvarez-Cohen L, McCarty PL. 1991. A cometabolic biotransformation model for halogenated aliphatic compounds exhibiting product toxic- oxic conditions. ity. Environ Sci Technol 25:1381–1387. Our results also indicate that Cr(VI) inhibits aerobic Bopp LH, Ehrlich HL. 1988. Chromate resistance and reduction in Pseu- growth and denitrification above certain levels. This may domonas fluorescens strain LB300. Arch Microbiol 150:426–431. influence the sequence in which electron acceptors are re- Daulton TL, Little BJ, Lowe K, Jones-Meeham J. 2002. Electron energy duced in a mixed waste setting. For example, the presence loss spectroscopy techniques for the study of microbial chromium(VI) reduction. J Microbiol Methods 50:39–54. of toxic levels of Cr(VI) may hinder denitrification and DiChristina TJ. 1992. The effects of nitrate and nitrite on dissimilatory iron prevent the redox potential of the system from decreasing to reduction in Shewanella putrefaciens 200. J Bacteriol 174:1891–1896. levels where other electron acceptors are reduced. Toxicity Eary L, Rai D. 1987. Kinetics of chromium(III) oxidation to chromium(VI) also appears to limit Cr(VI) transformation capacity; se- by reaction with manganese dioxide. Environ Sci Technol 21: quential spikes of lower levels of Cr(VI) allowed for a 1187–1193. larger transformation capacity over a longer duration than Egerton RF. 1996. Electron energy-loss spectroscopy in the electron microscope. New York: Plenum Press. single spikes of a higher Cr(VI) concentration. Although the Fein JB, Fowle DA, Cahill J, et al. 2002. Nonmetabolic reduction of Cr(VI) mechanism of Cr(VI) toxicity in MR-1 is not well under- by bacterial surfaces under nutrient-absent conditions. Geomicrobiol J stood, TEM and EELS studies revealed reduced chromium 19:369–382.

636 BIOTECHNOLOGY AND BIOENGINEERING, VOL. 83, NO. 6, SEPTEMBER 20, 2003 Rev Environ Contam Toxicol 178:93–164 © Springer-Verlag 2003

Chromium–Microorganism Interactions in Soils: Remediation Implications Sara P.B. Kamaludeen, Mallavarapu Megharaj, Albert L. Juhasz, Nabrattil Sethunathan, and Ravi Naidu

Contents I. Introduction ...... 94 A. Forms of Chromium ...... 95 B. Sources of Chromium in Soil ...... 95 C. Chromium Transformations in Soil ...... 97 II. Physicochemical Factors Governing Chromium Transformations in Soil ...... 97 A. Soil Physical Factors ...... 97 B. Soil pH ...... 97 C. Organic Matter ...... 97 D. Iron ...... 98 E. Manganese ...... 99 III. Microbiological Factors Governing Chromium Transformations in Soil ...... 103 A. Resistance or Tolerance to Cr(VI) ...... 103 B. Direct Cr(VI) Reduction ...... 104 C. Indirect Reduction ...... 119 D. Biotic–Abiotic Coupling in Mn Oxide-Mediated Oxidation of Chromium(III) ...... 121 IV. Implications of Chromium Transformations on Microorganisms and their Activities ...... 126 A. Microorganisms ...... 126 B. Effects on Soil Microbial Community ...... 130 C. Effect on Soil Microbial Processes and Activities ...... 132 V. Remediation of Chromium-Contaminated Water and Soils ...... 141 A. Remediation Technologies for Wastewater and Solutions ...... 141 B. Remediation Technologies for Chromium Wastes in Soils ...... 143 C. Bioremediation ...... 144 D. Applicability of Phytostabilization to Cr-Contaminated Soil ...... 147 VI. Challenges ...... 148 Summary ...... 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.

93 114 S.P.B. Kamaludeen et al. Reference 1995 Garnham and Green — Liu et al. 1995 Mechanism 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 Nonmetabolizing bacte- Fein et al. 2002 efficiency Cr-reducing of Cr(V) during Cr(VI) reduction Cr(VI) Cr(VI) 18 d to Cr(III);of 50% Cr(III) formed accumulated in the cells and 50% in the medium; Cr(VI) reduction associated with hetero- cysts — Transitory formation —— — Enzymatically reduced — Losi et al. 1994b — Enzymatically reduced — Losi et al. 1994b — Chromate reduced in level Cr tolerance — — — — Source — ). sp. sp. and Mougeotia Spirogyra Oscillatoria Chlorella Anabaena variabilis Identification Sporosarcina ureae Shewanella putrefaciens Algae Table 1. ( Continued Bacillus subtilis Chromium–Microorganism Interactions 117 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 (Viama- jala 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 Agrobacter- ium 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 pro- nounced 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. Evi- dence suggested a negative correlation between chromate reduction by the rest- Chromium–Microorganism Interactions 153

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