BIOHYDROMETALLURGY "A SUSTAINABLE TECHNOLOGY IN EVOLUTION"
Proceedings of the International Biohydrometallurgy Symposium, IBS 2003, held in Athens, Hellas, September 14-19, 2003
Part I Bioleaching Applications, Bioremediation Environmental Applications
Edited by Marios Tsezos Artin Hatzikioseyian Emmanouela Remoudaki
Associate Editors Pavlina Kousi Roza Vidali
NATIONAL TECHNICAL UNIVERSITY OF ATHENS School of Mining and Metallurgical Enginnering Laboratory of Environmental Science and Technology Heroon Polytechniou 9, 157 80 Zografou, Greece Tel: (+30) 2107722172, (+30) 2107722271, Fax: (+30) 2107722173 Contact: Professor Marios Tsezos, e-mail:
[email protected]
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First edition 2004
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Preface The present edition includes the proceedings of the 15th International Biohydrometallurgy Symposium (IBS 2003) held in Athens, Greece on September 14th19th, 2003. Continuing the effort on the understanding of the interactions between metals and microbial cells and on developing and applying biohydrometallurgical processes, International Biohydrometallurgy Symposia offer the opportunity of exchanging international experience on a wide range of topics from metal extraction to environmental remediation. During IBS 2003, the issues of sustainability and environmental remediation, worldwide priorities, were addressed from different points of view. Biohydrometallurgy is a sustainable innovative technology, which in many cases, during the last decade, has successfully replaced classical metal extraction processes, from minerals and rocks. Combining a competitive technology with minimum environmental impact is the challenge for optimization of technologies applied today and/or to be applied in the future. Recent advances towards the quantitative description of the interactions between metals and microbial cells as well as the identification of key parameters controlling these interactions, play an important role in metal extraction processes optimization and in the development of treatment technologies for liquid and solid metallurgical discharges. The 15th International Biohydrometallurgy Symposium opened by the invited plenary lecture: "Biohydrometallurgy: a sustainable technology in evolution" given by Professor Giovanni Rossi from the University of Cagliari, Italy. Professor Rossi honored the Symposium with his presence, reviewed the state of the art and pointed out to the future trends in different areas of biohydrometallurgy. The Symposium was organized along five sessions: Bioleaching Applications and Technology Developments. Bioremediation – Environmental Applications. Biosorption Fundamentals and Technology Developments. Microbiology Fundamentals. Molecular Biology and Taxonomy. All papers included in the present edition were previously reviewed by a minimum of two experts selected among the International Scientific Committee Members as well as among prestigious researchers in the biohydrometallurgy science and technology fields. Among the 160 papers included in this edition, the Organising Committee aimed at providing the opportunity to the Symposium participants to attend as many oral presentations as possible, according to originality and scientific merit. Sixty-five oral i
Preface
presentations were made during IBS 2003. The rest of the communications were presented in the poster session. A Closing Session, chaired by a panel of experts and pioneers in the corresponding areas, was organized to conclude the main topics of the Conference and to point out future trends in scientific areas of Biohydrometallurgy. From this position, we wish to acknowledge all the members of the International Scientific Committee: Antonio Ballester, Barrie Johnson, Bohumil Volesky, Borje Lindstrom, Carlos Jerez, Corale Brierley, David Holmes, Dominique Morin, Douglas Rawlings, Edgardo Donati, Eric Guibal, Giovanni Rossi, Gregory Karavaiko, Henry Erhlich, James Brierley, John Duncan, K. A. Natarajan, Kishore Paknikar, Klaus Bosecker, Marios Tsezos, Olli Tuovinen, Paul Norris, Piet Bos, Ralph Hackl, Ricardo Amils, Stoyan Groudev, Tomas Vargas, Tsuyoshi Sugio, Virginia Ciminelli, Wolfgang Sand, for participating in the reviewing and selection of the manuscripts submitted to IBS 2003. We also wish to express our appreciation to prestigious researchers non members of the International Scientific Committee for assisting the reviewing procedure: Anthimos Xenidis, Frantz Glombitza, Georgios Anastassakis, Konstantinos Komnitsas, Ludo Diels, Lynne Macaskie, Nymphodora Papassiopi, Styliani Agatzini-Leonardou. We also thank our colleagues at IBS 2003 from the National Organizing Committee: Anthimos Xenidis, Emmanouil Zevgolis, Georgios Anastassakis, Konstatina Tsaimou, Konstantinos Komnitsas, Nymphodora Papassiopi, Simos Simopoulos, Styliani AgatziniLeonardou, for their valuable assistance and support. Acknowledgements are also due to the many others who participated in the organization of the Symposium, the authors and the many participants who represented many countries around the world. Special thanks also to Mrs Pavlina Kousi and Mrs Roza Vidali, Ph.D candidate students of our Laboratory, for editing the final manuscripts for the preparation of the hardcopies of the IBS 2003 proceedings. Finally, we wish to thank the National Technical University of Athens (NTUA), The Ministry of Development: General Secretariat of Science and Technology, The Hellenic Ministry of Culture, The Hellenic Technical Chamber, The National Institute of Geology and Mineral Exploration for supporting the Symposium. Professor Marios Tsezos Dr. Emmanouela Remoudaki Dr. Artin Hatzikioseyian National Technical University of Athens, School of Mining and Metallurgical Engineering, Laboratory of Environmental Science and Engineering
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Table of contents Preface .................................................................................................................................. i Table of contents ................................................................................................................ iii
PART I PLENARY LECTURE Biohydrometallurgy: a sustainable technology in evolution Giovanni Rossi ...................................................................................................................... 3
CHAPTER 1 BIOLEACHING APPLICATIONS A novel bio-leaching process to recover valuable metals from Indian Ocean nodules using a marine isolate Mukherjee A., Raichur A.M., Modak J.M., Natarajan K.A. ............................................... 25 A novel biotechnological process for germanium recovery from brown coal Xianwan Y., Yun Z., Yuxia G., Banghui G. ......................................................................... 35 Aerobic and anaerobic bacterial leaching of manganese Zafiratos J.G., Agatzini-Leonardou S. ................................................................................ 41 Anaerobic iron sulfides oxidation Schippers A. ........................................................................................................................ 55 Bacterial growth and propagation in chalcocite heap bioleach scenarios Petersen J., Dixon D.G. ...................................................................................................... 65 Bacterial leaching studies of a Portuguese flotation tailing Costa M.C., Carvalho N., Iglesias N., Palencia I. ............................................................. 75
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Bacterial tank leaching of zinc from flotation tailings Panin V.V., Adamov E.V., Krylova L.N., Pivovarova T.A., Voronin D.Yu., Karavaiko G.I. .................................................................................................................... 85 Behaviour of elemental sulfur in the biohydrometallurgical processing of refractory gold-sulfide concentrates of various mineral types Sedelnikova G.V., Savari E.E. ............................................................................................ 91 Beneficiation of phosphatic ores from Hirapur, India Agate A.D. ........................................................................................................................ 101 Biohydrometallurgy of antimony gold-bearing ores and concentrates Solozhenkin P.M., Nebera V.P. ........................................................................................ 107 Bioleaching of Argentinean sulfide ores using pure and mixed cultures Frizan V., Giaveno A., Chiacchiarini P., Donati E. ......................................................... 117 Bioleaching of complex gold-lead ores Ulberg Z., Podolska V., Yermolenko A., Yakubenko L., Pertsov N. ................................. 127 Bioleaching of electronic scrap material by Aspergillus niger Ten W.K., Ting Y.P. .......................................................................................................... 137 Bioleaching of metallic sulphide concentrate in continuous stirred reactors at industrial scale – Experience and lessons Morin D., d’Hugues P., Mugabi M. ................................................................................. 147 Bioleaching of natural zeolite – the processes of iron removal and chamfer of clinoptilolite grains Styriakova I., Kolousek D., Styriak I., Lengauer C., Tillmanns E. ................................... 157 Bioleaching of pyrite by defined mixed populations of moderately thermophilic acidophiles in pH-controlled bioreactors Okibe N., Johnson D.B. .................................................................................................... 165 Biolixiviation of Cu, Ni, Pb and Zn using organic acids produced by Aspergillus niger and Penicillium simplicissinum Galvez-Cloutier R., Mulligan C., Ouattara A. ................................................................. 175 Biooxidation of pyrite by Acidithiobacillus ferrooxidans in single- and multi-stage continuous reactors Canales C., Gentina J.C., Acevedo F. .............................................................................. 185 Chemical chalcopyrite leaching and biological ferric solvent production at pH below 1 Kinnunen P.H.-M., Salo V.L.A., Pehkonen S.O., Puhakka J.A. ....................................... 193
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Comparative study of the bioleaching of two concentrates of chalcopyrite using mesophilic microorganisms in the presence of Ag(I) Lopez-Juarez A., Rivera-Santillan R.E. ............................................................................ 203 Comparison of air-lift and stirred tank batch and semi continuous bioleaching of polymetallic bulk concentrate Tipre D.R., Vora S.B., Dave S.R. ...................................................................................... 211 Effect of pH and temperature on the biooxidation of a refractory gold concentrate by Sulfolobus metallicus Nancucheo I., Gentina J.C., Acevedo F. .......................................................................... 219 Effect of the pulp density and particle size on the biooxidation rate of a pyritic gold concentrate by Sulfolobus metallicus Valencia P., Gentina J.C., Acevedo F. ............................................................................. 227 Enhancement of chalcopyrite bioleaching capacity of an extremely thermophilic culture by addition of ferrous sulphate Rubio A., Garcia Frutos F.J. ............................................................................................ 235 Evaluation of microbial leaching of uranium from Sierra Pintada ore. Preliminary studies in laboratory scale Paulo P.S., Pivato D., Vigliocco A., Lopez J., Castillo A. ................................................ 243 Extraction of copper from mining residues and sediments by addition of rhamnolipids Mulligan C.N., Dahrazma B. ............................................................................................ 253 Improving of film coating bioleaching using biorotor process Shahverdi A.R., Oliazadeh M., Rohi R., Davodi M. ......................................................... 261 Isolation and evaluation of indigenous iron- and sulphur-oxidising bacteria for heavy metal removal from sewage sludge Matlakowska R., Sklodowska A. ....................................................................................... 265 Kinetics of ferrous iron oxidation with Sulfolobus metallicus at 70°C Meruane G., Carcamo C., Vargas T. ............................................................................... 277 Kinetics of sulphur oxidation: pH and temperature influence on bioleaching Patino E., Sandoval R., Frenay J. .................................................................................... 285 Leaching of iron from China clay with oxalic acid: effect of acid concentration, pH, temperature, solids concentration and shaking Mandal S.K., Banerjee P.C. ............................................................................................. 291 Mathematical modeling of the chemical and bacterial leaching of copper ores in stack Zeballos F., Filho O.B., de Carvalho R.J. ........................................................................ 301 v
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Model for bacterial leaching of copper sulphides by forced aeration Sidborn M., Moreno L. ..................................................................................................... 311 Optimal oxygen and carbon dioxide concentrations for thermophilic bioleaching archaea de Kock S.H., Naldrett K., du Plessis C.A. ....................................................................... 319 Optimization study on bioleaching of municipal solid waste (MSW) incineration fly ash by Aspergillus niger Xu T.J., Ting Y.P. .............................................................................................................. 329 Production of an Acidithiobacillus ferrooxidans biomass using electrochemical regeneration of energetic substrate Morra C., Gondrexon N., Magnin J.-P., Deseure J., Ozil P. ........................................... 337 Removal of dibenzothiophene from fossil fuels with the action of iron(III)-ion generated by Thiobacillus ferrooxidans: Analytical aspects Beskoski V.P., Matic V., Spasic S., Vrvic M.M. ................................................................ 345 Solids loading in the bioleach slurry reactor: mechanisms through which particulate parameters influence slurry bioreactor performance Harrison S.T.L., Sissing A., Raja S., Pearce S.J.A., Lamaignere V., Nemati M. ............. 359 The development of a hybrid biological leaching-pressure oxidation process for auriferous arsenopyrite/pyrite feedstocks Dymov I., Ferron C.J., Phillips W. ................................................................................... 377 The development of the first commercial GEOCOAT® heap leach for refractory gold at the Agnes mine, Barberton South Africa Harvey T.J., Bath M. ........................................................................................................ 387 The electrochemistry of chalcopyrite bioleaching using bacteria modified powder micro-electrode Hongxu L., Dianzuo W., Yuehua H., Renman R. .............................................................. 399 The influence of crystal orientation on the bacterial dissolution of pyrite Ndlovu S., Monhemius A.J. ............................................................................................... 409 The influence of temperature and pH on the bioleaching of copper from a flotation concentrate of chalcopyrite Medrano-Roldan H., Salazar M.F.M., Pereyra-Alférez B., Solis-Soto A., Ramirez-Rodriguez D.G., Alvarez-Rosales E., Galan-Wong L.J. .................................... 419 The role of chemolitotrophic bacteria in the oxide copper ore heap leaching operation at Sarcheshmeh Copper Mine Seyed Baghery S.A., Shahverdi A.R., Oliazadeh M. ......................................................... 423
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Three-stage revolving drum biohydrometallurgical reactor for continuous operation Loi G., Trois P., Rossi G. ................................................................................................. 429 Use of biosurfactants for the mineral surfaces modification Sadowski Z., Maliszewska I., Polowczyk I. ....................................................................... 439
CHAPTER 2 BIOREMEDIATION ENVIRONMENTAL APPLICATIONS A new bench scale restoration method for a mercury-polluted soil with a mercury resistant Acidithiobacillus ferrooxidans strain SUG 2-2 Negishi A., Maeda T., Takeuchi F., Kamimura K., Sugio T. ............................................ 449 A novel type of microbial metal mobilization: cyanogenic bacteria and fungi solubilize metals as cyanide complexes Brandl H., Stagars M., Faramarzi M.A. ........................................................................... 457 An approach to cyanide degradation in wastewater of gold ore processing Podolska V., Ulberg Z., Pertsov N., Yakubenko L., Imanakunov B. ................................ 465 Available options for the bioremediation and restoration of abandoned pyritic dredge spoils causing the death of fringing mangroves in the Niger Delta Ohimain E. I. .................................................................................................................... 475 Bacterial reduction of TcO4- under the haloalkaline conditions Khijniak T., Medvedeva-Lyalikova N.N., Simonoff M. ..................................................... 483 Biodegradation of cyanides under saline conditions by a mixotrophic Pseudomonas putida Bipinraj N.K., Joshi N.R., Paknikar K.M. ........................................................................ 491 Bioleach of a fluvial tailings deposit material indicates long term potential for pollution Willscher S., Clark T.R., Cohen R.H., Ranville J.F., Smith K.S., Walton-Day K. ............ 497 Bioleaching of copper converter slag using A. ferrooxidans Seyed Baghery S.A., Oliazadeh M. ................................................................................... 507 Biooxidation of mine tailings using a mixed bacterial population Zahari M.A.K.M., Jaapar J., Bunyok M.A., Sohor S.H., Ahmad W.A. ............................. 513 Chromate reduction by immobilized cells of Desulfovibrio vulgaris using biologically produced hydrogen Humphries A.C., Penfold D.W., Forster C.F., Macaskie L.E. ......................................... 525 vii
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Clean-up of mine waters from a uranium deposit by means of a constructed wetland Groudev S.N., Komnitsas K., Spasova I.I., Paspaliaris I. ................................................ 533 Degradation of tetracyanonickelate (II) by Cryptococcus humicolus in biofilm reactors Kwon H.K., Woo S.H., Sung J.Y., Park J.M. .................................................................... 541 Development of a bio-process using sulfate-reducing bacteria to remove metals from surface treatment effluents Battaglia-Brunet F., Foucher S., Denamur A., Chevard S., Morin D., Ignatiadis I. ....... 549 Effects of total-solids concentration on metal bioleaching from sewage sludge Villar L.D., Garcia O. Jr .................................................................................................. 559 Enhancement of electrodialytic soil remediation through biosorption Jensen P.E., Ottosen L.M., Ahring B.K. ........................................................................... 567 Fundamentals of the uranium separation in constructed wetlands Glombitza F., Karnatz F., Fischer H., Pinka J., Janneck E. ............................................ 575 Geomicrobiological risk assessment of abandoned mining sites Bosecker K., Mengel-Jung G., Schippers A. ..................................................................... 585 Immobilisation and growth of Acidithiobacillus ferrooxidans on refractory clay tiles Donati E., Martinez L., Curutchet G. ............................................................................... 595 Investigation of bioremediation techniques for cleaning-up arsenic contaminated soils Vaxevanidou K., Papassiopi N., Paspaliaris I. ................................................................ 603 Leaching characteristics of heavy metals from sewage sludge by Acidithiobacillus thiooxidans MET Cho K.S., Moon H.S., Yoo N.Y., Ryu H.W. ....................................................................... 613 Mercury removal by polymer-enhanced ultrafiltration using chitosan as the macroligand Kuncoro E.K., Lehtonen T., Roussy J., Guibal E. ............................................................ 621 Microbial recovery of copper from printed circuit boards of waste computer by Acidithiobacillus ferrooxidans Cho K.S., Choi M.S., Hong J.H., Kim D.S., Ryu H.W., Kim D.J., Sohn J.S., Park K.H. .. 631 Oxidation of iron, sulfur and arsenic in mine waters and mine wastes: an important role for novel Thiomonas spp Coupland K., Battaglia-Brunet F., Hallberg K.B., Dictor M.C., Garrido F., Johnson D.B. ................................................................................................................................... 639 viii
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Oxidation of metallic copper by Acidothiobacillus Ferrooxidans Lilova K., Karamanev D. .................................................................................................. 647 Process monitoring of biodesulfurization of high sulfur coal in packed columns using molecular ecology methods Gómez F., Cara J., Carballo M.T., Moran A., Amils R., García Frutos F.J. .................. 653 Regeneration of hydrogen sulfide using sulfate reducing bacteria for photo catalytic hydrogen generation Takahashi Y., Suto K., Inoue C., Chida T. ........................................................................ 663 Remediation of sites contaminated by heavy metals: sustainable approach for unsaturated and saturated zones Diels L., Geets J., Vos J., Van Broekhoven K., Bastiaens L. ............................................ 671 Removal of chromium(VI) through a two-step process using sulphur-oxidising and sulphate-reducing bacteria Donati E., Viera M., Curutchet G. ................................................................................... 681 Removal of Mn(II) ions by manganese-oxidizing fungus at neutral pHs in the presence of carbon fiber Sasaki K., Endo M., Takano K., Konno H. ....................................................................... 689 Simultaneous removal of oil and heavy metals from waste waters by means of a permeable reactive barrier Groudeva V.I., Groudev S.N., Doycheva A.S. .................................................................. 697 The exploitation of sulphate-reducing bacteria for the reclamation of calcium sulphate sludges Luptakova A., Kusnierova M., Bezovska M., Fecko P. ..................................................... 703 The role of metal–organic complexes in the treatment of chromium containing effluents in biological reactors Remoudaki E., Hatzikioseyian A., Kaltsa F., Tsezos M. ................................................... 711 The selective precipitation of heavy metals by sulphate-reducing bacteria Luptakova A., Kusnierova M., Bezovska M., Fecko P. ..................................................... 719
APPENDIX Author index ................................................................................................................... A-3 Subject index ................................................................................................................. A-11
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PART II CHAPTER 3 BIOSORPTION A methodological approach to investigate the pH effect on biosorption process: experimental and modeling procedures Veglio F., Beolchini F., Pagnanelli F., Toro L. ............................................................... 731 A model for the copper biosorption in dried leaves de Carvalho R.P., De Sousa A-M.G., Freitas J.R., Rubinger C.P.L., Krambrock K. ...... 741 Agar-plate screening of effective metal biosorbents among year Podgorsky V.S., Lozovaya O.G., Kasatkina T.P., Fomina M.A. ...................................... 749 Bioremediation of chromium using Bacillus polymyxa Thyagarajan H., Subramanian S., Natarajan K.A. .......................................................... 759 Biosorption and bioaccumulation of heavy metals by bacteria isolated from contaminated sites of Karachi, Pakistan Nuzhat A., Uzma B., Fouad M. Qureshi, Fehmida F. ...................................................... 771 Biosorption equilibria with Spirogyra insignis Romera E., Fraguela P., Ballester A., Blazquez M.L., Munoz J.A., Gonzalez F. ............ 783 Biosorption of 226Ra and Ba by Sargassum sp. da Costa W.C., Garcia O. Jr., de Azevedo Gomes H. ...................................................... 793 Biosorption of arsenic and heavy metals on a ceramic-based biomass. Batch equilibrium experiments with Cu2+ model solutions Horak G., Willscher S., Werner P., Pompe W. ................................................................. 799 Biosorption of chromium (VI) by marine algal biomass Tan L.H., Chen J.P., Ting Y.P. ......................................................................................... 807 Biosorption of heavy metal ions from aqueous solutions by local seaweeds Sheng P.X., Chen J.P., Ting Y.P. ...................................................................................... 817 Biosorption of heavy metals onto an olive pomace: adsorbent characterisation and equilibrium modelling Pagnanelli F., Ubaldini S., Veglio F., Toro L. ................................................................. 825 Biosorption of Hg by vegetal biomasses Pimentel P.F., de Carvalho R.P., Santos M.H., Andrade M.C. ....................................... 835
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Biosorption of lead in aquatic environment by Mucor rouxii biomass Som Majumdar S., Saha T., Bandhapadhyay T., Chatterjee S., Guha A.K. ..................... 843 Cadmium(II) biosorption by Aeromonas caviae: kinetic modeling Loukidou M.X., Karapantsios T.D., Zouboulis A.I., Matis K.A. ...................................... 849 Chromium uptake by pretreated cells of Aeromonas hydrophila isolated from textile effluents Zakaria Z.A., Ahmad W.A. ................................................................................................ 859 Copper ion adsorbed on chitosan beads: Physico-chemical characterization Chatterjee S., Som Majumdar S., Chatterjee B.P., Guha A.K. ......................................... 869 Development of a process for biosorptive removal of mercury from aqueous solutions Tupe S., Paknikar K. ......................................................................................................... 877 Effects of ionic strength, background electrolytes and heavy metals on the biosorption of hexavalent chromium by Ecklonia biomass Park D., Park J.M., Yun Y.-S. ........................................................................................... 883 Evaluation of silver recovery from photographic waste by Thiobacillus ferrooxidans and chitin Thiravetyan P., Nakbanpote W., Songkroah C. ................................................................ 891 Influence of the treatment of fungal biomass on sorption properties for lead and mercury uptake Spanelova M., Svecova L., Guibal E. ............................................................................... 899 Lanthanum and neodymium biosorption by different cellular systems Palmieri M., Garcia O. Jr. ............................................................................................... 911 Modeling of chromium biosorption by seaweed Sargassum sp. biomass in fixedbed column in series Cossich E.S., Silva E.A., Tavares C.R.G., Mesquita H.M., Eidan L.S. ............................ 919 Modelling and optimisation of copper ion uptake by Acidithiobacillus ferrooxidans Boyer A., Baillet F., Magnin J.-P., Ozil P. ....................................................................... 925 Platinum and palladium recovery from dilute acidic solutions using sulfate reducing bacteria and chitosan derivative materials Chassary P., de Vargas Parody I., Ruiz M., Macaskie L., Sastre A., Guibal E. .............. 935 Preliminary study of lead sorption by selected sorbents Ly Arrascue M., Bauer-Cuya J., Peirano Blondet F., Roussy J., Guibal E. .................... 947
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Regeneration of biomass after sorption of heavy metals Massacci P., Migliavacca E., Ferrini M. ......................................................................... 957 Structural modeling of arsenic biosorption using X-Ray spectroscopy (XAS) Teixeira M.C., Duarte G., Ciminelli V.S.T. ...................................................................... 965 Uranium and thorium removal by a Pseudomonas biomass: sorption equilibrium and mechanism of metal binding Sar P., Kazy S.K., D’Souza S. F. ...................................................................................... 975
CHAPTER 4 MICROBIOLOGY FUNDAMENTALS A model for iron uptake in Acidithiobacillus ferrooxidans based upon genome analysis Quatrini R., Veloso F., Jedlicki E., Holmes D.S. .............................................................. 989 Activity and occurrence of leaching bacteria in mine waste at Cartagena, Spain, in the years 1991 until 2000 Sand W., El Korchi-Buchert D., Rohwerder T. ................................................................ 997 An AFM-study on the adhesion of Acidithiobacillus ferrooxidans and Leptospirillum ferrooxidans to surfaces of pyrite Kinzler K., Sand W., Telegdi J., Kalman E. ................................................................... 1003 An X-ray photoelectron spectroscopy study of the mechanism of microbially assisted dissolution of chalcopyrite Parker A., Klauber C., Stott M., Watling H.R., Van Bronswijk W. ................................ 1011 Analysis of chalcopyrite (CuFeS2) electrodes utilizing galvanic current in the presence of Acidithiobacillus ferrooxidans Bevilaqua D., Benedetti A.V., Fugivara C.S., Garcia O. Jr. .......................................... 1023 Application of the bacterial weathering of silicate minerals in the improvement of quality of non-metallics Styriakova I., Styriak I. ................................................................................................... 1029 Assessment of acid production potential of sulphide minerals using Acidithiobacillus ferrooxidans and microbial sulphate reduction using Desulfotomaculum nigrificans Chockalingam E., Subramanian S., Natarajan K.A., Braun J.J. .................................... 1037 Comparative study on pit formation and interfacial chemistry induced by Leptospirillum and Acidothiobacillus ferrooxidans during FeS2 leaching Tributsch H., Rojas-Chapana J. ..................................................................................... 1047 xii
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Composition of biofilm communities in acidic mine waters as revealed by combined cultivation and biomolecular approaches Kimura S., Coupland K., Hallberg K.B., Johnson D.B. ................................................. 1057 Computational fluid dynamics simulation of immobilized Acidothiobacillus ferrooxidans Metodiev B., Lilova L., Karamanev D. ........................................................................... 1067 Contribution to the quantification of the Acidithiobacillus ferrooxidans biomass concentration from the oxygen uptake rate Savic D.S., Veljkovic V.B., Lazic M.L. ............................................................................ 1077 Electrochemical and microbiological characterization of mercury in contact with mud Cruz F., Welzel A., Sampaio C., Englert G.E., Müller I.L. ............................................ 1085 Evaluating the growth of free and attached cells during the bioleaching of chalcopyrite with Sulfolobus metallicus Escobar B., Hevia M.J., Vargas T. ................................................................................. 1091 Experimental and modeling studies on inhibition effect of solution conditions on activity of Acidithiobacillus ferrooxidans during biooxidation of mixed sulphidic concentrates Chandraprabha M.N., Modak J.M., Natarajan K.A. ..................................................... 1099 Ferrous ion oxidation by an activated carbon cloth modified with Acidithiobacillus ferrooxidans de J. Cerino-Cordova F., Magnin J.P., Gondrexon N., Ozil P. ..................................... 1109 Heavy metal precipitation by off-gases from aerobic culture of Klebsiella pneumoniae M426 Essa A.M.M., Macaskie L.E., Brown N.L. ...................................................................... 1119 Influence of pH, Mg2+ and Mn2+ on the bioleaching of nickel laterite ore using the fungus Aspergillus niger O5 Coto O., Gutierrez D., Abin L., Marrero J., Bosecker K. .............................................. 1127 Mercury tolerance of thermophilic Bacillus sp. and Ureibacillus sp. Glendinning K.J., Brown N.L. ........................................................................................ 1137 Reduction of Pd(II) with Desulfovibrio fructosovorans, its [Fe]-only hydrogenase negative mutant and the activity of the obtained hybrid bioinorganic catalysts Mikheenko I.P., Baxter-Plant V.S., Rousset M., Dementin S., Adryanczyk-Perrier G., Macaskie L.E. ................................................................................................................. 1147 Removal of cobalt, strontium and caesium from aqueous solutions using native biofilm of Serratia sp. and biofilm pre-coated with hydrogen uranyl phosphate Paterson-Beedle M., Macaskie L.E. ............................................................................... 1155 xiii
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Removal of soluble manganese from mine waters using a fixed bed column bioreactor Johnson D.B., Miller H., Ukermann S., Hallberg K.B. .................................................. 1163 Sulfane sulfur of persulfides is the actual substrate of the sulfur-oxidizing enzymes from Acidithiobacillus and Acidiphilium spp. Rohwerder T., Sand W. ................................................................................................... 1171 Sulfate reduction at low pH by mixed cultures of acidophilic bacteria Sen A.M., Kimura S., Hallberg K.B., Johnson D.B. ....................................................... 1179 Sulfur assimilation in Acidithiobacillus ferrooxidans Valdes J., Jedlicki E., Holmes D.S. ................................................................................ 1187 Survival of acidophilic bacteria under conditions of substrate depletion that occur during culture storage Johnson D.B., Bruhn D. F., Roberto F.F. ...................................................................... 1195 Synthesis of nanophase hydroxyapatite by Serratia sp. N14 Yong P., Sammons R.L., Marquis P.M., Lugg H., Macaskie L.E. .................................. 1205 The effect of maintenance on the ferrous-iron oxidation kinetics of Leptospirillum ferrooxidans Dempers C.J.N., Breed A.W., Hansford G.S. ................................................................. 1215 The kinetics of thermophilic ferrous-iron oxidation Searby G.E., Hansford G.S. ............................................................................................ 1227 The role of microorganisms in dispersion of thallium compounds in the environment Sklodowska A., Golan M., Matlakowska R. .................................................................... 1237
CHAPTER 5 MOLECULAR BIOLOGY AND TAXONOMY A promiscuous, broad-host range, IncQ-like plasmid isolated from an industrial strain of Acidithiobacillus caldus, its accessory DNA and potential to participate in the horizontal gene pool of biomining and other bacteria Goldschmidt G.K., Gardner M.N., van Zyl L.J., Deane S.M., Rawlings D.E. ............... 1249 Analysis of salt-induced outer membrane proteins in Acidithiobacillus ferrooxidans NASF-1 Kamimura K., Yamakado M., Shishikado T., Sugio T. ................................................... 1261
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Bioinformatic analysis of biofilm formation in Acidithiobacillus ferrooxidans Barreto M., Rivas M., Holmes D.S., Jedlicki E. ............................................................. 1271 Diversity of Gram-negative bacteria at Malanjkhand copper mine, India Dave S.R., Tipre D.R. ..................................................................................................... 1279 Expression proteomics of Acidithiobacillus ferrooxidans grown in different metal sulfides: analysis of rhodanese-like proteins Ramirez P., Valenzuela L., Acosta M., Guiliani N., Jerez C.A. ..................................... 1287 Integration of metal-resistant determinants from the plasmid of an Acidocella strain into the chromosome of Escherichia coli DH5α Ghosh S., Mahapatra N.R., Nandi S., Banerjee P.C. ..................................................... 1297 Involvement of Fe2+-dependent mercury volatilization enzyme system in mercury resistance of Acidithiobacillus ferrooxidans strain MON-1 Sugio T., Fujii M., Takeuchi F., Negishi A., Maeda T., Kamimura K. ........................... 1305 Microbial diversity of various metal-sulphides bioleaching cultures grown under different operating conditions using 16S-rDNA analysis d’Hugues P., Battaglia-Brunet F., Clarens M., Morin D. .............................................. 1313 Molecular ecology of the Tinto River, an extreme acidic environment from the Iberian Prytic Belt González-Toril E., Llobet-Brossa E., Casamayor E.O., Amann R., Amils R. ................ 1325 Phenotypic characterization and copper induced stress resistance in the extremely acididophilic Archaeon Ferroplasma acidarmanus Baker-Austin C., Dopson M., Bowen A., Bond P. .......................................................... 1337 Pyrite oxidation by halotolerant acidophilic bacteria Norris P.R., Simmons S. ................................................................................................. 1347 Reversible loss of arsenopyrite oxidizing capabilities by Acidithiobacillus ferrooxidans is associated with swarming phenotype and presence of ISAfel Hurtado J.E. ................................................................................................................... 1353 Searching for physiological functions regulated by the quorum sensing autoinducer AI-1 promoted by afeI/afeR genes in Acidithiobacillus ferrooxidans Farah C., Banderas A., Jerez C.A., Guiliani N. ............................................................. 1361 Systematic analysis of our culture collection for "genospecies" of Acidithiobacillus ferrooxidans, Acidithiobacillus thiooxidans and Leptospirillum ferrooxidans Mitchell D., Harneit K., Meyer G., Sand W., Stackebrandt E. ....................................... 1369
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The strain genotypic heterogeneity of chemolithotrophic microorganisms Kondrateva T.F., Pivovarova T.A., Muntyan L.N., Ageeva S.N., Karavaiko G.I. .......... 1379
APPENDIX Author index ................................................................................................................... A-3 Subject index ................................................................................................................. A-11
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Biohydrometallurgy: a sustainable technology in evolution Giovanni Rossi Dipartimento di Geoingegneria e Tecnologie Ambientali – Università Piazza d’Armi, 19 – 09123 Cagliari, Italia Abstract In the mid to late 1990’s biohydrometallurgy used to be considered an innovative technology, but one can hardly continue to describe it as "innovative" today. Since the advent of biohydrometallurgical processing many major breakthroughs have been achieved and this economically profitable and sustainable technology now finds wide application in a variety of spheres, ranging from metal extraction to environmental remediation. Consequently, now that the pioneering days are over, what is required is a concerted effort by researchers to rationalize and optimize biohydrometallurgical processes. Three main areas of application can be identified: (i) environmental protection; (ii) metal extraction from minerals and rocks: (iii) pre-treatment of minerals to make them amenable to further processing. The fundamentals of biohydrometallurgy draw on a variety of disciplines, ranging from minerals engineering and mineralogy to microbiology, physical chemistry (with strong emphasis on surface science, colloid chemistry and electrochemistry) and solidstate physics. Researchers in all these fields have provided an equally important contribution to the development of this technology and this presentation endeavours to review their achievements and to briefly discuss those issues that remain open. Some of these issues continue to arouse controversy, stimulating the interest of scientists, which can also benefit other fields of science and technology. Based on his experience, the author would like to emphasize the need to establish a Biohydrometallurgical Society, to act as a point of reference to all those, from industry and academia, who are involved in implementing and further developing the technology. The Society should also provide a forum for information flow to decision makers in industry, about the potentials of biohydrometallurgy. Though it relies on the exploitation of the complex synergies between microoorganisms and minerals, this technology, when properly applied, is simple to implement, operationally stable and cost-effective. Finally, the author would like to invite academia to develop suitable curricula to ensure that new generations gain a specific working knowledge of biohydrometallurgy and industry to increase their funding for higher education and for private and academic research. 1.
INTRODUCTION In this review I will describe the evolution of a technology, biohydrometallurgy, which in my opinion offers promising prospects and can be highly rewarding for all those 3
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involved in providing humankind with the mineral resources indispensable for its progress with the minimum harm to the environment. Of course, I will focus most of my considerations on the engineering aspects of biohydrometallurgy and related physical-chemistry, solid state physics and earth sciences implications for two reasons with which I hope you will concur: first, in three recent excellent papers [1-3], all the most important aspects of microbiology applied to Biohydrometallurgy have been exhaustively reviewed in such a masterly way that there is really nothing significant to add at the present time; second, the degree of maturity attained by biohydrometallurgy as a new technology is such that its engineering developments and problems warrant attention. I would also like to point out that I have restricted the references to those necessary to justify some of my statements and that I have omitted many excellent contributions simply because otherwise this talk would have been more of a reference list than a presentation. In effect, on bioleaching kinetics alone I keep more than fifty papers in my files all of which deserve attention and mention in a paper concerning that topic. The Venn diagram shown in Figure 1 provides a visual representation of the interconnections among the various branches of science from which the fundamentals of biohydrometallurgy are derived. The engineering aspects are of paramount importance for commercial applications. In this regard three main groups of processes need to be distinguished: those concerned with metal extraction from rocks, those for mineral upgrading and environmental protection processes. For the sake of clarity these three groups will be examined separately.
Figure 1. Venn-Euler diagram showing biohydrometallurgy "parenthood"
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2.
ENVIRONMENTAL PROTECTION PROCESSES As is well known, the origins of biohydrometallurgy can be traced to the environmental problem created by the pollution of the Ohio River due to acid drainage of coalmines located in the river basin. Environmental protection continues to be one of the main areas of application of biohydrometallurgy and concerns either the development of remediation techniques aimed at low cost inhibition of dangerous effluents from old stopes or dumps or of systems for trapping toxic ions from effluents or for decontaminating polluted soils. Investigations into the properties of cell envelopes as adsorbents, of the so-called microbial derivatives [4,5], the proposed ingenious mathematical adsorption models, analysis of the factors influencing the processes, and the encouraging commercial applications [5] have opened new avenues and some promising research results have also been published recently [6,7] The interactions between microbes and solid surfaces play a very important role and three papers [8-10] on the fundamentals of this subject provide a useful background for those intending to advance in this field. The commercial applications rely on the knowledge of the mechanisms and the extent of adsorption of chemical elements or compounds by microorganisms. Providing the required information to the practitioners is the basic task of microbiology. 2.1 The biological fundamentals Organisms and microorganisms play the role of ion traps and, to some extent, can be considered the biological equivalents of inorganic exchange resins. This branch of biohydrometallurgy involves not only microorganisms but, more generally, all living things especially plants and algae. This distinctive feature already emerged at the time of what can be considered the First Symposium on Biohydrometallurgy, held in Braunschweig: on that occasion, the properties of the alga Hormidium fluitans (Gay) were described [11]. This is a broad field of research, as the ample literature published to date demonstrates. As pointed out by two recent reviews [5,12,13], the technology is promising but does not yet meet, for a number of reasons, the prerequisites for becoming a widespread cost-effective commercial application. For the time being the prospects of its application as a process in its own right seem to be limited. However, because of its characteristic feature of rapid intrinsic kinetics it is envisaged that biohydrometallurgy will be successfully integrated into water purification flowsheets consisting of hybrid technologies. For this type of technology the distinction between intra-biotechnological (IBT) or inter technological (IT) [12] depending on the type of associated processes can be helpful. The term IBT refers to biosorption, bioreduction or bioprecipitation, IT to biotechnology-based processes integrated with non-biotechnology based ones such as chemical precipitation, electrochemical processes, etc. 3.
BIOHYDROMETALLURGY AS A DEVELOPMENT OF EXTRACTIVE METALLURGY From its origins, in the 1940’s, up to the 1980’s, most biohydrometallurgical research work focussed on the ambitious target of developing an environmentally and costeffective process that could compete with pyrometallurgy-based processes for metal extraction from ores and concentrates.
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All of us old-timers certainly recall the enthusiasm with which we carried out our investigations and the encouraging results achieved by the first in-situ and dump leaching operations. Dare I say that almost all mining or metallurgical engineers involved in biohydrometallurgical research tested the process whenever an abandoned copper mine was available. The encouraging results obtained in several mining operations in the late 1970’s, that are still being confirmed today by the latest developments, like the Quebrada Blanca Mine [14] and the relatively low cost of the equipment required, certainly justify the impatience of several of us to start working on concentrate bioleaching. In effect, the results were quite frustrating. I still recall the disappointment after several months of very hard work in a well-equipped laboratory in Northern Italy. There, in 1977, in great secrecy, an Italian mining company entrusted an international team, of which I was part, with the task of developing a chalcopyrite flotation concentrate bioleaching flowsheet. We never succeeded in obtaining a 95% copper leaching in a single STR operated in batch. It was only two decades later that I began to understand the reasons for our failure, which depended simply on our ignorance of certain aspects of solid state physics typical of chalcopyrite. Subsequent research carried out in the light of the contributions of solid-state physicists and of new investigation methods like XPS, demonstrated the great potential of biohydrometallurgy. 3.1 Pre-treatment of run-of-mine ores or flotation concentrates and recovery of valuable metals The processes belonging to this group share the same type of problems and for this reason I will treat them together. The technology of this branch of biohydrometallurgy has already found successful application in cost-effective commercial operations. As summed up in the block diagram of Figure 2, their profitability depends on a series of interrelated factors whose investigation covers a wide range of basic sciences and of technological applications that had yet to be fully explored in the 1970’s. At this juncture I would like to stress the point that I purposely avoided the distinction between "pure" and "applied" sciences, as authoritatively stated more than a century ago by Pasteur [15]: ......No, a thousand times no, there does not exist a category of science to which one can give the name applied science. There are science and applications of science bound together as the fruit of the tree which bears it ..... The most encouraging commercial successes have been achieved in the pre-treatment of gold-bearing complex sulphide ores - notoriously refractory to conventional processing - for the subsequent cyanidation step. The excellent performance of numerous commercial plants is well documented [16,17]. Table 1 provides a summary of these achievements. It is no exaggeration to say that much of the recent research that has contributed to elucidating a great many problems posed by this technology has culminated in these successes. Research efforts directed to coal desulphurisation have produced some interesting results. A semi-commercial pilot plant, the first of its kind, jointly designed, built and operated in partnership between four European research groups in the framework of a project funded by the Commission of European Communities, demonstrated the practical
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Table 1. Operating parameters of some commercial bioleaching operations Plant and location Fairview South Africa Sao Bento Brazil Olympia Greece
Ore minerals
Reactor type
% Solids concentration
Total useful bioreactor volume, m3
Daily throughput per bioreactor unit useful volume, tonn/m3.day
Residence time, hours
Reference
P, A
STR
20
90
0.444
96
[16]
A, P, Pr
STR
20
580
0.138
21
[16,17]
Complex Cu, Fe, As, Zn, Pb sulphides
STR
20
15,936 (3 moduli of 4 1,328 m3 each STR’s)
0.048
96
[16,18]
STR
20
23,376 (4 moduli of 6,974 m3 each STR’s)
0.047
96
[17]
P, A, Stb
STR
20
6x470 = 2,820
0.045
120
[17]
A, P, Pr, Mrc
STR
20
16,200 (3 moduli of 6 900 m3 each STR’s )
0.0444
96
[17,19]
Amantaytau Uzbechistan Wiluna Australia Ashanti Sansu Ghana
P = Pyrite; A = Arsenopyrite; Pr = Pyrrhotite; Mrc = Marcasite; Stb = Stibnite; C = complex Cu, Zn, Pb, As, Fe sulfides
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feasibility of coal biodepyritization. Removal of the so-called "organic sulphur" from coal, though investigated in depth by the same research groups, was not as successful, but did provide the guidelines for future research [20-22]. Solving this problem is of great environmental and economic significance, as billions of tons of fossil coal could be utilized were it possible to remove the organic sulphur therefrom [23].
Figure 2. Block diagram showing the factors affecting reactor bioleaching profitability 3.2 Base metal recovery from minerals As far as I am aware, no commercial bioleaching plant has ever been built for the extraction of base metals from mineral sulphides concentrates. One of the likely reasons for this is that the state of the art technology cannot yet compete with conventional pyrometallurgical processes. The problem here is obviously one of profitability. As Figure 2 shows, two main factors affect process profitability: bioleaching kinetics and reactor characteristics. Both factors are currently being researched, though bioleaching kinetics has received far more attention and will be dealt with first. The development of analytical expressions that mathematically relate the characteristics of: (i) microbial strains, (ii) solid substrate, (iii) process environment and (iv) reactor features, and their interactions with process kinetics has been and continues to be the main objective of bioleaching research. The composition of microbial populations and their synergies have been the almost exclusive hunting ground of microbiologists and can reasonably expected to be so in the future. A better understanding of the role played by the solid substrate requires the involvement of several fields of specialized nonmicrobiological knowledge, ranging from mineralogy associated to solid state physics, chemistry, electrochemistry and mineral engineering. Bioleaching is essentially the product of the the microbial population interacting with the solid substrate but the process 8
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is strongly conditioned by the physico-chemical environment, specifically the liquid phase pH and Eh, and consequently the related electrochemical phenomena, chemical composition, solids concentration, particle size distribution and evolution, and temperature. 3.2.1 Bioleaching microbiology Microorganisms can be considered, latu sensu, as the biocatalysts of mineral oxidation and solubilization processes. Up to now the following research lines have been pursued: (i) identification of the microorganisms involved in the process, (ii) how microorganisms interact among themselves and with the solid substrate, (iii) development of the most suitable microorganisms or microbial associations, (iv) enumeration of bioleaching microorganisms. Identification of the microorganisms involved in the bioleaching process that began with Temple and Hinkle’s discovery in 1948 still continues today. Major breakthroughs include the discovery of Leptospirillum ferrooxidans [24] and of the ability of Sulfolobus to bioleach metal sulphides [25]. These discoveries opened up a whole new world that continues to be investigated today. Incidentally, the discovery of L. ferrooxidans, may well be cited as yet another example of serendipity, because later investigations showed the organism to be quite atypical compared to those species that turned out to be so important in commercial operations [26]. The pioneering work on the physiology of Thiobacilli carried out by Kelly [citations in 27] the application of the theory of chemiosmotic mechanism to A. ferrooxidans proposed by Ingledew, Cox and Helling and by Ingledew [citations in 27] and the work by Don Kelly and Norris [citations in 27] on the moderate thermophiles and on the inhibition by ferrous and ferric ions have provided the tools for interpreting and the approaches for optimizing some laboratory and commercial performances. The results of investigations aimed at gaining a better understanding of the evolution of how bioleaching systems evolve in commercial operations carried out in CFSTR’s (work that appears to be continuing today) have provided a sound understanding of the roles of Thiobacilli and of Leptospirillum ferrooxidans [28-31] in metal sulphide bioleaching and of the importance of properly calibrating the physico-chemical environment on their oxidation activity. Undeniably, this information is invaluable to plant engineers. Identification methodologies involving nucleic acids represent the turning point for unequivocally distinguishing between strains of microorganisms as well as between species and genera and were exhaustively reviewed in a recent paper [32]. The confirmation of the wide diversity among Thiobacillus ferrooxidans strains [33], already observed by a number of researchers and reported two decades ago [34], as well as among the Leptospirillum ferrooxidans strains [35] are significant achievements of this methodology. Another group of methodologies are those based on immunological methods. This diversity is also of great practical significance. I still recall that about forty years ago, I was rather puzzled by the statement of a distinguished colleague as to the ubiquity of T. ferrooxidans. As matter of fact, I myself collected from numerous mines throughout Italy several strains of a microorganism that complied with the characteristics of T. ferrooxidans reported in the literature then available. However, the various strains responded quite differently under identical experimental conditions, a situation that seriously intrigued me, as at the time the differences in performance were inexplicable, and naturally I attributed it to some mistake in the laboratory procedure. The problem 9
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became dramatic when, a few years later, I was testing different strains for developing an industrial bioleaching process. As an engineer I would like to remark, however, that a major drawback of these methodologies is the need for fairly specialized equipment and skilled personnel [32]. I still recall the sophisticated van that housed a fine mobile laboratory where my Dutch colleagues very effectively carried out real time monitoring of the microorganisms used in the Porto Torres coal biodepyritization semi-commercial scale pilot plant [21]. It appears to me that this type of monitoring is very difficult to propose for a normal industrial operation. One potential area that warrants investigation is the development of kind of "diagnostic kit" that requires training accessible to technicians with a bachelor degree education background. The same considerations can be made for enumeration methodologies. In recent years microbial associations have been progressively identified and their importance recognized. Thus, it has been ascertained that several strains of heterotrophs contribute significantly to the activity of autotrophs f.i. acting as scavengers of their metabolites sometimes behaving as inhibitors when in sufficiently high concentrations [36-39]. Looking on the optimistic side, mutualism and synergism [37] should be considered important factors, for process optimisation also in continuous bioreactor systems, which so far have not received the attention they deserve. This is one way of remedying the other bad habit of many biohydrometallurgists – and I regret to say that I am one of them – of using enrichment cultures using substrates such as ferrous iron, sulphur and pyrite for isolating acidophiles. As vigorously pointed out recently [38], this unfortunately widespread practice may result in the selection of a relatively narrow range of acidophiles, depriving the cultures of the contribution of some important bacteria. The interactions between components of the microflora present in leaching operations were brought to light many years ago by Groudev [citations in 27]. In more recent times the existence and importance of this interaction seem to have been confirmed, in relation for instance to the role played by chemoorganotrophic microorganisms, such as Acidiphilium sp., [35] as a possible stimulating agent of Extracellular Polymeric Substance (EPS) excretion [40] by Leptospirillum ferrooxidans and as a kind of mineral surface scavenger/cleaner of the residual EPS after detachment of microorganisms [41] as well as the predatory activity of some protozoa on Thiobacilli. There is no doubt that the advances made, that started with some fundamental work by Costerton [42] and Characklis [43] on the microbial capsula and its importance in cell attachment to solid surfaces, fuelled the as yet unresolved controversy, that dates back more than thirty years [44,45] that has generated the fascinating and sometimes heated dispute between the supporters of direct [46] and those of indirect attack [47,40]. This issue has generated some very interesting and to a certain extent instructive critical reviews of the research work carried out on the topic. Probably the solution lies, as often happens, in a compromise and the papers mentioned above are the first premonitory signs. The mechanism by which microorganisms enhance metal sulphides oxidation and leaching are still surrounded by controversy. Consensus appears to have been momentarily reached on the current lack of evidence as to the existence of an enzyme that justifies the theory of the direct attack. The discovery of the extracellular polymeric substance (EPS) has given rise to some discrepancy in the interpretation of its function between the supporters of the purely electrochemical mechanism [47] and those who uphold a (surface-) chemical process where EPS is the localized environment cell/mineral surface for the action of an energy carrier (e.g. cysteine-based sulphur carrier) produced by some 10
Plenary Lecture
chemically-induced mechanism in the microorganism or for the local artificial increase in the concentration of an electron extracting agent (Fe3+) [46]. The peculiar properties of cysteine [48] appear, in this regard, extremely interesting and further research on this subject may be very rewarding. Consensus on the important role played by the EPS is of paramount importance to the process engineer because this provides vital information for the correct design of the bioreactor, as will be discussed later on. However, there still seems to be some disagreement as to the agents that generate the phenomena evolving within the EPS layer: Fe3+ concentration, hydrogen ion concentration, occurrence of thiol groups (e.g. cysteine). Experimental evidence based on redox potential measurements seems to favour the purely electrochemical mechanism [47]. However, further experiments are warranted in the light of recent contributions suggesting that redox potential at the surface of an electrode differs from the bulk of the solution in which it is immersed [49]. Similarly, further work is required to gain a better understanding of the ability of Thiobacillus ferrooxidans to excrete cysteine or other thiol group compounds. Hence, at least until the enzyme mediating the direct attack of the mineral by the microorganism is discovered, it seems quite appropriate to replace the term “direct attack” with the more cautious "direct contact" coined several decades ago in the very early days of Biohydrometallurgy [50]. Assessment of the significance of EPS is in my opinion a very important issue: the latest results obtained in research work being conducted by my group substantiate its crucial role in bioleaching. During the bioleaching of a gold-bearing pyrite/arsenopyrite concentrate in the new continuous bioreactor designed and constructed by our team, for which a patent is pending, the Biorotor [51] we came across a situation that we had never encountered when using batch bioreactors. The final leach liquor was observed to contain as much as 50 gram.dm-3 total iron, much higher than is consistent with its pH (0.9), with similar amounts of Fe3+, Fe2+. The intriguing point is that conventional analytical methods have proven unsatisfactory, owing to the fact that the iron seems to be encapsulated in something very similar to the recently reported glucuronic acid/iron ion complex containing 53 gram dm-3 [52,53], hence possibly the remnants of the EPS after pyrite corrosion and detachment of the microbial cell. This seems to prevent the precipitation of iron compounds that significantly impair bioleaching performance. Five of these substances’ properties appear to be quite remarkable, from the practical process engineering standpoint: (i) the fact that their chemical composition and surface activity are influenced by the substratum (ii) the fact that they form a particular, enlarged reaction space for the microbial cells, (iii) the ability of the cells to replenish their capsular material in a few hours when they loose it for any reason (for instance owing to mechanical action), (iv) EPS mediate the contact and (v) the microorganism looses its "catalyst" action if, for any reason, it is deprived of the EPS. This latter property appears to be significant [3] from the practical viewpoint: in the STR’s, at high suspension solids concentration, abrasion may seriously affect microbial action by tearing away the EPS from the cells. Two remedies are possible: either to modify the bioreactor design so as minimize abrasion and shear stresses or to design, employing genetic engineering methods, microorganisms resistant to abrasion and shear stresses. The Biorotor was designed in an attempt to pursue the first remedy. The opportunities offered by genetic engineering over the past twenty years, can be likened to medieval alchemists search for the philosophers’ stone with the obvious difference that genetic engineering is a very rigorous and well developed science. And like 11
Plenary Lecture
their search that contributed enormously to the development of modern chemistry, so too genetic engineering has provided a better understanding of the mechanisms underlying microbe-minerals interactions and enabled a far more accurate identification of microbial strains. Genetic engineering aims to develop microorganisms tailor-made for leaching, with fast kinetics and highly resistant to metal ions, the individual mineral. However, it needs to be said that the potential of "wild" or "indigenous" microorganisms has likely not yet been fully exploited through the development f.i. of suitable bioreactors. Genetic engineering may contribute, for instance, to suitably modifying the bacterial genomes of those microorganisms whose attachment to the mineral surface depends on the EPS, so that they are tailor engineered for the particular mineral. The observation, that as commercial biohydrometallurgical operations are not sterile, the risk of modified organisms being released into the environment may discourage or even prohibit the resort to such a technique is quite realistic [54]. A broad range of microbiological investigations has been conducted in the context of metal oxides bioleaching and the recovery of important elements from silicates as exhaustively discussed in a recent review paper by Ehrlich [2]. Attempts to use microorganisms for enhancing the solubilization kinetics of rocks, in particular of silicates, date back to the beginning of the last century, hence well before the discovery of Acidithiobacillus ferrooxidans. The solubilization of leucite, a potassium and aluminium silicate, to extract potassium and aluminium at low cost, was attempted in 1906 by De Grazia and Camiola [55] using molds. While the interesting properties of Acidithiobacillus ferrooxidans, were being investigated, "silicate" solubilizing bacteria were claimed to have been isolated from agricultural soils [56]. It seems that the silicate bacteria claimed by Aleksandrof and his school are in effect strains of Bacillus circulans, [57,58] but research in this area probably still has a long way to go before commercial applications can be contemplated. Some controversy still surrounds this discovery and compared to the advances made with Thiobacilli little progress has been reported in this field. Over the past fifty years no major advances have been achieved in research on the use of molds or, at least, they do not offer any promising commercial prospects. One of the major drawbacks of this technology is the dramatic volume of biomass involved and the practical problems posed by its handling. Some interesting proposals for applications of silicate solubilizing bacteria concerned the beneficiation of bauxites [59], but again no industrial applications have been reported. An interesting niche is represented by the investigations into the microbial recovery of manganese from manganese oxide [60-62]. Research carried out in this field is unravelling the mechanisms of microbial oxides reduction, and the results obtained so far seem quite encouraging. This area warrants further investigation as it offers interesting commercial prospects, for instance for recovering manganese from ocean nodules. 3.2.2 Microbe/solid matter interactions As far as solid matter is concerned, a distinction needs be made between rocks (minerals) and solid matter other than rocks. The latter is concerned more with environmental applications and has been dealt with above. Mineral bioleaching is a subject that embraces several topics. As sulphide minerals oxidation is a physico-chemical process, a better understanding of how the presence of microorganisms enhances electron transfer from the mineral to the end acceptor is crucial for optimising any bioleaching operation. 12
Plenary Lecture
The amenability of a given mineral to oxidation is related to the solid-state physics of both its bulk and its surface. A number of recent contributions have provided an insight into the importance of the semiconductor nature of the minerals involved and the modifications produced by the presence of foreign elements in the crystal lattices [22,6366]. Molecular orbital theory, valence bond theory and mineralogy may provide a decisive contribution to elucidating the solubility of metal sulphides in bioleaching systems. Thus, based on their solid state physics, Sand et al. [40] were able to classify metal sulphides into two main groups: the first, consisting of pyrite, molybdenite and tungstenite, and the second of sphalerite, galena, chalcopyrite, hauerite, orpiment and realgar, exhibiting strong differences in solubility in acid. One typical case is sphalerite: mining engineers and metallurgists are well aware of the fact that quite often sphalerite is not a stoichiometric compound of zinc and sulphur but contains dissolved iron in proportions that usually differ from one orebody to another. These varieties are called "marmatites" and have been widely investigated [67]. The importance of the proportion of iron dissolved in the crystal lattice of sphalerite had already been recognized in the mid 1950’s by flotation engineers. Now, in the light of considerations that have emerged in recent years [64], the importance of iron has been fully elucidated. The plot reported by Crundwell [68] demonstrating that the rate of reaction for the oxidative dissolution of marmatites is a linear function of the iron concentration in the zinc sulphide lattice, shows dramatically how iron affects the electrochemical, and hence the leaching behaviour, of marmatites. Similar considerations were made for pyrite [22]. Since the elements dissolved in the lattice are kinds of fingerprints of the mineral, it is clear that its origin warrants special attention and may well explain the differences in leaching behaviour. Mechanisms based on the valence bond theory have recently been proposed [46,64] for the oxidation of pyrite sulphur moiety by iron(III) ions and, though requiring further refinements, they are worthy of mention in that they are on the right track to solving the problem. Further research will likely provide the information necessary for process modulation. The surface structure and composition of minerals and their modifications during the bioleaching process as well as a kind of "activation" by some metals, like silver, have been recognized in some instances as important factors for the evolution of the process. Chalcopyrite is, in this respect, quite typical since the acidic ferric sulphate leaching or bioleaching evolve according to a well-documented pattern [27,69]. Copper dissolution proceeds in two phases: the initial phase is characterized by a relatively high dissolution rate, which can be expressed by a parabolic law. Over time, the rate decreases pronouncedly, the dissolution rate being represented by a gently sloping straight line. The dissolution rate of the first phase can only be restored by regrinding the leached residue, which usually contains no less than 50% of the copper in the initial feed. Among the explanations advanced for this behaviour, the one based on electroanalytical observations [70] seems to have been confirmed by XPS investigation of surface layers of bioleached chalcopyrite [71]. According to this model, during the first phase, a diffusion layer of copper-depleted chalcopyrite is formed on the mineral surface and it is this layer that governs the subsequent dissolution rate. The presence of silver in the reacting suspension has a catalytic effect [72] that strongly enhances copper solubilization and is explained by the formation of conductive
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compounds in the retarding barrier, which neutralize its passivating effect by acting as channels for electron flow. This is just one of many significant, still debated examples, of the interactions of a multiplicity of physico-chemical, electrochemical and solid-state physics factors. The intriguing micrographs showing the preferential adhesion of microbial cells to areas where emergence of dislocations is visible [40, 44, 73] seem to justify further and more sophisticated research on this subject. The resort to instruments that, like XPS, provide information on the chemical bonding between elements in the surface layers, may unveil the importance of increasing the surface area-to-bulk volume ratio and the extent to which ions in solution interact with the mineral surface and affect oxidation and solubilization. The importance of chemical composition is also significant, as sometimes the elements passing into solution are toxic to the microorganism, thus requiring continuous adaptation. The inhibitory action of these elements may affect the growth kinetics of the microbial populations and this again relates to the physiology of microorganisms. 3.2.3 Bioleaching kinetics In the light of the above considerations it is now possible to briefly discuss what may be considered one of the key issues in biohydrometallurgy: the bioleaching kinetics from a process engineering standpoint. As this audience is well aware, numerous researchers have explored this topic, most of whom are chemical engineers or hydrometallurgists, and at a rough estimate more than 50 papers have been published on the subject over the past sixty years. Further research is expected to provide useful data for modulation of the process. The dramatic increase in mineral sulphides oxidation and solubilization kinetics produced by the pseudocatalytic mediation of Thiobacillus sp. organisms has roused great interest in practitioners and researchers in extractive metallurgy since the very beginnings of biohydrometallurgy. In effect, it was just around this time that a greater awareness of environmental issues was developing, which was soon to place increasing constraints on new industrial projects. Up until then extractive metallurgy had been dominated by pyrometallurgical processes, which had a negative environmental impact. Clearly the development of mathematical models capable of predicting the economic performance of the new processes became a pressing need. As shown in Figure 2, the performance of a bioleaching process depends on a multiplicity of factors and, as pointed out in recent years [27,74], its kinetics are a combination of microbial growth kinetics, microbe-minerals and chemical compounds interaction kinetics and particulate solids oxidation and dissolution kinetics (a process that involves several branches of science and technology). The early models, which already endeavoured to develop a general theory of bioleaching, were relatively simple [69,75-77] but deserve mention in recognition of the originality and the initiative of those researchers. Developments in subsequent years are well documented in several excellent reviews [74,78,79] and numerous of proposed models (fifteen) are exhaustively summarized in a recent review paper [80]. The above-mentioned classification of metal sulphides proposed by Sand et al. [40] and the continuing speculation surrounding the importance of the glycocalyx in the biosolubilization process, would appear to definitively preclude the possibility of developing a general kinetic theory that embraces the entire field of bioleaching. At this point in time it seems expedient to revisit the existing models in the light of the advances in the fundamentals: a deeper insight into some of these may well lead to their unification. 14
Plenary Lecture
There seems to be general consensus that bioleaching kinetics is the result of the combination of several subprocesses, the most important being ferrous iron oxidation, bacterial growth (planktonic and sessile), and mineral sulphide dissolution reaction. The kinetic models The four recently published reviews mentioned above, propose a classification of the kinetic models. Three of them concern ferrous iron oxidation and mineral bioleaching, the fourth ferrous iron oxidation. Crundwell [81], proposes a classification criterion based on the relative importance of what are unanimously defined as "contact leaching" and "indirect leaching" [46] mechanisms and refers to two categories: (i) models that postulate a well defined mode of evolution of the leaching process (shrinking-core, propagating-pore) and derive bacterial growth; and (ii) models that postulate a well defined bacterial growth law (rate of growth related to attachment to- and detachment from the solid substrate, Monod law, the logistic equation) and derive the evolution of the leaching process. Haddadin et al. [80] based their classification (concerning both ferrous iron oxidation kinetics and mineral bioleaching) on the mass balance of each reactional system and on the underlying assumptions of each model and identified three categories: (i) well agitated reactors operating on a liquid phase; (ii) well agitated reactors operating on liquid and solid phases; (iii) the so-called "biofilm reactors". Hansford [74] considers three classes: (i) empirical models, based on the logistic equation; (ii) models based on attachment of microorganisms; (iii) the two subprocess mechanism (bacterial ferrous iron oxidation and chemical ferric leaching of the sulphide mineral, assumed to be an electrochemical process for which the Volmer-Butler equation holds) for pyrite bioleaching. The implications of this mechanism, as reported by Hansford [74] are rather interesting, insofar as it seems to be able to explain the findings that L. ferroooxidans is the dominant bacterial species in the bioleaching of arsenopyrite and pyrite [28,29] and that arsenopyrite is preferentially leached ahead of pyrite [82]. However, none of these bioleaching kinetics models take into account the population balance approach for describing in quantitative terms, by means of suitable distribution functions, the influence of material properties distribution on the overall behaviour of biohydrometallurgical systems, exhaustively described by extractive metallurgists in the late 1970’s [83,84]. The significance of the material properties appear to have been first understood by researchers investigating coal biodepyritization in relation to the size distribution of pyritic coal [85] but the population balance approach was not adopted. Crundwell [81] deserves recognition for his model based on the population balance for particulate leaching, for the bacterial cell number and material balance describing solution reactants and products. This model was found to provide an excellent fit to experimental data reported by other researchers for pyrite bioleaching, although he used the shrinking-core rather than the more realistic propagating-pore mechanism [74]. It should however be pointed out that the population balance approach can also have pitfalls, associated for instance with the practical difficulty, of which mineral processing practitioners are well aware, of developing a realistic function for the grain size distribution of a mineral powder. These difficulties can increase the complexity of such a model making its application impracticable. The kinetics of ferrous iron oxidation has long been recognized as playing a key role in bioleaching and several papers have been published on this subject over the past sixty 15
Plenary Lecture
years. I will not review every single paper here but will limit myself to a few general remarks, spotlighting the main features of this history, as they provide a good description of the evolution of biohydrometallurgy. Over the years the significance of the following factors in bioleaching processes listed in chronological order - has been recognized: Total initial iron concentration Temperature Type of culture (continuous or batch) Wall growth Total iron concentration Inhibition by ferric iron Threshold concentration of ferrous iron Product inhibition pH Bacterial concentration (mass or numeric) Bacterial decay Maintenance Carbon dioxide transfer Nemati et al. [86] reviewed the work carried out on the kinetics of ferrous iron oxidation by Thiobacillus ferrooxidans distinguishing two major classes: (i) freely suspended cells; (ii) immobilized cells. Whereas, for the immobilized cells processes, these authors compared the performances of different types of reactors (packed-bed, fluidised-bed, rotating biological contactors), in the paper by Pesic et al. [87] the experimental set-up consisted of a thermostated 300 cm3 plexiglass cell containing 250 cm3 of ferrous solution and Thiobacillus ferrooxidans inoculum and provided with a magnetic stirrer: no information was provided on the rpm of this stirrer. In another paper [88] ferrous conversion kinetics was experimentally determined using a set-up consisting of a small, thermostated, reaction cell (volume 20 cc) containing 10 cc of reaction liquid plus cell suspension, sparging air through the liquid contained in the cell at a flow rate of 8,3 cm3.s-1. This set-up can be regarded as an "ideal" reactor, where the bacterial cells are very likely only subjected to a minimum of mechanical stress. It is possible that experimental conditions in both cases ensured a fairly quiescent environment thus avoiding undue stress to the bacterial cells. These operating conditions differ significantly from those existing in the particulate mineral suspension of an STR fitted, for instance, with a Rushton-type turbine. Only recently, however, a mechanistically-based model adopting the initial rate method and relying on Michaelis-Menten kinetics [79,88] and a model based on Ingledew’s chemiosmotic theory, on the electrochemical theory and on the Monod and Michaelis- Menten models, take into account all of the above influencing factors and appear to provide a satisfactory fit to the bacterial ferrous iron oxidation data reported in the literature [74]. This is undoubtedly an important achievement, though the data fit refers to experiments carried out in a variety of bioreactors where, in terms of reactor dynamics, optimum operating conditions may well not have been ideal for bacterial cells. So, all the models for ferrous iron bioleaching kinetics or for mineral bioleaching kinetics appear to neglect the influence of reactor dynamics, which, as far as I am aware, was only mentioned [80] in relation to a paper [89] dealing with the geometry and operating characteristics of an air–lift bioreactor (Pachuca tank) designed by researchers of Bergbauforschung and still being used in my laboratory.
16
Plenary Lecture
Practically all the authors who have provided an overview of bioleaching kinetics models agree as to the difficulty of generalizing the results of the individual models, on account of the numerous influencing factors that characterize each substrate, such as mineralogy, solid-state physics, chemical composition, particle size, specific gravity, electrochemistry, and the microbial population (temperature, pH, strains, associated microbial strains…). Moreover, several authors have demonstrated that strong shear stresses are generated in conventional bioreactors (STR’s and ALR’s) during agitation of the suspension and that, beyond a certain threshold, abrasion produced by solid particles has to be taken into account. Such situations may become a limiting and determining factor seriously affecting both bacterial growth and bacterial adhesion and ultimately process kinetics. Quantification of these effects is an open and very likely promising research field, although I still doubt that general theories can be formulated. Probably, the most practicable and profitable way is to design suitable bioreactors where these effects are minimized by suitably selecting the proper operating conditions. There is general consensus that bioleaching process kinetics is directly dependent upon the number of active microbial cells present in the system. Already about twenty years ago this number had been estimated to be in the order of magnitude of 1012 cells.cm-3 [85,90,91] which, as far as I know, has not yet been achieved. It is therefore reasonable to attempt to maximize microbial growth. The question then arises as to what is the maximum number of microorganisms that can be achieved and if the bioreactors currently in use are the most suitable for attaining this population density in steady state regime and how such a reactor should be designed. The problems are very similar to those confronted by mineral beneficiation researchers and flotation plant designers and operators for about 100 years, from the very early days of flotation technology. Research has played a very important role in elucidating the problems, but no general theory could be developed. Basically, the flotation process is the same for any ore, but each ore requires equipment, reagents, and physico-chemical environment to be properly adjusted. Process performance can only be predicted by a wise combination of fundamentals (that, except for the biological agent, are practically the same as in biohydrometallurgy) and of bench- and laboratory scale pilot testing using devices that simulate commercial operation, that provide the experimental numeric parameters for the mathematical expressions for describing the specific process kinetics - supplemented by an up-to-date knowledge of microbiology. The future progress of Biohydrometallurgy much depends on the solutions of these problems: the fascinating aspect but also the intriguing feature of biohydrometallurgy is its multidisciplinarity. It is fascinating because it reveals how intimate the liaison between the various branches of science and technology can be; intriguing, because this technology requires a working knowledge of several disciplines, that are sometimes so different from one another and that cannot be reasonably mastered by a single specialist. Real advances will only be made possible by the cooperation of people skilled in the various facets of biohydrometallurgy. The timely exchange of information and a very complete documentation are the prerequisites for gaining an identity that this new technology seems to have not yet achieved. The importance of up-to-date documentation will prevent researchers from expensive and time-wasting repetition of investigations successfully carried out elsewhere. What I am going to say now may sound trivial, but it
17
Plenary Lecture
should be remembered that a well-documented failure is as useful as a successful achievement. 4.
A BIOHYDROMETALLURGY SOCIETY Based on his experience, the author would like to emphasize the need to establish a Biohydrometallurgical Society, to act as a point of reference to all those, from industry and academia, who are involved in implementing and further developing the technology. The Society should also provide a forum for information flow to decision makers in industry, about the potentials of biohydrometallurgy. Though it relies on the exploitation of the complex synergies between microoorganisms and minerals, this technology, when properly applied, is simple to implement, operationally stable and cost-effective. One task of the Society might be, for instance, the preparation and further updating of a Recommended Standard Terminology and Nomenclature for Biohydrometallurgy, like the one published several years ago by the Institution of Chemical Engineers Fluid Mixing Group [92]. This idea came to mind when I was reading a comprehensive review written by a distinguished microbiologist, who proposed an important terminology, justifying his intervention with the fact that the related topic had been studied by researchers from diverse cultural backgrounds other than microbiology. Last, but not least, the Society would be able to benefit from the experience and knowledge of all those involved in industry and academia to develop new university curricula aimed at training a new generation of specialists. REFERENCES 1. 2. 3. 4.
Ehrlich, Hydrometallurgy, 59 (2-3) (2001) 127. Ehrlich, Minerals and Metallurgical Processing, 19 (4) (2002) 220. Rawlings, D. E., Ann. Rev. Microbiol., 56 (2002) 65. Brierley, J.A., Brierley, C.L., and Goyak, G.M., Fundamental and Applied Biohydrometallurgy, Lawrence, R.W., Branion, R.M.R. and Ebner, H.G. (Eds.), Elsevier, Amsterdam (1986) 291. 5. Tsezos, M., Microbial Mineral Recovery, Ehrlich, H. L. and Brierley, C. L., (Eds.), McGraw-Hill Publishing Company, New York, (14) (1990) 325. 6. Pümpel, T., Ebner, C., Pernfuß, B., Schinner, F., Diels, L., Keszthelyi, Z., Stankovic, A., Finlay, J. A., Macaskie, L. E., Tsezos, M., and Wouters, H., Hydrometallurgy, 59 (2-3) (2001) 383. 7. Hatzikioseyian, A., Tsezos, M., and Mavituna, F., Hydrometallurgy, 59 (2-3) (2001) 395. 8. van Loosdrecht, M.C.M. and Zehnder, A.J.B., Experientia, 46 (1990) 817. 9. van Loosdrecht, M.C.M., Norde, W., Lyklema, J. and Zehnder, A.J.B., Aquatic Sciences, 52 (1) (1990) 103. 10. van Loosdrecht, M.C.M., Lyklema, J., Norde, W., Gosse, S. and Zehnder, A.J.B., Appl. Environm. Microbiology, 53 (8) (1987) 1893. 11. Madgwick, J.C., and Ralph, B.J., Conference Bioleaching, Schwartz, W. (Ed.) Weinheim (Germany) Verlag Chemie, (1977) 85. 12. Tsezos, M., Hydrometallurgy, 59 (2-3) (2001) 241. 13. Volesky, B., Hydrometallurgy, 59 (2-3) (2001) 203. 14. Schnell, H.A., Biomining: Theory, Microbes and Industrial Processe, Rawlings, D. E. (Ed.), Springer, Berlin, (11) (1997) 21. 15. Pasteur, L., La révue scientifique, 2e Sér. 1 (4) (1871) 73. 18
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16. van Answegen, P.C. and Marais, H.J., MINERAL BIOTECHNOLOGY, Kawatra, S.K. and Natarajan, K.A. (Eds.), Littleton, Colorado, USA Society for Mining, Metallurgy, and Exploration, Inc., (2001) 121. 17. Dew, D.W., Lawson, E.N., and Broadhurst, J.L., Biomining, Rawlings, D.E. (Ed.), Berlin, Springer, (1997) 45. 18. Taxiarchou, M., Adam, K., and Kontopoulos, A. Hydrometallurgy, 36 (2) 1994 169. 19. Nicholson, H.M., Smith, G.R., Stewart, R.J., Kock, F.W., and Marais, H.J., Biomine ’94, Australian Mineral Foundation: Glenside, South Australia, 1994 2-1. 20. Loi, G., Mura, A., Trois, P., and Rossi, G., Memorie dell'Associazione Mineraria Sarda – Iglesias, I (1) (1994a) 29. 21. Loi, G., Mura, A., Trois, P., and Rossi, G., Fuel Processing Technology, 40 (1994b) 26. 22. Rossi, G., Fuel, 72 (12) (1993) 1581. 23. International Scientific Committee of the Fourth Intrntl.Symp. on Biological Processing of Fossil Fuels, Rossi, G. (Guest Ed.), Coal Processing Technology, 40 (1994), 379. 24. Markosyan, G.E., Biol. Zh. Armenii, (1972) 25. 25. Brierley, C.L., and Murr, L.E., Science, 179 (1973) 488. 26. Rawlings, D.E., Biomining: Theory, Microbes and Industrial Processe, Rawlings, D. E. (Ed.), Springer Berlin, (1997) 229. 27. Rossi, G., Biohydrometallurgy, McGraw-Hill GmbH, Hamburg, (1990). 28. Rawlings, D. E., Tributsch, H., and Hansford, G. S., Microbiology, 145 (1999) 5. 29. Boon, M., Brasser, H. J., Hansford, G. S., and Heijnen, J. J., Hydrometallurgy, 53 (1) (1999) 57. 30. Breed, A. W., Dempers, C. J. N., Searby, G. E., Gardner, M. N., Rawlings, D. E., and Hansford, G. S., Biotechnology and Bioengineering, 65 (1) (1999) 44. 31. Battaglia-Brunet, F., d'Hugues, P., Cabral, T., Cezac, P., Garcia, J. L., and Morin, D. 1998, Minerals Engineering, 11 (2) (1998) 195. 32. Jerez, C. A. 1997, Biomining: Theory, Microbes and Industrial Processes. Rawlings, D. E. (Ed.), Springer, Berlin (1997) 281. 33. Kelly, D. P. and Wood, A. P., International Journal of Systematic and Evolutionary Microbiology 50 (2000) 511. 34. Harrison, A. P. J., Archives of Microbiology, 131 (1982) 68. 35. Hallmann, R., Friedrich, A., Koops, H., Pommering-Röser, A., Rohde, K., Zenneck, C., and Sand, W., Geomicrobiology Journal, 10 (1992) 193. 36. Johnson, B. and Roberto, F. F., Biomining: Theory, Microbes and Industrial Processe, Rawlings, D. E. (Ed.), Springer Berlin, (13) (1997) 260. 37. Johnson, D. B., FEMS Microbiol. Ecol., 35 (1998) 307. 38. Johnson, D.B., Hydrometallurgy, 59 (2-3) (2001) 147. 39. Johnson, B., Bacelar, Nicolau P., Okibe, N., Yahya, A., and Hallberg, K.B., Biohydrometallurgy - “Fundamentals, Technology and Sustainable Development”, Proceedings of the International Biohydrometallurgy Symposium IBS-2001, Part A, Ciminelli, V. S. T. and Garcia Jr., O. (Eds.), vol. A, Elsevier Amsterdam, (2001) 461. 40. Sand, W., Gehrke, T., Jozsa, P.-G. and Schippers, A., Hydrometallurgy, 59 (2-3) (2001) 159. 41. Sand W, Jozsa P-G, and Schippers A., Biohydrometallurgy and the Environment toward the Mining of the 21st Century, Amsterdam, Elsevier, Vol. A (1999) 27. 42. Costerton, J. W., Irvin, R. T., and Chen, K.-J., Ann Rev Microbiol, 35 (1981) 299. 43. Characklis, W. G., Biotechnology and Bioengineering, 23 (1981) 1923. 44. Berry, V. K. and Murr, L. E., Metallurgical Transactions, B, 6B, (1975) 488. 19
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45. Bennett, C. and Tributsch, H., J. Bacteriology, 134 (1) (1978) 310. 46. Tributsch, H., Hydrometallurgy, 59 (2-3) (2001) 177. 47. Crundwell, F.K., Biohydrometallurgy - “Fundamentals, Technology and Sustainable Development”, Proceedings of the International Biohydrometallurgy Symposium IBS2001, Part A, Ciminelli, V.S.T. and Garcia Jr., O. (Eds.), Elsevier, Amsterdam (2001) 149. 48. Rojas-Chapana, J. A., and Tributsch, H., 35 (2000) 815. 49. Nicol, M. J. and Lázaro, I., Hydrometallurgy, 63 (1) (2002) 15. 50. Silverman, M.P., and Ehrlich, H.L., Advan. Appl. Microbiol., 6 (1964) 153. 51. Loi, G., Trois, P., and Rossi, G., Biohydrometallurgical Processing, Vargas, T., Jedrez, C.A., Wiertz, J.V. and Toledo, H. (Eds.), University of Santiago, Chile, Vol. 1, (1995) 263. 52. Gehrke, T., Hallmann, R., Kinzler, K., and Sand, W., Appl. Environ. Microbiol., 65 (1997) 159. 53. Gehrke, T., Telegdi, J., Thierry, D., and Sand, W., Appl. Environ. Microbiol., 64, 7 (1998) 2743. 54. Rawlings, D.E., Hydrometallurgy, (2-3) 59 (2001) 187. 55. De Grazia, S. and Camiola, G., Le stazioni sperimentali agrarie italiane, 39 (9) (1906) 829. 56. Aleksandrov, V.G., Kiev Izv. Akad. Nauk. S. S. S. R., (1958) 62. 57. Tesic, Z.P. and Todorovic, M.S., Zemljiste i biljka, 1 (1) (1952) 3. 58. Tesic, Z.P. and Todorovic, M.S., Zemljiste i biljka, 8 (1-3) (1958) 233. 59. Groudeva, V.I. and Groudev, S.N., 5th International Congress of ICSOBA, Zagreb, Yugoslavia, Academie Yugoslave des Sciences et des Arts, 83 (18) 257. 60. Remezov, V. D., Agate, A. D., and Yurchenko, V. A., Biogeotechnology of Metals – Manual, Karavaiko, G. I., Rossi, G., Agate, A. D., Groudev, S. N., and Avakyan, Z. A. (Eds.), Centre for International Projects – GKNT Moscow U.S.S.R., (1988) 304. 61. Ehrlich, H. L. (and Rossi, G.), Microbial Mineral Recovery, Ehrlich, H. L., and Brierley, C. L. (Eds.), (7), MacGraw-Hill Publishing Company, New York, (1990) 149. 62. Ehrlich HL., Biohydrometallurgical Technologies, Torma, A. E., Apel, M. L., and Brierley, C. L. (Eds.), v. 2, The Minerals, Metals & Materials Society, Warrendale, Penna, USA, (1993) 415. 63. Pridmore, D.F., and Shuey, R.T., American Mineralogist, 61 (1976) 248. 64. Crundwell, F.K., AIChE Journal, 34(7) (1988) 1128. 65. Nesterovich, L. G., Biogeotechnology of Metals - Manual. Karavaiko, G. I., Rossi, G., Agate, A. D., Groudev, S. N., and Avakyan, Z. A. (Eds.), Centre for International Projects GKNT, Moscow, U.S.S.R., (1988) 101. 66. Thomas, B., Ellmer, K., Bohne, W., Rohrich, J., Kunst, M., Tributsch, H., Solid State Communications, 111 (1999) 235-240. 67. Kullerud, G., Norsk Geologisk Tidsskrift, 32 (1953) 61. 68. Crundwell, F.K., Hydrometallurgy, 21(2) (1988) 155. 69. McElroy, R.O. and Bruynesteyn, A., Metallurgical Applications of Bacterial Leaching and Related Microbiological Phenomena, Murr.L.A., Torma, A. E., and Brierley, J. A. (Eds.) Academic Press, New York (1978) 441. 70. Ammou-Chokroum, M., Cambazoglu, M. and Steinmetz, D., Bull. Soc. Fr. Miner. Cristallogr. 100 (1977) 161. 71. De Filippo, D., Rossi, A., Rossi, G., and Trois, P., International Biotechnology Symposium Proceedings, Durand, G., Bobichon, G., and Florent, J. (Eds.), J. Soc. Française de Microbiologie, Paris, France (1988) 1131. 20
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72. Miller, J.D., and Portillo, H.Q., Proceedings XIII Int. Min. Proc. Cong., Vol. 2 (A), Laskowski, J. (Ed.), Elsevier Scientific Publishing Comp., (1979) 851. 73. Rojas-Chapana, J.A., and Tributsch, H., Biohydrometallurgy and the Environment toward the Mining of the 21st Century, Amsterdam, Elsevier, Vol. 2 (1999) 597. 74. Hansford, G. S., Biomining. Rawlings, D. E., Springer: Berlin, (8) (1997) 153. 75. Torma, A. E., CANMET, Ottawa, Canada (1985) 5. 76. Torma, A. E. and Panneton, J. J., Ministère des Richesses Naturelles - Direction Générale des Mines-Centre de Recherches Minérales, Québec, Canada BLS-1 (1973). 77. Legault, G., and Torma, A.E., 421ème Congrès Annuel, Association Canadienne Française pour l'Avancement des Sciences, Province de Québec (1974) 1. 78. Barrett, J., Hughes, M. N., Karavaiko, G. I., and Spencer, P. A., Metal Extraction by Bacterial Oxidation of Minerals. Ellis Horwood Limited, NewYork. (1993) 103. 79. Nemati, M, and Webb, C., Biotechnology Letters, 20 (9) (1998) 873. 80. Haddadin, J., Dagot, C., and Fick, M., Enzime and Microbial Technology, 17 (1995) 290. 81. Crundwell, F.K., The Chemical Engineering Journal, 54 (1994) 207. 82. Miller, D. M. and Hansford, G. S., Minerals Engineering 5 (7) (1992) 737. 83. Sepulveda, J. E. and Herbst, J. A., AIchE Symposium Series, 74 (173) (1978) 41. 84. Herbst, J. A., Rate Processes of Extractive Metallurgy. Sohn, H. Y. and Wadsworth, M. E. (Eds.) Plenum Press New York, (2) (1979) 53. 85. Andrews G.F., Bioprocessing of Coal Workshop III, Idaho National Engineering Laboratory, Idaho Falls, ID, USA (1988) 234. 86. Nemati, M., Harrison, S.T.L., Hansford, G.S. and Webb, C., Biochemical Engineering Journal, 1 (1998) 171. 87. Pesic, B., Oliver, D.J., and Wichlacz, P., Biotechnology and Bioengineering, 33 (1989) 428. 88. Nemati, M. and Webb, C., Biotechnology and Bioengineering, 53, (5) (1997) 478. 89. Beyer, M., Ebner, H.G., and Klein, J., Appl. Microbiol. Biotechnol., 24 (1986) 342. 90. Andrews G.F. and Quintana J., First International Symposium on Biological Processing of coal, E.P.R.I. Palo Alto, California, U.S.A., (1990) 5-69. 91. Stevens, C. J., Noah, K. S., and Andrews, G. F., Fuel, 72 (12) (1993) 1601. 92. Anonymous, The Chemical Engineer, (1980) 557.
21
C HAPTER 1 Bioleaching Applications
15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
A novel bio-leaching process to recover valuable metals from Indian Ocean nodules using a marine isolate Amitava Mukherjeea, Ashok M. Raichura, Jayant M. Modakb, K.A. Natarajana a
Department of Metallurgy Department of Chemical Engineering Indian Institute of Science, Bangalore – 560012, India b
Abstract A novel bio-leaching process to recover valuable nonferrous metals from Indian Ocean nodules at near neutral pH and ambient temperature is presented in this paper. Poly-metallic manganese nodules contain a lucrative resource for valuable strategic metals like Cu, Co and Ni. In view of rapid depletion of land -based resources of copper and nonavailability of significant resources of Ni and Co in India, these nodules have an enormous potential to effect future metal extraction trends in the country. The significant role of indigenous microorganisms in solubilization of valuable trace metals from the nodules was studied in detail. A marine organism was isolated from the nodules. The isolate was shown to be a Gram-positive heterotroph of Bacillus species. It was grown in artificial seawater medium. The growth studies of the isolate with respect to pH, temperature and salinity of the medium proved that, the isolate grows well at neutral pH, 30°C temperature and 0.25M NaCl concentrations. A growing culture of the isolate as well as cell-free spent liquor containing metabolites was employed to recover transition metals from the nodules. A few chemical leaching experiments in moderate acidic environment were also conducted to generate baseline data. The bioleaching of the nodules was carried out in two different ways (a) leaching during the growth of the isolate and (b) leaching by spent liquor after removal of fullygrown cells. The leaching efficiency was observed to be more or less the same in both the cases confirming that metabolites produced during growth of the microorganism, played a pivotal role in the leaching process. Around 50% cobalt and 30% of the Cu and Ni came out in the leach liquor at pH of 8.2 in a four-hour interval, indicating that the effect of metabolites was specific to cobalt recovery in comparison to copper and nickel. Kinetic studies revealed that the metal recovery stabilized after four hours. The results for cobalt were quite comparable to those achieved in highly acidic conditions. A strong reducing environment is required to break down the MnO2 and/or Fe-oxide matrix encapsulating the remaining part of the transition metals. So, the effect of adding a reductant like starch in spent liquor of the isolate was investigated. Starch enhanced the recovery of all the transition metals to about 80%-85%. The effect of pulp density and pH of the leach liquor on the bio-leaching process was investigated. The recovery was observed to be almost independent of size fraction of nodules over a wide range. 25
Bioleaching Applications
1.
INTRODUCTION In view of continuous depletion of land-based resources along with increasing consumption of valuable metals in India, development of environment friendly technologies for tapping alternate sources of metals has gained importance lately. One of them is recovery of strategic metals Cu, Ni and Co from polymetallic Indian Ocean nodules, by biological processing. Fuerstenau and Han (1983) extensively reviewed processing and extraction of valuable metals from manganese nodules discussing both hydro and pyrometallurgical routes. High porosity of the nodules resulting in high moisture content coupled with polluting effluent gases pose major hurdles in pyrometallurgical processing of the nodules. Therefore, hydrometallurgical techniques are potentially viable for extraction of metals from nodules. But often, slow kinetics and poor recovery with dilute acids and corrosiveness of the concentrated ones restrict application of hydrometallurgical extraction. Researchers have studied addition of reducing agents with either mineral acids or ammonia. The reducing environment, so created, enhances leaching, by breaking up the nodule matrix occluding the valuable metals (Niinae et al 1996; Kanungo et al 1988; Jana et al 1999; Trifoni et al 2001; Zhang et al 2001). These processes have varying degrees of success. However, most require high temperature pretreatment and/or costly, corrosive reagents to obtain a sizeable amount of metal recovery with favourable kinetics. As the nodules are low-grade ores of Cu, Co and Ni, use of costly chemical reagents as reducing agents may not be economically viable for large-scale runs. Ehrlich and his co-workers have studied the biogenesis and microbiology associated with Atlantic Ocean nodules extensively over the last four decades. Ehrlich (1963) isolated and characterized the Mn-reducing organism Bacillus 29 from the Atlantic Ocean nodules. Growing cultures of Bacillus 29 are able to reduce MnO2 aerobically and anaerobically using glucose as an electron donor. However, microbial ecology of the Indian Ocean nodules has not been studied in detail until now; utilizing microorganisms isolated from the nodules themselves to leach valuable metals from the nodules, still remains an unexplored route. In the recent past, researchers have been looking into bioprocessing as an alternative route of metal recovery from the nodules. Konishi et al (1997) showed nodule leaching by acidophilic sulphur oxidizing bacteria and thermophilic A. brierleyi. A. Kumari and Natarajan (2001) have been able to extract valuable metals from the nodules by electrobioleaching using acid producing chemolithotrophs like Acidithiobacillus ferrooxidans and Acidithiobacillus thiooxidans. But these processes employed a highly acidic environment supplemented with thermal or electrical energy for recovering a sizeable amount of metals. Major objectives of our present work were to isolate an organism from Indian Ocean nodules, characterize and grow the isolate, and ultimately use the isolate or its growth products for solubilisation of Cu, Co and Ni from the nodules. The recovery of metals through a biological route is compared with that of chemical leaching using dilute acids. The effect of reducing additives in the supernatant, pH of the leaching medium and pulp density were also investigated.
26
Bioleaching Applications
2.
MATERIALS AND METHODS
2.1 Materials Ocean nodule sample was collected from the bed of the Indian Ocean by the National Institute of Oceanography, Goa, India. The sample was ground in a mortar and pestle and sieved to obtain appropriate size fractions. The partial chemical composition of the different size fractions of the as-received sample is presented in Table 1. Phases revealed in the X-ray diffraction pattern of the nodule sample are depicted in Table 2. Table 1. Partial chemical analysis of the ocean nodule
Table 2. XRD analysis of the ocean nodules
Quartz
Chemical composition SiO2
1.15
Asbolan
NiMn2O3(OH)4
Ni
1.12
Co
0.16
Cobalt manganese oxyhydroxide
(Co, Mn) OOH
Fe
8
Element
Wt %
Mn
22
Cu
Phases present
2.2 Microorganisms 2.2.1 Isolation of the unidentified marine species The marine bacteria occurred inherently on the nodule sample as a spore-former. To remove the surface contaminants and germinate the spores, the nodule sample was boiled in water for 30 minutes (Ehrlich, 1963). After boiling, hot water was removed by decantation. The lump of nodule was transferred to an autoclaved porcelain mortar (outer diameter, 150mm) in an inoculating hood, which had been pre-sterilized for at least 30 min by UV rays. The lump was pulverized by pounding in the mortar with a sterilized pestle. One g of this nodule powder was added aseptically to 20ml of sterilized artificial seawater nutrient broth whose composition is given in Table 3. Henceforth the medium will be referred to as ASWNB medium. ASWNB with the nodule powder was incubated at 37°C for 24 hrs. From this enrichment media a loopful of the inoculum was streaked on to artificial seawater nutrient agar plates and incubated at 37°C for 24hrs. The next day round, creamy, smooth surface colonies were found growing on the same medium. A colony was picked from the plate and sub-cultured several times on the same medium to finally obtain the pure strain of the marine isolate. Periodic streaking was done to check for the purity of the isolated strain. Henceforth the bacterium will be referred to as "marine isolate". Table 3. Composition of Artificial Sea water nutrient broth (ASWNB) Chemical components
Wt. in g. in 100 ml of distilled water
Sodium chloride
28.13
Potassium chloride
0.77
Calcium chloride dihydrate
1.60
Magnesium chloride hexahydrate
4.80
Sodium bicarbonate
0.11 27
Bioleaching Applications Chemical components
Wt. in g. in 100 ml of distilled water
Magnesium sulphate heptahydrate
3.5
Peptone
5.0
Beef extract
3.0
Sodium chloride
28.13
Potassium chloride
0.77
Calcium chloride dihydrate
1.60
Magnesium chloride hexahydrate
4.80
Sodium bicarbonate
0.11
Magnesium sulphate heptahydrate
3.5
Peptone
5.0
Beef extract
3.0
2.2.2 Growth of the marine isolate The marine isolate was grown in ASWNB in 250-ml baffled Erlenmeyer flasks at 30°C on a rotary shaker (200 rpm). Ten percent of an active inoculum (from the late exponential phase) containing at least 109 cells/ml was added to the sterilized ASWNB medium. The growth of the microorganism was monitored, by measuring the cell count using a Petroff-Hauser counter employing phase-contrast microscopy. The sodium chloride concentration was kept at 0.25M, which was arrived at by testing at different salt concentrations. Growing cells as well as cell free growth supernatant, containing metabolites produced during growth, were used as bioleaching agents. 2.3 Methods 2.3.1 Chemical leaching experiments Chemical leaching experiments were carried out in 250-ml Erlenmeyer flasks on an incubated rotary shaker at 200 rpm at 30°C. In all the cases, 1 g of properly crushed nodule of 50-75 µm size fraction was used and the solid: liquid ratio was kept at 1:100(W/V). All of these experiments were conducted to generate baseline data to compare with bioleaching results. In order to optimize different parameters, the duration of leaching was fixed at four hours in some cases. Parameters such as choice of mineral acids, time of leaching, requirement of reducing agents, and the effects of organic and inorganic additives were investigated. HCl, HNO3 and H2SO4 solutions of 2.5M concentrations were used as leaching agents. H2SO4 solution (pH 2.0) alone or with reducing agents like sodium thiosulfate was employed for leaching also. After leaching, leach liquor was filtered using Whatman 42 filter papers and the collected residue was digested in 1:1 HCl at 60-70°C. The resultant solution, after proper dilutions were made, was analyzed for Cu, Co, Ni, Mn and Fe with an Inductively Coupled Plasma spectrophotometer (ICP). All chemicals used were of reagent grade. 2.3.2 Bioleaching experiments Leaching with growing culture: One g of pre-sterilized, pulverized ocean nodule was placed in 90 ml of sterilized ASWNB media in 250 ml conical flasks; 10% v/v actively growing culture (109cells/ml) of the marine isolate was added as inoculum. Growth flasks were removed from the rotary shaker after appropriate time intervals and the solution analyzed for leached metal content. 28
Bioleaching Applications
Leaching with cell-free growth supernatant: To obtain cell free growth supernatant, a fully-grown culture (after 10 hours of growth) was centrifuged at 10,000 rpm for 15 minutes followed by pressure filtration using Millipore ultra-filtration unit. The absence of any cells in the resultant supernatant was assured by observing under phase contrast microscope. One g of pulverized ocean nodule was added to100 ml of the growth supernatant and solid: liquid ratio was kept at 1:100. To optimize recovery of metals in leaching, pH of the growth supernatant was varied from an acidic to an alkaline range by adding 10N H2SO4or 0.1N NaOH. The duration of leaching was kept constant at four hours. The size fraction of the crushed nodules was in the range of 50 to 75 microns for all tests. Leaching with starch added to the growth supernatant: To observe the effect of starch addition to the growth supernatant, increasing proportions of starch were added to 100 ml of cell free growth supernatant and the solid: liquid ratio was kept constant at 1:100. The duration of leaching was maintained at four hours. The size fraction of the crushed nodule was in the range of 50 to 75 microns. Solution pH in all the cases was within 8-8.5 ranges In all the above tests, leach liquor collected after appropriate time intervals was filtered using Whatman 42 filter papers and the residue was digested in 1:1 HCl at 6070°C. The resultant solution, after proper dilutions were made, was analyzed for Cu, Co, Ni, Mn and Fe by an ICP spectrophotometer. 3.
RESULTS AND DISCUSSIONS
3.1 Characterization of the marine isolate A marine bacterium was isolated from an Indian Ocean nodule sample following the procedure discussed in section 2.2. A preliminary morphological and physiological examination of the marine isolate was carried out. Results are presented in Table 4. The isolated strain was rod shaped, motile and able to reduce Mn (IV) to Mn (II). Table 4. Characterization of the marine isolate Shape
Bacilli
Gram reaction
+
Motility
+
Manganese reduction
+
Catalase activity
+
Oxidase activity
+
Aerobic metabolism
+
3.2 Chemical leaching experiments Baseline leaching tests of the nodules in HCl, HNO3 and H2SO4 of 2.5M concentrations were carried out and the results are summarized in Table 5 below. We can observe from the table that, though Cu recovery is around 80% in all the cases and Ni recovery is about 60%, Co and Mn recoveries are very low. When leaching with HCl, only 30% Co leaches is leached. Co recovery is negligible for the other two acids. Low Co recovery in all the above tests may be attributed to low Mn recovery. A major part of Co in nodules is supposedly occluded in the MnO2 matrix, so disintegration of the matrix is an essential prerequisite for Co solubilization. Therefore presence of a reducing 29
Bioleaching Applications
agent, like glucose, under acidic condition with HCl remarkably enhances the recovery of Mn and Co. With glucose in a 2.5M HCl solution 50% Co is extracted while Mn recovery increased to 80%. Again the disparity in Co and Mn recovery proves that Mn and Co recovery do not follow a simple 1:1 correlation. It is possible, however, that all of the cobalt in the nodules might not be directly associated with Mn. Table 5. Leaching of ocean nodules by mineral acids with and without reducing agents Leaching reagent
% Recovery of metals Co
Cu
Ni
Mn
Fe
2.5 M HCl
30
80
55
30
60
2.5M HCl + 20% glucose
50
85
85
80
65
2.5M HNO3
<5
50
50
<5
<5
2.5M H2SO4
<5
80
55
<5
60
pH 2 H2SO4
25
40
26
10
40
pH 2 H2SO4 + 1% thiosulfate
35
45
40
20
50
Since 2.5 M acid solution is not well suited for industrial applications due to corrosiveness and handling hazards, dilute solutions of acid having a pH of 2.0 were applied. From Table 5 it can be observed that 25% Co, 25% Ni and 40% Cu are recovered with almost 10% Mn and 40% Fe extraction. With 1% sodium thiosulfate in solution with pH 2 H2SO4, metal recoveries are observed to increase by 5-10%. In this case it can be observed that adding a reducing agent in the leaching medium enhances the recovery. 3.3 Bioleaching experiments 3.3.1 Leaching by growing cells and cell-free growth supernatant Leaching tests with uninoculated ASWNB at near neutral pH were carried out to verify whether media constituents could dissolve metal values from nodules. Though 10% Cu came out in solution, recovery of Co, Ni, Fe and Mn was negligible. Cu recovery may be attributed to complexation by media constituents. Then leaching of the nodules was carried out using the growing culture of the marine isolate as well as the cell-free growth supernatant at near neutral pH. Results are shown in Table 6. It was observed from the kinetic studies carried out in the lab that within four hours most metal recoveries stabilized; therefore, leach times were maintained at four hours. Comparing leaching by the growing cells and the supernatant, it can be observed that, although there is a small increase in recovery of Co and Fe in case of leaching by the supernatant, recovery of Cu, Ni, Mn are similar in both the cases. During leaching by the growing culture, attachment of the cells onto the nodule surface might have resulted in a decrease in area available for reaction, thereby decreasing the recovery of Co and Fe. Another important observation is, in spite of no significant Mn recovery, sufficient Co came into solution in both test cases. Therefore, Co recovery in case of biological leaching may not be directly related to Mn recovery. It has already been shown in section 3.2 that all the Co in nodules may not be directly associated with the MnO2 matrix. It should be stressed that in both the test cases the pH of the medium was 8.1-8.5. The thermodynamic solubility of all the metals concerned is negligible at that pH range. But the solubility of the transition metals in the complexed state can be markedly different from that in the 30
Bioleaching Applications
uncoordinated state. Therefore complexation at near neutral pH by the metabolites present in the growth supernatant of the marine isolate may be responsible for the unusual solubility of the transition metals. Comparing the results of leaching by a chemical reagent like 2.5M HCl and bioleaching by the growth supernatant of the marine isolate, some interesting conclusions can be drawn. The recoveries of Co, Mn and Fe are quite comparable in both the cases. So in spite of operating at a near neutral pH, metal recoveries similar to that in highly acidic conditions are achieved. This can have significant industrial application with respect to recovering metals from ocean nodules. Table 6. Leaching of ocean nodules by the growing cells, growth supernatant of the marine isolate and 2.5M HCl solution Leaching conditions
% Recovery of metals Cu
Co
Ni
Mn
Fe
Leaching by the growing cells
20
30
20
25
40
Leaching in growth supernatant
25
40
25
25
50
2.5 M HCl
80
30
55
30
60
3.3.2 Leaching by the growth supernatant at different pH values Leaching experiments using the growth supernatant at different pH values were carried out to observe the effect of pH on metal solubilization by the supernatant. The results are shown in Table 7. Until the pH increases above 10 the leaching recoveries of metals do not change much. But at the highly alkaline pH above pH 12.0 the recoveries of all the metals are considerably enhanced. This indicates a change in chemical nature of the metabolite/metabolites present in the growth supernatant of the marine isolate, intensifying the complexation effect in the leach solution. A titration of the growth supernatant against 0.1M NaOH showed that the pKa value lies between 11.5 and 12.5, confirming deprotonation of the metabolites above that pH. Table 7. Effect of pH of the medium on leaching of ocean nodules by the growth supernatant of the marine isolate Leaching conditions (pH of the growth supernatant)
Co
Cu
Ni
Mn
Fe
2.9
53
25
23
10
40
6.0
50
25
26
25
50
8.1
45
20
22
24
45
10.0
56
27
30
22
48
14.0
75
50
45
60
75
% Recovery of metals
31
Bioleaching Applications
3.3.3 Leaching by the growth supernatant at different pulp densities From Table 8 it is evident that increasing the in pulp density (g. of solid/100ml of liquid) up to 10% has little effect on recovery of metal values from the nodule sample. Solution pH in all the cases was within a range of 8.0-8.5. Table 8. Effect of pulp density on leaching of ocean nodules by the growth supernatant of the marine isolate % Recovery of metals
Leaching conditions (pulp density)
Co
Cu
Ni
Mn
Fe
1%
45
20
25
25
45
5%
40
20
25
24
43
10%
45
20
25
10
43
3.3.4 Leaching by the growth supernatant with addition of starch Starch was added to the growth- supernatant in increasing proportions: 1%, 3% and 5%. Addition of starch in the leaching media enhanced Cu and Ni extraction significantly while Co, Mn and Fe extraction also improved. Starch provides a reducing environment, which helps to break down the MnO2 matrix thus liberating Cu, Co and Ni occluded therein. Table 9 shows the enhancement in recovery due to reducing action of starch. In all the cases pulp density was kept at 1% and starting pH was within a range of 8.0-8.5. Table 9. Leaching of ocean nodules by growth supernatant of the marine isolate with addition of starch Conditions of leaching
% Recovery of metals Co
Cu
Ni
Mn
Fe
Growth - supernatant only (marine isolate)
45
20
29
25
45
Supernatant + 0.1% starch
65
35
35
30
50
Supernatant + 0.5% starch
85
70
60
83
80
Supernatant +1% starch
80
80
65
85
86
Supernatant +3% starch
80
83
82
83
85
Supernatant +5% starch
83
85
86
87
84
4. CONCLUSIONS The following major conclusions can be drawn based on the above study: 1) Marine organism isolated from the Indian Ocean nodules grows well in artificial seawater and near neutral pH. It can reduce MnO2. 2) The growing cells as well as the cell-free growth supernatant can leach Cu, Ni and Co from the nodules. 3) Significant dissolution of Co and Fe, comparable to chemical leaching by 2.5M HCl, can be achieved by leaching with the growth supernatant of the marine isolate at near neutral pH. 4) Metabolites produced during the growth of the isolate solubilize transition metals at neutral pH by complexation. 32
Bioleaching Applications
5) The recovery of metals by the growth supernatant remains unaffected by increasing the pulp density to 10%. 6) Recoveries of Cu, Co and Ni are enhanced in highly alkaline growth supernatant having a pH over 12.0. 7) A considerable enhancement in recovery of Cu, Co and Ni is achieved by introducing a reducing agent like starch into the medium. ACKNOWLEDGEMENTS The authors are grateful to Department of Science and Technology, Government of India for providing necessary financial assistance for carrying out the above work. REFERENCES 1. Ehrlich, H.L., 1963. Applied Microbiology 16,197-202. 2. Fuerstenau, D.W., Han, K.N., 1983. Mineral Processing and Technology Review 1, 183. 3. Jana, R.K., Pandey, B.D., Premchand, 1999. Hydrometallurgy 53, 45-56 4. Kanunugo, S.B., Jena, P.K., 1988. Hydrometallurgy 21, 41-58. 5. Konishi, Y., Asai, S., Sawada, Y., 1997. Metalll. Mater. Trans. 28B, 25-32. 6. Kumari, A., Natarajan, K.A., 2001. Hydrometallurgy 62, 125-134. 7. Niinae, M., Komatsu, N., Nakahiro, Y., Wakamatsu, T., Shibata, J., 1996, Hydrometallurgy 40, 111-121. 8. Trifoni, M., Toro, L., Veglio, F., 2001. Hydrometallurgy 59, 1-14. 9. Zhang, Y., Liu, Q., Sun, C., 2001. Minerals Engineering 14, 525-537.
33
15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
A novel biotechnological process for germanium recovery from brown coal Yang Xianwana, Zhu Yuna, Guo Yuxiaa and Guo Banghuia a
Kunming University of Science and Technology, Yunnan, 650093, P.R. China (Tel: 011-86871-5182554)
Abstract A novel biotechnological process, including two main steps - bioleaching and desorption, to recover Ge from brown coal has been developed. The kinetics of the leaching has also been studied. A mathematical model of the leaching process has been deduced. The reaction order of the process has been calculated as 0.826, and the activation energy of the reaction as 63.641kJ/mol. Keywords: biotechnology, Ge, brown coal 1.
INTRODUCTION Coal, especially brown coal from certain coal mines such as Kaiyuan Coal Mine in Yunnan province and Shenhua Coal Mine in Inner Mongolia, contains a certain amount of rare metals, such as Ge and Ga. The conventional process to recover Ge from brown coal takes five steps, i.e. 1) burning of the brown coal; 2) recovering of Ge from ashes by sulfuric acid leaching; 3) precipitating of Ge with tannin; 4) roasting of Ge-containing tannin to produce Ge concentrate with the grade of 11%. This process is complex and has a low recovery of 60%[1], which is sure to bring about a great waste of resource. In current study, a novel process to recover Ge from brown coal in the presence of microorganism has been developed. A Ge recovery of up to 85.33% has been achieved, and kinetic study on the leaching process has also been conducted. 2.
PRINCIPLE It is concluded[2] that 97.3% of Ge in the brown coal exists in the humus, which is the most common component in the brown coal. The humus in brown coal is macromolecule organic substance with a relatively high molecular mass around 300~30,000. According to its solubility in the acid or alkali solution, the humus is classified as three parts: humic acid, fulvous acid, and humin. When humus is treated with an alkali solution, humic acid and fulvous acid can dissolve in the solution, while humin remains insoluble. Neutralizing the solution, humic acid will precipitate while fulvous acid remains in the solution. Research on some brown coal samples shows that 86~89% of Ge in the brown coal is bound to humin, 10~12% to humic acid and 1~2% to fulvous acid, which is why Ge in brown coal can not be recovered by acid leaching or alkali leaching. In order to liberate Ge bound to humus, therefore, Ge-bearing organic complex should be degraded first. 35
Bioleaching Applications
The Ge leaching from brown coal in the presence of microorganism mainly takes place in two stages: 1) biodegradation of brown coal for breaking-down of Ge-containing complex of humus; 2) the recovery of Ge from Ge-containing solution. The flowsheet of this novel process is shown in Fig. 1. The microbes adopted in the experiments for brown coal degradation were selectively cultured from the brown coal. They are classified into three types: bacteria, actinomycetes, and mould, all of which are Gram-positive. There are two types of bacteria known as spheroplast and corynebacterium. It is observed that at the beginning of the leaching, actinomycetes’ function in neutral solution remains important, and then mould and bacteria play an important role in the acid solution. The natural population of the microbes in natural brown coal is very small, thus, a solution with high microbes population (which is referred to as microbe-rich solution) was prepared and used for the degradation-bioleaching. The brown coal samples were firstly leached in the microbes-rich solution for several days; most Ge was found to be leached into solution. However, some of Ge was found being adsorbed in the pores of coal. Thus the second step, desorption was adopted to extract Ge trapped in the pores from coal. The average original aperture of brown coal r0 was 1.13×10-4cm (11.3µm), while the length of microorganism lm is 1.6µm. ro > 2lm. Therefore the leaching process can be considered as a reaction of an aqueous species A with a porous material, the product of the reaction is also an aqueous species. The leaching rate of such process can be presented on the basis of "Pore Model"[3] and Petersen Model[4]: ε dCGe 1 = k 0 (1 − ) CnA N ϕ dt r0 G −1 (1)
where, k is logic rate constant of the bioleaching reaction; r0-average original aperture of raw material; ε0 is original porosity of raw material; CA is concentration of leaching agents; φ is reproducing rate of microorganism (=1-[C΄(t/τ)+C΄΄(t/τ)2], where C΄ and C΄΄ are constants); N is the population of microorganism; t is leaching time; G is generation of organism. Brown coal (Ge: 200g/t) Dilute sulfuric acid solution Crush
Microbes colony
Bioleaching Desorption High Ge concentration resolution ([Ge]T>800mg/L) Figure 1. Flowsheet of the recovery of Ge from brown coal Integration of Equation (1) gives: α=
36
εο bkCnA Ν ο τ × × (t + Bt 2 + Ct 3 ) 1 − εο ro ρ G
(2)
Bioleaching Applications
where: α is recovery of Ge; B and C are constants; ρ is the density of raw material; n is reaction order; N0 is the original population of the bacteria; τ is the time of thorough reaction. ε bkCnA Ν ο τ Assuming: Α = ο × × (3) 1 − εο roρ G The logarithm of Equation (3) is given by: lg A = lg k + n lg CA + lg
εο bΝ τ × ο 1 − ε ο roρG
(4)
Thus, from Equation (4), the logic rate constant k and reaction order n of the leaching reaction can be determined. 3.
EXPERIMENTAL The brown coal samples used for the experiments were crushed into three size fractions, say, 0.0998~0.147mm, 2~4mm and 12~14mm. The composition of the brown coal is 56.02%C, 5.21%O, 2.17%H, 1.34%S, 29.33% ash, and 0.0312% Ge. Its combustion value is 16,798 kJ/kg. The porosity of brown coal was 42.25% that was determined by immerging samples into water for 4 hours and measuring the water in the samples. The density of dry coal was 1.12g/cm3. The leaching experiments were carried out in a static beaker. The leaching solution is microbe-rich solution with one of leaching agents in it, which are industrial pure sodium hydroxide, sulfuric acid, and ammonia sulfate, respectively. The pH varied from 3.5 to 9.0, the operation temperature varied from 22 to 55°. To assure that the concentration of leaching agents remained constant, the ratio of liquid to solid (mL/mS) was controlled as high as 10. During the leaching, the leaching solution was sampled and Ge concentration was analyzed once every day. 4.
RESULTS AND DISCUSSION Following findings and evidences show that microbes has played an essential role in the leaching process: 1. Gas bulbs were observed to be produced on the surface of the coal, and pH of the solution decreased, as shown in Table 1. 2. Oxygen concentration in the coal decreased, as shown in Table 1. 3. In the presence of the microbes, Ge concentration in the solution increased continuously, as shown in Table 1, while in the absence of the microbes, no Ge has been leached; 4. The population of the microbes was observed to increase in the solution. Table 1. Data indicating the degradation of the Ge-bounded humus complex in the presence of the microbes Leaching time /d
1
2
3
4
5
6
7
8
pH of the solution
7
6.5
6.5
6.0
5.0
4.5
4.0
3.0
Oxygen concentration in the coal/ %
5.21
5.02
4.67
4.12
3.68
3.13
2.82
2.63
Ge concentration in the solution in the presence of microbes /mg/L
0
1
3
5
8
10
11
13 37
Bioleaching Applications
The influences of coal particle size, leaching agent and its concentration, temperature are discussed in details below. In addition, the mathematical model is also developed. 4.1 Effect of coal particle size on the leaching rate of Ge Coal samples of three size samples were leached in microbe-rich solution with 0.001mol/L sulfuric acid in it: 8~12mm, 0.533~2mm, 0.074~0.998mm. The results were shown in Fig. 2. It can be seen from Fig. 2 that the recovery of Ge increases with the decrease of the coal particle size, and the optimal size is 2mm. 70
1
60
1
40
Ge recovery (α) %
Ge recovery (α) %
50
2 30 20
3 10
50
2
40 30
3
20 10 0
0 0
2
4
6
8
10
0
12
5
10
time, day
15
20
25
time, day
Figure 2. Effect of coal size on Ge recovery (leaching agent as sulfuric acid)
Figure 3. Effect of leaching agents on Ge recovery
1- 0.074~0.998mm 2- 0.533~2mm 3- 8~12mm
1- H2SO4, 0.001mol/L 2- NaOH, 0.001mol/L 3- (NH4)2SO4, 0.001mol/L
4.2 Effect of leaching agents on the recovery of Ge Three agents, sulfuric acid, sodium hydroxide and ammonia sulfate, were used for leaching and all the experiments were conducted in the presence of microorganism. The mass ratio of liquid to solid, e.g. mL/mS, remains 10 in all experiments. The pH of the leaching solution varied from 2.5 to 10.0. The experiments lasted for 26d. The Ge recovery vs. leaching time in the presence of microorganism with different leaching agents is shown in Fig. 3. It is obvious that H2SO4 solution is a preferable leaching agent to recover Ge in the presence of microorganism. 1
4
70
0,8
60 50
3
40
lgA
Ge recovery (α) %
80
2
30 20 0 0
4
8
12
time,day
16
20
Figure 4. Effect of sulfuric acid concentration on leaching rate 1- 0.0004 mol/L; 2- 0.0012 mol/L; 3- 0.002 mol/L; 4- 0.004 mol/L; 38
0,4 0,2
1
10
0,6
24
0 -3,5
-3,1
lgC
-2,7
Figure 5. lgA-lgC diagram in the presence of H2SO4
-2,3
Bioleaching Applications
4.3 Effect of leaching agent concentration The result of experiments under different acid concentration is shown in Fig. 4. It can be seen, from the Fig. 4, that the higher the concentration of sulfuric acid is, the higher the recovery of Ge is. But the concentration of agent cannot be increased too high because the pH of solution has to be controlled in the range that is favorable for microbes. The regression of experimental data is summarized as following: α1=1.2024t - 0.0168t2 α2=2.4769t - 0.0569t2 α3=4.5083t - 0.1156t2 (5) 2 α4=7.8766t - 0.2115t Supposing A is the quotieties of the first order of the equation (2) and cH2SO4 is the concentration of sulfuric acid, the dependence of lgA as a function of lgcH2SO4 can be described as a linear function as shown in the Fig. 5, from which the reaction rate constant of the leaching K and the reaction order n is determined to be 9.113×10-4/d and 0.826, respectively. The reaction order, n=0.826, indicates that the leaching is a slow process. 4.4 Effect of temperature The experimental results under different temperatures in the presence of microbes are shown in Fig. 6. The regression of these results are presented in Equation (6): α1=0.7684t - 0.00956t2 α2=1.9517t - 0.0339t2 α3=4.486t - 0.11456t2 (6) 2 α4=5.7726t - 0.1348t - 2. 4
70
- 2. 6
60
4
- 2. 8 3
40
lgk
α, %
50 30
- 3. 2
2
20
-3
- 3. 4
1
10
- 3. 6
0 0
4
8
12
16
20
24
t, day
Figure 6. Effect of temperature on reaction rate
- 3. 8 0. 0031
0. 0032
0. 0033 1/ T, K
0. 0034
0. 0035
Figure 7. lgK-1/T plot of the leaching process
T/°C: 1-20°C; 2-30°C; 3-40°C; 4-46°C
1 is shown in Fig. 7. The slope of the line is –3323.79. T From the slope value, the activation energy of the reaction W is determined to be 63.641 KJ/mol, which further indicates that the process is controlled by chemical reaction. A recovery of 72% was achieved for experiments carried out with 0.002mol/L sulfuric acid solution as the leaching agent in 16d at 40~42°. The recovery is not high because, after leaching, the coal residue was porous and some Ge would be trapped in the The Arrhenius plot of lg K −
39
Bioleaching Applications
coal residue. Therefore, the desorption step after bioleaching was adopted for recovering trapped Ge and total Ge recovery of 85.33% was obtained. 5.
CONCLUSIONS 1. A novel biotechnological process including two main steps, bioleaching and desorption, has been developed to recover Ge from brown coal of some coal mines. 2. The coal particle size, leaching agents and their concentrations, and temperature are important factors affecting the leaching rate of Ge. The optimal operation condition had been determined as following: leaching with 0.002 mol/L sulfuric acid solution at 40~42° for 16 days; the coal particle size is 2mm. The recovery of Ge could be up to 72%, the desorption after bioleaching can further increase the recovery of Ge up to 85.33%; 3. A mathematical model of the leaching process has been deduced, which gives a good fit to both the calculated and experimental data; 4. The reaction order of the process has been calculated as 0.826, and the activation energy of the reaction as 63,641 kJ/mol.
REFERENCES
1. Tselikiman A.H., Metallurgy of Rare Metals (in Russian) [M], Metallurgy Industry Press, Beijing, 1980:344. 2. Wu X.L., Metallurgy of Ge (in Chinese) [M], Metallurgy Industry Press, Beijing, 1987:24. 3. Yang X.W., Hydrometallurgy [M]. Metallurgy Industry Press, Beijing, 1998:329. 4. Sohn H.Y., Rate and Process of Extractive Metallurgy [M], Metallurgy Industry Press, Beijing, 1984:24.
40
15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Aerobic and anaerobic bacterial leaching of manganese J.G. Zafiratos and S. Agatzini-Leonardou Laborarory of Metallurgy, National Technical University of Athens 15780 Zografos, Athens, Greece Abstract Heterotrophic bacteria are widely used in the study of bacterial leaching of manganese from manganese dioxide ores and glucose or other organic compounds are used as a source of energy, rendering their commercial utilization uneconomic. In the present research, autotrophic bacteria and a mineral source, such as elemental sulphur as their growth substrate, were used instead in order to develop an economic bacterial leaching process for low-grade manganese ores. Active cultures of bacteria were isolated from mining areas such as Laurium, in Attica (Greece), Kassandra Mines, in Chalkidiki (Greece), Skouriotissa, in Cyprus and Wales (UK). Low-grade manganese dioxide ore samples were obtained from ELBAUMIN S.A., a company which operates a manganese mine in Drama (Greece). Recovery of manganese from these samples by bacterial leaching was studied in stirred tank reactors at both aerobic and anaerobic conditions. A process flow sheet was constructed for the production of electrolytic manganese dioxide from low-grade ores in Greece and a preliminary economic evaluation of this conceptual project was performed. Keywords: manganese, bioleaching, Thiobacillus, low-grade ores 1.
INTRODUCTION Manganese is a strategic metal. It is essential in the production of iron and steel and no adequate substitute has been found to date. Its non-steelmaking applications, mainly dry-cell batteries, are rapidly growing as well [1]. Manganese ore, however, is mined in large quantities only in a few countries of the world (U.S.S.R., S. Africa), and it is not presently commercially mined in the steel-making countries (U.S.A., Japan, E.U.). This is so because the latter possess low-grade manganese ore reserves, which are not amenable to conventional enrichment and extraction techniques. The above facts have imposed the need for the development of novel manganese extraction techniques which would eventually solve the supply problems for the industrial countries, which are currently importing enormous quantities of manganese ore or ferromanganese. The hydrometallurgical treatment of manganese ores has a great potential for future application. Chemical reductive leaching has been tested with a variety of leaching agents with satisfactory results but it creates environmental (SO2.H2O leaching) or iron removal (FeSO4 / H2SO4 leaching) problems [2, 3]. 41
Bioleaching Applications
On the other hand, the biological aqueous processing of ores has been successfully applied, on an industrial scale, for the extraction of copper, uranium and precious metals from sulphide mineral containing deposits [4, 5]. This large-scale application has proved that bacterial leaching, using autotrophic microorganisms, is a low-cost and environmentally safe process but the sulphide ores are under the course of constant depletion, worldwide. Therefore, the scientific and economic interest for the extraction of metals from non-sulphide resources using microorganisms is ever growing. Until recently, research efforts on manganese bioleaching have been concentrated on heterotrophic bacteria, which directly reduce the manganese dioxide to the soluble Mn2+ form. Marine bacteria and, probably, soil and freshwater species (e.g. Bacillus, Micrococcus, Pseudomonas, Achromobacter, Enterobacter) biodegrade pyrolusite by enzymatic reduction, using a variety of carbohydrate nutrients (e.g. glucose, molasses) under aerobic and microaerobic growth conditions. Also, certain species of fungi (e.g. Aspergillus niger) are known to act on Mn4+ by formation of a mixture of extracellular reducing organic compounds, consisting, mainly, of citric, formic and oxalic acids, during fermentative metabolism of sugars. Complete manganese solubilization can be achieved by heterotrophic microorganisms, but the cost of nutrients such as glucose, and other refined sugars, makes their application prohibitively expensive [3, 6-16]. At the Laboratory of Metallurgy of the National Technical University of Athens a research project was undertaken with the aim of investigating the possibility of applying biological leaching to the Greek low-grade manganese ores, which are similar to many pyrolusitic deposits around the world, using autotrophic bacteria [17, 18]. The tasks of this project included: a) Microorganism strain selection between mixed cultures of autotrophic bacteria (Thiobacillus ferrooxidans, Thiobacillus thiooxidans and Leptospirillum ferrooxidans) of different origins, grown on elemental sulphur. b) Manganese bioleaching using the selected bacterial strain. c) Flowsheet design and preliminary economic evaluation of the conceptual project. Task (b) results have already been presented elsewhere [18]. In the present work tasks (a) and (c) are presented. 2.
MATERIALS AND METHODS
2.1 Microorganisms and culture media The mixed culture of T. ferrooxidans like strains was selected among four different strains of autotrophic bacteria: a) a strain (named Culture C) isolated from a bacterial heap leaching solution, provided by Hellenic Mining Company Ltd. (Cyprus), b) a strain (named Culture K) isolated from acid mine drainage of the Kassandra mixed sulfide mine (Greece), c) a strain (named Culture L) isolated from a sulfide waste dump in Lavrion (Greece) and d) a strain (named Culture W) provided by Professor A. Kontopoulos (sent by Dr. F. D. Pooley, University of Wales, UK).
42
Bioleaching Applications
Initially, the microorganisms were cultivated in an iron-free 0K standard nutrient medium. Later, the nutrient mediums used were ID and MS0b [19], both supplemented with elemental sulphur (0.5 or 1% w/v S). 2.2 Ore and elemental sulphur The low-grade manganese dioxide ore was provided by ELBAUMIN S.A. mining company (Athens, Greece) and came from a battery-grade ore beneficiation site in Drama, Greece. Its chemical analysis is given in Table 1. The X-Ray diffraction analysis of the ore showed that its mineralogical components were: pyrolusite, quartz and calcite. The elemental sulphur was obtained from an oil refinery in Northern Greece and was ground to –60 and –100 mesh. Table 1. Chemical analysis of the ore sample M2W (average of two analyses performed on two specimens of the ore sample) Constituent
Dry ore composition (wt %)
SiO2
25.05
MnO2
30.65
MnO
3.30
MnTOTAL
21.92
Pb
0.12
Fe
1.07
Zn
0.10
Cu
0.009
Insolubles (in HCl)
28.27
CaO (CaTOTAL)
3.80
CaCO3
-
CO2
-
2.3 Culture optimization, strain selection and adaptation procedures The adaptation of the mixed culture to use elemental sulphur as its sole energy source was done in shake flasks by serial subculture in a sulphur-containing medium. The adaptation of the bacteria to grow on elemental sulphur was carried out in 500ml shake flasks, incubated in a thermostatic water-bath (30oC) with magnetic stirring. The phase of adaptation of the culture, in the presence of manganese, was carried out in air agitated Pachuca-type reactors. 3.
RESULTS AND DISCUSSION
3.1 Factorial Experiment of Strain Selection A 23 full factorial experiment was conducted using the levels of factors shown in Table 2.
43
Bioleaching Applications
Table 2. Factorial design for strain selection. Minimum and maximum levels of variables FACTORS
VARIABLES
LOW LEVEL
HIGH LEVEL
A
Sulphur Pulp Density
0.5% w/v
1.0% w/v
B
Bacterial Cell concentration in the inoculum
107cells ml-1
109 cells ml-1
C
Nutrient Medium
MS0b
ID
D
Sulphur Grain Size
-60mesh
-100mesh
The responses investigated were the sulphur conversion (oxidation) factor, the production of biomass and the production of sulphurous ions. Biomass growth was monitored by optical density measurements of the culture, correlated to actual bacterial mass using the Micro-Kjeldahl technique. The conversion of elemental sulphur to sulfate was monitored by sulfate analysis of the culture medium using a turbidimetric technique after precipitation of barium sulphate. Sulphurous acid concentrations were determined using colorimetry. Parameters, with constant values throughout the experiments, were the leach solution initial pH (1.5) and the temperature (30oC). The experimental results are shown in Table 3. Statistical analysis of the data is depicted in Figures 1-6. Due to space limitations, the results for the Kassandra culture are the only given here. Besides, all four cultures exhibited the same behaviour in relation to the factors studied. Regarding the response "sulphur to suphate conversion factor", the main effect of factor A as well as the interactions AB and ABC were found statistically significant. The existence of significant interaction, means that the model describing this response is not linear, i.e. the experimental region corresponds to the curvature of the response curve (near a maximum or minimum). More specifically, the sulphur pulp density had a negative effect to the sulphur oxidation, which was attributed to the inhibition of bacterial activity by solid particles, commonly observed in bacterial leaching tests. The interaction AB was significant and negative. This means that the effect of pulp density on the sulphur conversion factor is much greater when a low initial biomass concentration is used, whithin the present factor limits. Regarding the response "sulphurous acid production", the main effects of factors A and B as well as the interactions AB, BC and ABC were found statistically significant. The existence of significant interactions, again, means that the model is not linear. Specifically, the sulphur pulp density again had a negative effect to the sulphurous acid production. On the contrary, the initial biomass concentration had a positive effect on this response. The interaction of factors A, B was again significant and negative. This means that the effect of pulp density on the production of sulphurous acid is much greater when a low initial biomass concentration is used, whithin the present factor limits. The interaction BC was significant and positive. This means that the effect of the initial biomass concentration on the production of sulphurous acid is much greater when the ID culture medium is used than with the MS0b medium. As far as the response "biomass production" is concerned, only the main effect of factor A and the interaction AB were found statistically significant. Specifically, the sulphur pulp density again had a negative effect to biomass production. The interaction AB was again significant and negative. This means that the effect of pulp density on the 44
Table 3. Strain Selection. Responses: Sulphur Conversion Factor, Sulphurous Acid and Biomass Production. TREATMENT CODE
Sulphurous Acid Production (mg/l)
Sulphur Conversion Factor
Biomass Production (cells ml-1 x108)
I
II
Average
I
II
Average
I
II
Average
(1)
0.55
0.44
0.49±0.05
14.97
23.35
19.16±3.06
45.54
70.02
57.78±9.82
a
0.20
0.29
0.25±0.05
10.78
6.59
8.68±3.06
34.74
40.32
37.53±9.82
b
0.41
0.48
0.45±0.05
13.17
12.57
12.87±3.06
44.82
70.02
57.42±9.82
ab
0.23
0.19
0.21±0.05
9.58
2.99
6.29±3.06
14.04
32.40
23.22±9.82
c
0.39
0.29
0.34±0.05
11.37
8.38
9.88±3.06
38.52
63.90
51.21±9.82
ac
0.37
0.45
0.41±0.05
11.37
10.18
10.76±3.06
41.76
65.70
53.73±9.82
bc
0.51
0.61
0.56±0.05
19.76
30.53
25.14±3.06
60.48
81.00
70.74±9.82
abc
0.20
0.11
0.16±0.05
8.38
1.20
4.79±3.06
9.00
12.40
10.70±9.82
d
0.57
0.38
0.47±0.05
14.97
20.35
17.66±3.06
45.18
71.10
58.14±9.82
ad
0.34
0.14
0.24±0.05
10.78
4.19
7.48±3.06
26.10
32.40
29.25±9.82
bd
0.55
0.45
0.50±0.05
17.96
26.34
22.15±3.06
50.76
71.10
60.93±9.82
abd
0.24
0.22
0.23±0.05
10.18
3.59
6.88±3.06
19.98
40.32
30.15±9.82
cd
0.48
0.34
0.41±0.05
11.97
10.78
11.37±3.06
44.46
65.70
55.08±9.82
acd
0.24
0.39
0.31±0.05
10.78
7.18
8.98±3.06
38.52
62.46
50.49±9.82
bcd
0.84
0.84
0.84±0.05
25.14
31.13
28.14±3.06
62.10
81.00
71.55±9.82
abcd
0.18
0.19
0.18±0.05
8.98
1.80
5.39±3.06
10.80
12.40
11.60±9.82
Bioleaching Applications
biomass production is much greater when a low initial biomass concentration is used, whithin the present factor limits. Based on the above results, the suggested models for each of the responses considered are: YR(sulphur conversion factor) = 0.38 - 0.13 * X1 - 0.066 * X1* X2- 0.063 * X1 * X2 * X3 YR(sulphurous acid production) mg/l = 12.85 - 5.44 *X1 + 1.10 *X2 - 2.68 *X1*X2 + 1.70 *X2*X3 - 2.53 *X1*X2*X3 YR(biomass production) *109 cells/ml = 4.560 - 1.476 * X1 - 0.8360 * X1 * X2 where: Y’s are the predicted values of the responses, X1, X2, X3 are the coded variables corresponding to the natural variables A, B, C. The coded variable is equal to +1 or –1 when the corresponding natural variable is at its high or low level, respectively. The maximum sulphur conversion factors obtained during these tests by each culture were the following: C: 0.999, K: 0.871, L: 0.970 and W: 0.622. As can be readily seen, strain C, effected a nearly complete sulphur conversion. For this reason, it was selected for further experimental work. DESIGN-EXPERT Plot
Sulphur conversion A: Pulp Density B: Initial Biomass C: Nutrient Medium D: Sulphur Particle Size
99
A Half-Normal Probability (%)
97 95
AB ABC
90
AD BD ACD C BC D BCD CD B
85 80 70 60
ABD AC
40 20
ABCD
0
0.00
0.06
0.13
0.19
0.26
|Effect|
Figure 1. Half-Normal plot of effects on sulphur conversion factor
46
Bioleaching Applications DESIGN-EXPERT Plot
Sulphurous Acid Production A: Pulp Density B: Initial Biomass C: Nutrient Medium D: Sulphur Particle Size
99
A Half-Normal Probability ( % )
97 95
AB ABC
90
BC
85
B BD AD
80 70 60
D ABCD BCD ABD
40 20 0
CD C ACD AC
0.00
2.72
5.44
8.17
10.89
|Effect|
Figure 2. Half-Normal plot of effects on sulphurous acid production DESIGN-EXPERT Plot
Biomass Production A: Pulp Density B: Initial Biomass C: Nutrient Medium D: Sulphur Particle Size
99
A Half-Normal Probability ( % )
97 95
AB ABC
90 85
B
80 70
BC
60 40 20 0
C BD ABD BCD AD AC ABCD D ACD CD
0
738062500 1476125000 2214187500 2952250000
|Effect|
Figure 3. Half-Normal plot of effects on biomass production
47
Bioleaching Applications
0.70 Sulphur conversion factor
0.57 0.43 0.30 0.17
1.00E+09 1.00 6.70E+08
0.88 0.75
3.40E+08 Β: Initial Biomass (cells/ml) 10.00E+06
0.63 Sulphur pulp density (% w/v) 0.50
Figure 4. 3-D Response surface of sulphur conversion factor
Biomass Production (cells/ml)
7.11E+09
1.00E+09 1.00 0.88 5.05E+08
0.75
Initial Biomass (cells/ml) 10.00E+06
0.50
0.63 Sulphur Pulp Density (% w/v)
Figure 5. 3-D Response surface of biomass production
48
Bioleaching Applications
Sulphurous acid production (mg/l)
26.64 21.25 15.86 10.48 5.09
1.00E+09 1.00 6.70E+08
0.88 0.75
3.40E+08 Initial biomass (cells/ml) 10.00E+06
0.63 Sulphur pulp density (% w/v) 0.50
Figure 6. 3-D Response surface of sulphurous acid production 3.3 Bioleaching at Aerobic and Anaerobic Conditions At aerobic conditions, the bacterial population derives energy from the oxidation of sulphur to SO42-, according to the oxidation pathway below [20]:
(1) where, APS: adenosine phosphosulfate, [S]: colloidal sulphur It is believed that the main intermediate product SO32- subsequently reduces MnO2 during leaching, according to the reaction [21-24]: (2) MnO2 + SO32- + 2H+ → Mn2+ + SO42- + H2O At anaerobic conditions, i.e. in the absence of oxygen, according to literature, T. ferrooxidans couples the oxidation of inorganic sulphur to the reduction of Fe3+ to Fe2+ [25-27]. The product of this anaerobic process is beneficial to manganese leaching, because Fe2+ ions readily react with MnO2 in the acidic culture medium, according to the equation: (3) MnO2 + 2Fe2+ + 4H+ → Mn2+ + 2Fe3+ + 2H2O Therefore, ferric ions were added to the culture medium at a concentration of 1g/l, thus producing culture medium MS1b. 3.4 Factorial Experiment of Bioleaching A 23 full factorial experiment was conducted using the levels of factors shown in Table 4.
49
Bioleaching Applications
Table 4. Factorial Design of the Bioleaching Experiment - Minimum and Maximum Levels of Variables FACTORS
VARIABLES
A
Ore pulp density
B
“Ore / elemental sulphur” (O/S) mass ratio 3+
C
Fe concentration in the leach solution
D
Gas composition
LOW LEVEL
HIGH LEVEL
5% w/v
15% w/v
8.6/1
8.6/2 3+
0g/l Fe
air + 1% CO2
1g/l Fe3+ N2 + 1% CO2
The bioleaching results have been presented elsewhere [18]. Briefly, the main effects A, C and D and also the interactions AD, CD were statistically significant at α=0.01. In this range of variables, too, the Fe3+ concentration in the leach solution had the largest effect with a wide gap separating it from the remaining contrasts. Regarding the manganese leaching rate, this was found to be dependent on all of the studied factors, particularly on the presence of Fe3+ in the solution, which speeded up the reaction. Therefore, the following conclusions were drawn: - Bacterial leaching of manganese dioxide ores by autotrophic species is possible in stirred tank reactors with dispersion of air plus CO2 or N2 plus CO2. - Manganese extraction is favored at low pulp densities and in the presence of ferric iron at anaerobic conditions. - The mass ratio ore / elemental sulphur does not affect the manganese extraction under the experimental conditions used. - The low values of pH, resulting during leaching, cause the complete dissolution of iron contained in the ore. 4.
FLOWSHEET DESIGN AND ECONOMIC EVALUATION Taking into consideration the existence of an electrolytic manganese dioxide (EMD) plant in Thessaloniki (Northern Greece), a flowsheet was designed which employs ore pretreatment at the mine site (Drama) and bacterial leaching of the washed ore at the electrolysis plant site. A preliminary economic evaluation of a project based on the above flowsheet was conducted. The assumptions made for this purpose were the following: i. Head ore grade: 20% Mn ii. Mining Method: underground iii. Yearly run of mine ore production: 63,000 tons iv. Project lifetime: 20 years v. Overall Mn recovery: 60% vi. Fe dissolution: 65% vii. Yearly EMD production: 12,000 tons The evaluation results are given in Table 5. In order to economically evaluate the project, the Discounted Cash Flow Method was used (DCF). The performance indicator rate (i) of the investment (ROI), determined by solving the equation: Net Present Value = 0, was found to be equal to 0.2437, which is considered very promising.
50
Bioleaching Applications
Table 5. Commercial Bioleaching Plant Costs (in Euros) OPERATING COST CAPITAL COST 56700 t/y 56700 t/y Labour 365,124 Beneficiation Equipment 1,684,226 Supplies 57,051 Leaching Equipment 2,342,480 Maintenance 22,820 Neutralization Equipment 2,546,850 Reagents 524,866 Construction 1,173,001 Energy 718,838 Process Plant Overall 7,746,557 Various 34,230 Design and Supervision 373,294 Electrowinning 2,971,701 Unforseen 604,842 Overall Leach-EW 4,694,629 Working Capital (W) 376,522 Ore Transportation 458,260 TOTAL (Q) 9,101,215 Mining-Beneficiation 1,427,569 Q-W 8,724,693 TOTAL 6,580,458 Remaining Value S=20%Q 1,820,243
35000 t/y 1,261,004 1,753,778 1,906,787 878,148 5,799,717 279,601 452,913 281,814 6,814,045 6,532,231 1,362,809
5.
CONCLUSIONS The results of the strain selection experimental data led to the establishment of the following optimized conditions for growth of the mixed bacterial culture on the elemental sulphur substrate: - Nutrient medium: MS0b (iron-free) - Cell concentration in the inoculum: 109 cells.ml-1 - Sulphur pulp density: 0.5%w/v - Sulphur grain size: -60mesh - Culture selected for further work: Culture C The results of the bioleaching experimental data led to the following optimum factor levels for the process of manganese leaching by the mixed bacterial culture: - Ore pulp density: 5%w/v - O/S mass ratio: 8.6/1 - Fe3+ concentration in the leach solution: 1g.l-1 (Nutrient medium: MS1b) - Dispersed gas composition: N2 + 1% CO2 Based on the above parameters, and the flowsheet design, the preliminary economic evaluation for an electrolytic manganese dioxide project in Greece showed favourable results.
51
Figure 7. Flowsheet of a conceptual manganese beneficiation – bioleaching – electrowinning project in Greece
Bioleaching Applications
REFERENCES
1. S.A. Weiss. Manganese, the Other Uses - A Study of the Non-steelmaking Applications of Manganese. Metal Bulletin Books, England (1978) 1-355. 2. L.Y. Tavares and L.A. Teixeira. “Precipitation of jarosite from manganese sulphate solutions” in: J.E. Dutrizac and A.J. Monhemius (eds): Iron Control in Hydrometallurgy, Ellis Horwood Ltd, England (1986) 431-453. 3. C. Abbruzzese, M.Y. Duarte, B. Paponetti, L. Toro. Biological and chemical processing of low-grade manganese ores. Minerals Eng. 3(3/4) (1990) 307-318. 4. A.E. Torma, K. Bosecker. Bacterial leaching. In: Progress in Industrial Microbiology Vol.16, Bull M.J. (ed), Elsevier, Amsterdam (1982) 77-118. 5. K. Adam, A. Kontopoulos, M. Stefanakis and M. Taxiarchou “The Use of Novel Biooxidation Techniques in the Treatment of Mixed Sulfide Ores” METBA Final testwork report, March 1988 - January 1990 (1990) 74pp. 6. C. Abbruzzese, M.Y. Duarte, A. Marabini, B. Paponetti and L. Toro. “Manganese recovery from MnO2 ores by Aspergillus niger: Role of metabolic intermediate”, in: B.J. Scheiner, S.K. Kawatra and F.M. Doyle (eds): Biotechnology in Minerals and Metal Processing. S.M.E. Inc, Littleton (1989) 33-37. 7. F. Novielli, E.C. Perkins. Bacterial leaching of manganese ores. U.S.B.M. RI 6102 (1962) 1-11. 8. E.G. Baglin, J.A. Eisele, D.L. Lampshire and E.G. Noble “Bioleaching of manganese from ores using heterotrophic microorganisms” in: R.W. Smith and M. Misra (eds): Mineral Bioprocessing. TMS, Warrendale (1991) 233-245. 9. E.G. Baglin, J.A. Eisele, D.L. Lampshire, E.G. Noble. Biological leaching of manganese ores. Paper presented at the SME Annual Meeting, February 25-28, Denver, Colorado (1991). 10. V.D. Remezov, A.D. Agate, V.A. Yurchenko. Leaching and beneficiation of manganese ores. In Karavaiko G.I. et al (eds). Biogeotechnology of Metals - Manual. UNEP/GKNT, Moscow (1988) 304-313. 11. J.M. Kozub and J.C. Madgwick “Microaerobic microbial manganese dioxide leaching” Proc. Australas. Inst. Metall., 288 (1983) 51-54. 12. P.J. Holden and J.C. Madgwick “Mixed culture bacterial leaching of manganese dioxide”, Proc. Australas. Inst. Metall., 286 (1983) 61-63. 13. J.C. Madgwick and C.M. Silverio, “Biodegradation of manganese dioxide by purified leaching microorganisms”, Bull. Proc. Australas. Inst. Metall., 290(7) (1985) 63-66. 14. M.J. Hart, J.C. Madgwick. Biodegradation of manganese dioxide tailings. Bull. Proc. Australas. Inst. Metall. 291(3) (1986) 61-64. 15. T.I. Mercz, J.C. Madgwick. Enhancement of bacterial Manganese leaching by microalgal growth products. Proc. Aus.IMM 283 (1982) 43-46. 16. H. Buys et al. Acid removal of manganous ion adsorbed on ore surfaces during microbial Manganese dioxide leaching. Bull. Proc. Aus.IMM 291(4) (1986) 71-73. 17. J.G. Zafiratos and S. Agatzini, “The use of microorganisms in the leaching of manganese ores” Mineral Wealth, 85 (1993) 1, 5-22. 18. J.G. Zafiratos and S. Agatzini “Continuous bacterial leaching of a low-grade manganese dioxide ore” Paper presented at the TMS Annual Meeting, Orlando, Florida (1997). 19. R.T. Espejo and P. Romero, “Growth of Thiobacillus ferrooxidans on elemental sulphur”, Appl. Env. Microbiol., 53(8) (1987) 1907-1912.
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20. D.G. Lundgren, R. Reed and M. Valkova-Valchanova, “Chemical reactions important in bioleaching and bioaccumulation” in: H.L. Ehrlich and D.S. Holmes (eds), Biotechnology and Bioengineering Symposium No.16. J. Wiley & Sons, New York (1986) 7-22. 21. A.E. Torma “The microbiological extraction of less common metals” J. of Metals, 41(6) (1989) 32-35. 22. J. Ghosh and K. Imai, “Leaching of manganese dioxide by Thiobacillus ferrooxidans growing on elemental sulphur”, J. Ferment. Technol., 63(3) (1985) 259-262. 23. J. Ghosh and K. Imai, “Leaching mechanism of manganese dioxide by Thiobacillus ferrooxidans”, J. Ferment. Technol, 63(3) (1985) 295-298. 24. K. Imai “On the mechanism of bacterial leaching” in: J.A. Brierley, L.E. Murr and A.E. Torma (eds): Metallurgical Applications of Bacterial Leaching and Related Microbiologi-cal Phenomena. Academic Press, New York (1978) 275-282. 25. T. Sugio, et. al., “Role of ferric ion - reducing system in sulphur oxidation of Thiobacillus ferrooxidans” Appl. Env. Microbiol., 49 (1985) 1401-1406. 26. T. Sugio, et. al., “Production of ferrous ions as intermediates during aerobic sulphur oxidation in Thiobacillus ferrooxidans” Agric. & Biol. Chem., 50 (1986) 2755-2761. 27. T. Sugio, et. al., “Purification and some properties of sulphur: ferric ion oxidoreductase from Thiobacillus ferrooxidans” J. Bacteriol., 169 (1987) 4916-4922.
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Anaerobic iron sulfides oxidation Axel Schippers Section Geomicrobiology, Federal Institute for Geosciences and Natural Resources (BGR), Stilleweg 2, D-30655 Hannover, Germany e-mail:
[email protected] Abstract Several experiments using FeS2 and FeS have been done to find out if metal sulfides are oxidized under anaerobic conditions at circumneutral pH. In chemical experiments, FeS2 and FeS were oxidized by MnO2 but not with NO3- or amorphous Fe(III) oxide. With MnO2 as oxidant, elemental sulfur and sulfate were the only products of FeS oxidation, whereas FeS2 was oxidized to a variety of sulphur compounds, mainly sulfate plus intermediates such as thiosulfate, trithionate, tetrathionate, and pentathionate. Thiosulfate was oxidized by MnO2 to tetrathionate while other intermediates were oxidized to sulfate. The reaction products indicate that FeS2 was oxidized via the thiosulfate mechanism and FeS via the polysulfide mechanism under anaerobic conditions which previously had been found for aerobic metal sulfide oxidation. For anaerobic FeS2 oxidation with MnO2 the reaction rates related to the FeS2 surface area were 1.02 and 1.12 nmol.m-2.s-1 for total dissolved S and total dissolved Fe, respectively. These values are in the same range as previously published rates for the oxidation of FeS2 by Fe(III). In presence of MnO2, an Fe(II)/Fe(III) shuttle should transport electrons between the surfaces of the two solid compounds, FeS2 and MnO2. At the FeS2 surface, Fe(III) oxidizes FeS2 and is thereby reduced to Fe(II) which is reoxidized to Fe(III) by MnO2. Bacteria could be enriched from anaerobic marine sediments, which anaerobically oxidize FeS, but not FeS2, with NO3- as electron acceptor. Bacteria were not obtained with amorphous Fe(III) oxide as electron acceptor. The result that bacteria do not attack FeS2 under anaerobic conditions has been confirmed in experiments using 55FeS2 as tracer. Keywords: pyrite oxidation, metal sulfide oxidation, thiosulfate mechanism, polysulfide mechanism, manganese oxide, iron oxide 1.
INTRODUCTION Biological metal sulfide oxidation at low pH below 4 in presence of oxygen, known as bioleaching, is well documented in the literature [for recent reviews see 1-3]. In the literature about bioleaching it has been regularly stated that bioleaching organisms oxidize metal sulfides by two different ways, "direct" and "indirect". "Direct" means that organisms are attached to the metal sulfide surface, dissolving the metal sulfide without a soluble electron shuttle. “Indirect” means that organisms are not attached to the mineral surface and that the metal sulfide is oxidized via the electron shuttle Fe(II)/Fe(III). So far, 55
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it has not been shown, how organisms oxidize metal sulfides in a "direct" way. High amounts of Fe bound in a layer of extracellular polymeric substances (EPS) of Acidithiobacillus ferrooxidans and of Leptospirillum ferrooxidans have been detected [46]. Recently, Ehrlich [2] suggested that this EPS bound Fe may serve as an electron shuttle, as Fe also does in the "indirect" way. Following this suggestion, Fe(III) is generally the oxidant for biological metal sulfide dissolution, irrespective if cells are attached ("direct") or not attached ("indirect") to the mineral surface. This inte rpretation is supported by a SEM study of Edwards et al. [7] who detected similar leaching patterns on metal sulphide surfaces in case of bioleaching and of abiotic Fe(III) leaching. As well, Rawlings [3] highlighted the dominant role of EPS bound Fe for bioleaching and stated in his review about the mechanisms of bioleaching, that the mechanism is strictu sensu indirect. A direct contact of a cell to the mineral surface is not essential for bioleaching but increases the rate of bioleaching. He suggests to replace the term "direct leaching" by the term "contact leaching". However, bioleaching of metal sulfides is carried out by acidophilic Fe(II) oxidizing organisms providing Fe(III) to oxidize metal sulfides via the thiosulfate or the polysulfide mechanisms. The intermediary sulfur compounds are either oxidized chemically by Fe(III) or biologically by acidophilic sulfur/-compound oxidizing organisms [5, 6, 8 -10]. Biological metal sulfide oxidation at neutral to alkaline pH in presence of oxygen is less well studied. Bioleaching organisms can not live at this pH and Fe is insoluble. Thus, it is not possible that metal sulfides are biologically oxidized in a similar way as described for low pH. A biological dissolution of the acid soluble metal sulfide FeS at neutral pH has been shown for moderately acidophilic sulphur compound oxidizing organisms like Thiomicrospira frisia [11, 12]. These organisms produce protons by sulfur oxidation which dissolve the acid soluble metal sulfide. According to the polysulfide mechanism, intermediary sulfur compounds like elemental sulfur are formed which are biologically oxidized. In case of the acid insoluble FeS2, moderately acidophilic sulfur compound oxidizing organisms like Thiomonas intermedia only oxidize intermediary sulfur compounds formed by the chemical FeS2 oxidation and do not increase the chemical FeS2 dissolution rate [13, 14]. Growth of microaerophilic, neutrophilic Fe(II) oxidizing organisms with FeS as substrate has been reported [15], but it is not known if these organisms increase the metal sulfide dissolution rate and which sulfur compounds are formed. In the absence of oxygen, an anaerobic biological metal sulfide oxidation at low pH below 4 has also been demonstrated. In presence of Fe(III) as sole electron acceptor, Acidithiobacillus ferrooxidans enhanced the solubilization of Cu from a CuFeS2 containing concentrate [16]. At low pH, Fe(III) is soluble and efficiently oxidizes metal sulfides including CuFeS2. According to the polysulfide mechanism, elemental sulfur accumulates in the course of the chemical CuFeS2 oxidation. Acidithiobacillus ferrooxidans, Acidithiobacillus thiooxidans and Sulfolobus acidocaldarius are able to oxidize elemental sulfur to sulfuric acid by reduction of Fe(III) [17, 18]. The production of sulfuric acid enhances the dissolution of the metal sulfide. However, Fe(II) as a product of the biological elemental sulfur oxidation accumulates because it cannot be oxidized by acidophilic microorganisms in the absence of oxygen. As a consequence, the availability of the oxidant Fe(III) is limited, and a continues supply of Fe(III) is necessary for an anaerobic biological metal sulfide oxidation at low pH. This process seems to relevant in miningimpacted lake sediments [19]. Furthermore, a purely chemical oxidation of FeS2 with MnO2 as sole electron acceptor at low pH has been demonstrated, and this process
56
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has been suggested for the dissolution of low grade ores or ocean bed nodules in acid media [20, 21]. In this paper, I report about several experiments being done to find out if and how metal sulfides are oxidized under anaerobic conditions at pH 8 in presence of different electron acceptors. To elucidate the anaerobic metal sulfide oxidation mechanisms, intermediate sulfur compounds were analyzed. Anaerobic metal sulfide oxidation at neutral to alkaline pH seems to be relevant in the environment, e.g. in sulfidic mine tailings or in marine sediments. The experiments have been previously described in detail by Schippers and Jørgensen [22, 23]. 2.
MATERIALS AND METHODS
2.1 Anaerobic chemical iron sulfides oxidation experiments FeS was obtained form Aldrich (34,316-1, -100 mesh, 99.9 %, troilite and pyrrhotite were the main minerals). FeS2 and 55FeS2 were prepared from pure chemicals as previously described [22] (pyrite and marcasite in minor amounts were the only detectable minerals. The surface area was 2.9 m2.g-1 measured by the BET-method). As oxidant NaNO3 was used as a pure chemical, amorphic Fe(III) oxide and MnO2 were freshly prepared as described elsewhere [24, 25]. For the experiments, 0.5 g of FeS or FeS2 was weighed into 250 ml flasks. To each assay, 50 ml of a 1 M NaHCO3 solution and 50 ml of either a NaNO3 solution (25 g/l), a suspension of amorphic Fe(III) oxid (1 mol L-1) or a suspension of MnO2 (1 mol L-1) were added. Five parallel assays were prepared for each combination of iron sulphide and oxidizing agent. Additional assays with amorphous Fe(III) oxide in presence of Fe-complexing organic compounds such as salicylic acid, oxalic acid, and citric acid, or in the presence of the electron transporting compound AQDS (2,6-anthraquinone disulfonate) in concentrations of 10 mM, 1 mM, or 0.1 mM each were prepared. Suspensions containing either iron sulfides or oxidizing agents were prepared as controls. The pH remained at 8 (+ 0.5) for all experiments. The flasks were closed with air-tight butyl rubber seals, evacuated, and gassed with a mixture of CO2/N2 (10/90, v/v) three times. All assays were incubated at 20°C in the dark. Chemical analysis was done as previously described [22]. 2.2 Enrichment of anaerobic iron sulfides oxidizing bacteria To enrich anaerobic, neutrophilic FeS and FeS2 oxidizing bacteria using NO3- or amorphous Fe(III) oxide as electron acceptor more than 300 assays were inoculated with material from more than 10 different anoxic marine sediments as previously described [23]. 3.
RESULTS
3.1 Anaerobic chemical iron sulfides oxidation FeS and FeS2 were chemically oxidized by MnO2 in a bicarbonate buffered solution at pH 8, whereas NO3- and amorphic Fe(III) oxide did not oxidize these iron sulfides. FeS and FeS2 were also not oxidized by amorphous Fe(III) oxide in presence of Fe-complexing organic compounds such as salicylic acid, oxalic acid, and citric acid, or in the presence of the electron transporting compound AQDS (2,6-anthraquinone disulfonate) in a carbonate buffered solution at pH 8. In the control experiments with iron sulfides and without oxidizing agents, a formation of oxidation products did not take place and amorphic 57
Bioleaching Applications
Fe(III) oxide or MnO2 were not dissolved in the absence of iron sulfides. Furthermore, elemental sulfur was not oxidized by any of the oxidizing agents tested (data not shown). The products from the chemical oxidation of FeS and of FeS2 by MnO2 are shown in Fig. 1. In the case of FeS, elemental sulfur and sulfate were the only oxidation products detected, whereas the oxidation of FeS2 produced a variety of oxidation products in high amounts. Sulfate was the main product and increased over a month to 30 mM. Sulfate formation was highest during the first 15 days. Tetrathionate accumulated until day 15 to 12 mmol S L-1 and then deceased again. Thiosulfate and trithionate increased to around 5 mmol S L-1, whereas pentathionate occurred only at concentrations below 1 mmol S L-1. Longer-chained polythionates or elemental sulfur were not formed. The concentration of Mn(II) increased to 120 mmol L-1 due to MnO2 reduction. The concentration of tetrathionate as the main sulfur intermediate decreased after the day 15 and sulphate increased further, indicating an oxidation of tetrathionate to sulfate by MnO2. At the end of the experiment when 2/3 of the total dissolved sulfur had been completely oxidized to sulfate, the dissolved Mn/S molar ratio was 2.5 and the Mn/Fe ratio was 5. -1
mmol S L 5 4,5 4 3,5 3 2,5 2 1,5 1 0,5 0 0
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pentathionate
Figure 1. Products of the chemical oxidation of (A top) FeS and of (B bottom) FeS2 by MnO2 at pH 8. For FeS2 also the formation of Mn(II) due to reduction of MnO2 is shown. The following concentrations were used: 0.5 mol L-1 MnO2 and 57 mol L-1 FeS (0.5 g) or 42 mol L-1 FeS2 (0.5 g) [22] 58
Bioleaching Applications
To verify if the sulfur compound intermediate thiosulfate is oxidized by MnO2 as oxidant as well, a separate experiment has been carried out. Within 5 hours, 2/3 of the initial amount of 10 mM thiosulfate was oxidized by MnO2, mainly to tetrathionate, some sulfate and a little pentathionate (data not shown). 3.2 Anaerobic biological iron sulfides oxidation From anaerobic marine sediments bacteria could not be enriched with amorphous Fe(III) oxide as electron acceptor. Instead, bacteria could be enriched, which anaerobically oxidize FeS, but not FeS2, with nitrate as electron acceptor. The result that bacteria do not attack FeS2 under anaerobic conditions has been confirmed in an experiment using 55FeS2 as tracer. In this experiment, a FeS oxidizing and nitrate reducing enrichment culture obtained from sediment of the estuary of the Rio Tinto, Spain, was further anaerobically cultivated at 30°C with 2 mM Fe2+ and a few mg So as substrates and 10 mM NO3- as electron acceptor in the presence of 50 mg "tracer-marked" 55FeS2 to test for co-oxidation of FeS2. Samples were taken from assays inoculated with bacteria or from control assays without bacteria and analyzed for concentrations of SO42-, NO3-, Fe2+, and 55Fe in the medium as previously described [22, 23]. After an incubation period of 2.5 month, SO42was formed and NO3- and Fe2+ were consumed in the assays with bacteria, presumably due to bacterial Fe2+ and So oxidation coupled to NO3- reduction. Values of 55Fe were not higher in the assays with bacteria than in the controls, which means that an anaerobic microbial dissolution of 55FeS2 could not be detected (data not shown). 4.
DISCUSSION
4.1 Anaerobic chemical iron sulfides oxidation The formation and degradation of sulfur compound intermediates in the course of FeS2 oxidation has been shown previously for the bioleaching of FeS2 in presence of oxygen by which FeS2 is degraded via the thiosulfate mechanism [5, 6, 8-10]. According to the molecular-orbital theory Fe(III) hexahydrate ions attack FeS2, oxidize S22- to thiosulfate and consequently cleave the chemical bonding between the Fe(II) and the S22in the FeS2 lattice. As a consequence, thiosulfate and Fe(II) occur as dissolution products. The Fe(II) is oxidized to regenerate Fe(III) for further attack, while thiosulfate is oxidized via tetrathionate, disulfane-monosulfonic acid and trithionate to mainly sulfate in a cyclic pathway. The results of the present study are in agreement with the thiosulfate mechanism. Conclusively, the anaerobic FeS2 oxidation by MnO2 proceeds via the thiosulfate mechanism as well. Since sulfate is the endproduct of the overall reaction, the oxidation of FeS2 by MnO2 proceeds by the following reaction: FeS2 + 7.5 MnO2 + 11 H+ ----> Fe(OH)3 + 2 SO42- + 7.5 Mn2+ + 4 H2O (1) According to this equation, the stoichiometry of the products is 3.75 for Mn/S and 7.5 for Mn/Fe. At the end of the experiments with FeS2, 2/3 of the dissolved pyritic sulfur was completely oxidized to sulfate. Stoichiometries of 2.5 for Mn/S and 5 for Mn/Fe were calculated which are exactly 2/3 of the stoichiometries above and, therefore, in agreement with equation (1). Oxidation rates were calculated per surface area for the FeS2 oxidation during the first 15 days of the experiment as described by Peiffer and Stubert [26]. For the calculation, the amounts of total dissolved S or Fe after 15 days and the total surface area of the FeS2 were
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used. For total dissolved S and total dissolved Fe the rates were 1.02 and 1.12 nmol.m-2.s-1, respectively. Both FeS2 and MnO2 are insoluble solid compounds and it is not known how electrons are transported from FeS2 to MnO2. Alternatively, a tight contact of the surfaces of the two solid compounds might enable a direct electron transfer, or electrons might be transported via an electron-shuttling compound. It is assumed that Fe(II)/Fe(III) cycling transports electrons based on the following reasons: A) Fe(III) was detected in our experiments and is a well-known oxidant for FeS2: FeS2 + 14 Fe3+ + 8 H2O ----> 15 Fe2+ + 2 SO42- + 16 H+
(2)
Despite the low solubility of Fe(III) at circumneutral pH, Fe(III) may serve as oxidant if it remains adsorbed onto the surface of FeS2 [26, 27]. B) The surface related reaction rates of 1.02 and 1.12 nmol.m-2.s-1 for FeS2 oxidation by MnO2 are in the low range of the rates described for the oxidation of FeS2 by Fe(III) at circumneutral pH, 1.1-39 nmol.m-2.s-1 [26]. This could imply that in our experiments Fe(III) is the oxidant for FeS2 as well. C) It was shown that Fe(II) is oxidized by MnO2. Postma and Appelo [28] describe this reaction by the following equation: 2 Fe2+ + MnO2 + 2 H2O ----> 2 FeOOH + Mn2+ + 2 H+
(3)
Fe(III) generated by the reaction of Fe(II) with MnO2 could react with FeS2 before it precipitates as FeOOH. D) The coupling of the redox pairs FeS2/Fe(III) and Fe(II)/MnO2 has been suggested for the dissolution of low grade ores or ocean bed nodules in acid media [20, 21]. E) Molecular-orbital theory considerations by Luther [29, 30] support Fe(II)/Fe(III) cycling. Fe(II) is a d6 (t2g6) electron configuration and Mn(IV) is a d3 (t2g3) electron configuration. The t2g orbitals are filled in case of Fe(II) and half-filled in case of Mn(IV) which imparts the stability for these metal ions. For both solids, FeS2 and MnO2, to react with each other, a ligand would have to dissociate as both reactants touched, but this does not occur. Soluble Fe(II), which has a t2g4 eg*2 electron configuration, is high spin and labile, thus it can adsorb to and react with MnO2 to form Fe(III). Soluble Fe(III) has d5 (t2g3 eg2) electron configuration and is therefore a labile cation that can undergo ligand exchange and is therefore able to react with the S22- ligand of FeS2. Summarizing the results, a model of FeS2 oxidation by MnO2 is shown in Fig. 2. It is postulated that electrons are transported via the Fe(II)/Fe(III)-shuttle if FeS2 and MnO2 are in a close contact. Fe(II) and Fe(III) should be adsorbed onto the surface of FeS2 because our experiments with amorphic Fe(III) oxide have shown that precipitated Fe(III) does not oxidize FeS2. However, while amorphic Fe(III) oxides precipitated to the FeS2 surface do not alone oxidize FeS2, they may serve as an electron conduit [31]. Electrons might flow from FeS2 via Fe(III) oxides to adsorbed Fe(III) or via Fe(III) oxides directly to MnO2. In our experiments with MnO2, only precipitated Fe(III) but not Fe(II) was detected by extraction with HCl, indicating that the reaction between Fe(II) and MnO2 is faster than the reaction between Fe(III) and FeS2. All reactions in this model were shown to be purely chemical, however, biological catalysis could be involved in the degradation of sulfur intermediates [9, 32, 33].
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Figure 2. Model of anaerobic FeS2 oxidation by MnO2 via the Fe(II)/Fe(III)-shuttle [22]
In nature, Fe(III) might be stabilized in solution complexed to organic ligands, thus the Fe(II)/Fe(III)-shuttle might transport electrons even if FeS2 and MnO2 are not in close contact. Soluble complexed Fe(III) has been shown to exist in sediment pore waters [34, 35]. Furthermore, the FeS2 oxidation rate increases in presense of Fe(III)-chelating ligands [26, 30]. Besides MnO2, other manganese oxides might oxidize FeS2 as well because the standard redox potential of the couples MnO2/Mn2+, Mn3O4/Mn2+, and MnOOH/Mn2+ are all around 600 mV [36]. In contrast to FeS2 oxidation, FeS oxidation by MnO2 only produced elemental sulfur and some sulfate as oxidation products. According to the polysulfide mechanism [5, 6, 10], acid soluble metal sulfides are dissolved by Fe(III) and proton attack. The sulfide is oxidized via radicals and polysulfides mainly to elemental sulfur besides some sulfate, which is in agreement with our results. Therefore, the chemical FeS oxidation is described by the following equation: FeS + 1.5 MnO2 + 3 H+ ----> Fe(OH)3 + So + 1.5 Mn2+
(4)
Unlike MnO2, amorphous Fe(III) oxide was not an oxidant for FeS2 or FeS at pH 8 in the experiments of this study, even not in the presence of organic Fe-complexes or of the electron transporting compound AQDS. Luther et al., [30] showed a chemical FeS2 oxidation by 1 mM ferrihydrite and 10 mM salicylic acid in the pH range of 4 to 6.5. Ferrihydrite and salicylic acid form a Fe(III) salicylate complex which reacts with FeS2. Liu and Millero [37] showed that the solubility of Fe(III) in the presence of Fe(III) complexing humic acids is two orders of magnitude higher at pH 4-6 than at pH 8. Presumably, the concentration of complexed Fe(III) in the experiments of this study at pH 8 was too low to enable FeS2 dissolution. FeS2 and FeS were also not chemically oxidized by NO3- in the experiments of this study. Ottle y et al. [38] have shown, that a chemical oxidation of Fe(II) by NO3- can be catalyzed by metals such as copper. Thus, a chemical oxidation of FeS2 by NO3- might exist as well. A FeS2 oxidation by the reduction of NO3- has been suggested for aquifers based on geochemical data [39, 40] but clear experimental evidence is lacking. Bacteria may be involved in this process.
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4.2 Anaerobic biological iron sulfides oxidation From anaerobic marine sediments bacteria could be enriched which anaerobically oxidize FeS with nitrate as electron acceptor. This finding is in agreement with results of Garcia-Gil and Golterman [41] who described a FeS-mediated denitrification for a marine sediment. FeS belongs to the acid soluble metal sulfides which are chemically oxidized via polysulfides to mainly elemental sulfur and some sulfate [5, 6, 10]. Due to its acid solubility, protons dissolve FeS according to:
FeS + H+ ----> Fe2+ + HS-
(5) -
Both products of this reaction may be oxidized by NO3 reducing bacteria. The Fe2+ can be oxidized according to Straub et al. [42]: 10 FeCO3 + 2 NO3- + 24 H2O ----> 10 Fe(OH)3 + N2 + 10 HCO3- + 8 H+
(6)
-
HS may be oxidized by e.g. Thiobacillus denitrificans or Thiomicrospira denitrificans [11]: 5 HS- + 8 NO3- + 3 H+ ----> 5 SO42- + 4 N2 + 4 H2O
(7)
In equation 6, protons are produced which continue to dissolve FeS. Bacteria might be attached to the FeS surface embedded in extracellular polymeric substances (EPS). Bacteria produce EPS to create a microenvironment which favours their metabolisms [5, 6]. In such a microenvironment, the pH might be much lower than 8 enabling FeS dissolution. Consequently, Fe2+ or HS- oxidizing and NO3- reducing bacteria can grow with FeS as a substrate, and I was able to enrich these bacteria from different marine sediments. With FeS2 as a substrate, bacteria did not grow since FeS2 cannot be dissolved by protons. Precipitation of Fe(III) hydroxide might explain the absense of 55FeS2 dissolution in a o S and Fe2+ oxidizing and NO3- reducing bacterial culture. The bacteria oxidize Fe(II) to Fe(III) which has to diffuse from the Fe-oxidizing enzyme of the bacteria to the FeS2 surface to serve as an oxidant for FeS2. Obviously, Fe(III) precipitates immediately and therefore cannot serve as an oxidant for FeS2. In aquifers where slightly acidic pH values were detected, a FeS2 oxidation by the reduction of NO3- has been suggested based on depth profiles of NO3- and SO42- [39, 40]. There, Fe2+ oxidizing and NO3- reducing bacteria and soluble organic Fe(III) complexes could probably catalyze an anoxic FeS2 oxidation with NO3- as electron acceptor. REFERENCES
1. H. Brandl, in H.-J. Rehm and G. Reed in cooperation with A. Pühler and P. Stadler (eds.), Biotechnology, Vol. 10, Wiley-VCH, Weinheim, Germany (1991) 191. 2. H.L. Ehrlich, Geomicrobiology, Marcel Dekker Inc., New York, 2002. 3. D.E. Rawlings, Ann. Rev. Microbiol., 56 (2002) 65. 4. T. Gehrke, J. Telegdi, D. Thierry and W. Sand, Appl. Environ. Microbiol., 64 (1998) 2743. 5. W. Sand, T. Gehrke, P.-G. Jozsa and A. Schippers, in R. Amils and A. Ballester (eds.), Biohydrometallurgy and the environment toward the mining of the 21st century, Part A, Elsevier, Amsterdam (1999) 27. 6. W. Sand, T. Gehrke, P.-G. Jozsa and A. Schippers, Hydrometallurgy 59 (2001) 159. 7. K.J. Edwards, B. Hu, R.J. Hamers and J.F. Banfield , FEMS Microbiol. Ecol., 34 (2001) 197. 8. Schippers, P.-G. Jozsa and W. Sand, Appl. Environ. Microbiol., 62 (1996) 3424. 62
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9. A. Schippers, T. Rohwerder and W. Sand, Appl. Microbiol. Biotechnol., 52 (1999) 104. 10. A. Schippers and W. Sand, Appl. Environ. Microbiol., 65 (1999) 319. 11. J.G. Kuenen, L.A. Robertson and O.H. Tuovinen, in A. Balows, H.G. Trüper, M. Dworkin, W. Harder and K.-H. Schleifer (eds.), The Prokaryotes, Springer Verlag, Berlin (1992) 2638. 12. T. Brinkhoff, G. Muyzer, C.O. Wirsen and J. Kuever, Int. J. Syst. Bacteriol., 49 (1999) 385. 13. G.J.M.W. Arkesteyn, Plant and Soil, 54 (1980) 119. 14. A. Schippers, H. von Rège and W. Sand, Minerals Engineering, 9 (1996) 1069. 15. D. Emerson and C. Moyer, Appl. Environ. Microbiol., 63 (1997) 4784. 16. A. Das and A.K. Mishra, Appl. Microbiol. Biotechnol., 45 (1996) 377. 17. T.D. Brock and J. Gustafson, Appl. Environ. Microbiol., 32 (1976) 567. 18. J.T. Pronk and D.B. Johnson, Geomicrobiol. J., 10 (1992) 153. 19. K. Küsel, U. Roth, T. Trinkwalter and S. Peiffer, Environ. Experim. Bot., 46 (2001) 213. 20. S.B. Kanungo, Hydrometallurgy, 52 (1999) 313. 21. R.K. Paramguru and S.B. Kanungo, Can. Metal. Quart. 37 (1998) 389. 22. A. Schippers and B.B. Jørgensen, Geochim. Cosmochim. Acta, 65 (2001) 915. 23. A. Schippers and B.B. Jørgensen, Geochim. Cosmochim. Acta, 66 (2002) 85. 24. D.R. Lovley and E.J.P. Phillips, Appl. Environ. Microbiol., 51 (1986) 683. 25. D.R. Lovley and E.J.P. Phillips, Appl. Environ. Microbiol., 54 (1988) 1472. 26. S. Peiffer and I. Stubert, Geochim. Cosmochim. Acta, 63 (1999) 3171. 27. C.O. Moses and J.S. Herman, Geochim. Cosmochim. Acta, 55 (1991) 471. 28. D. Postma and C.A.J. Appelo, Geochim. Cosmochim. Acta, 64 (2000) 1237. 29. G.W. III Luther, Geochim. Cosmochim. Acta, 51 (1987) 3193. 30. G.W. III Luther, J.E. Kostka, T.M. Church, B. Sulzberger and W. Stumm, Mar. Chem., 40 (1992) 81. 31. C.M. Eggleston, J.-J. Ehrhardt and W. Stumm, Amer. Mineral. 81 (1996) 1036. 32. B. Thamdrup, K. Finster, J. Würgler-Hansen and F. Bak, Appl. Environ. Microbiol., 59 (1993) 101. 33. K. Finster, W. Liesack and B. Thamdrup, Appl. Environ. Microbiol., 64 (1998) 119. 34. M. Huettel, W. Ziebis, S. Forster and G.W. III Luther, Geochim. Cosmochim. Acta, 62 (1998) 613. 35. M. Taillefert, A.B. Bono and G.W. III Luther, Environ. Sci. Technol. 34 (2000) 2169. 36. B. Thamdrup, in B. Schink (ed.) Advances in Microbial Ecology, Kluwer Academic/Plenum Publishers, New York (2000) 41. 37. X. Liu and F.J. Millero, Geochim. Cosmochim. Acta, 63 (1999) 3487. 38. C.J. Ottley, W. Davison and W.M. Edmunds, Geochim. Cosmochim. Acta, 61 (1997) 1819. 39. D. Postma, C Boesen, H. Kristiansen and F. Larsen, Water Resources Research 27 (1991) 2027. 40. P. Engesgaard and K.L. Kipp, Water Resources Research, 28 (1992) 2829. 41. L.J. Garcia-Gil and H.L. Golterman, FEMS Microbiol. Ecol. 13 (1993) 85. 42. K.L. Straub, M. Benz, B. Schink and F. Widdel, Appl. Environ. Microbiol., 62 (1996) 1458.
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Bacterial growth and propagation in chalcocite heap bioleach scenarios J. Petersena and D.G. Dixonb a
Department of Chemical Engineering, University of Cape Town, Rondebosch 7700, South Africa b Department of Metals and Materials Engineering, University of British Columbia, 6350 Stores Road, Vancouver, B.C. V6T 1Z4, Canada Abstract Heap bioleaching of chalcocite ores is widely practised as a relatively low cost process option, especially for marginal deposits. However, the rates of copper extraction achieved in actual operations are often slower than expected. A recent study by the authors has shown that chalcocite heap bioleaching is controlled by a two-stage mechanism. The first stage is rapid and ultimately controlled by the supply of acid to the reaction sites. The second stage is intrinsically much slower and controlled by mineral oxidation kinetics. In this work, the results of a column operated with a bacterial consortium grown under idealised laboratory conditions are compared to those of a column run with a native culture and high-TDS (total dissolved solids) raffinate solution from the mine site. The results clearly indicate that while the leach reactions follow the two-stage mechanism in both cases, the overall rate of copper extraction in the high-TDS column is significantly retarded. Bacterial growth in this column is slower and much more limited. It is postulated that in such high-TDS environments the rate of copper leaching is controlled by the rate of bacterial growth and oxidation. Bacterial growth and propagation trends observed in the experiments have been reproduced with a comprehensive heap simulation tool developed by the authors. The bacterial growth and oxidation model employed is briefly introduced. The simulations confirm that, if the bacterial growth rate is significantly retarded, it can become rate controlling over other factors. In full-scale heaps, however, the principal rate-limiting factor is still the diffusion of acid into large stagnant zones. Keywords: heap leaching, bio-oxidation, growth kinetics, yield, TDS, growth inhibition 1.
INTRODUCTION Heap bioleaching of chalcocite ores has become widely practised for the economic extraction of copper from low-grade deposits. However, in many operations the rate of copper extraction falls well behind what can be achieved in laboratory columns. Some concerns have been expressed that field conditions (high altitude, low temperatures, high TDS (total dissolved solids) of process waters) are adverse to bacterial growth, and may therefore be the cause for slow extraction rates. Some successes with improving heap performance by adapting cultures to such conditions have been reported [1], but there 65
Bioleaching Applications
appears to be only a limited understanding of the interactions between minerals and bacteria in chalcocite heap leach situations. The authors have conducted a comprehensive study into the dynamics of chalcocite heap bioleaching, and some results are presented in a previous paper [2]. Based on fundamental electrochemical investigations [3], extensive column studies and simulations with a comprehensive heap leach modelling tool, it has been found that the rate of chalcocite leaching by ferric ions, both in columns and in full scale unconfined heaps, is driven by a two-stage mechanism. In stage 1, approximately 40% of copper from chalcocite is released to form an interim pseudo-covellite compound: (1) Cu2S + 0.8 Fe2(SO4)3 → 0.8 CuSO4 + Cu1.2S + 1.6 FeSO4 The mineral undergoes significant structural changes during stage 1 leaching, such that the second stage, which releases the remaining copper, must be viewed as a separate reaction: (2) Cu1.2S + 1.2 Fe2(SO4)3 → 1.2 CuSO4 + S° + 2.4 FeSO4 In bioleaching the necessary ferric is in each case regenerated continuously through bio-oxidation: 2FeSO4 + 0.5 O2 + H2SO4 → Fe2(SO4)3 + H2O (3) The first stage of chalcocite oxidation is kinetically very rapid and is controlled primarily by the supply of ferric to the reaction site, and hence by the rate of ferrous to ferric oxidation facilitated by bacteria. As ferric consumption is virtually instantaneous, solution potentials are generally quite low (< 500 mV Ag/AgCl). In heaps the only acid available for reaction (3) during this stage is that coming in with the feed, since pyrite oxidation is typically limited and no elemental sulfur is being generated for subsequent bio-oxidation to sulfate. This restricts the stage 1 oxidation reaction to a narrow zone, which progresses downwards, commensurate with the rate of acid supply. In unconfined heaps this phenomenon is complicated further by transport effects in stagnant zones. These have been used to explain the large discrepancies in leach rates between full-scale heaps and column experiments [2]. Stage 2 oxidation proceeds in the wake of the stage 1 zone. This step is kinetically much slower at ambient temperatures. During bioleaching, the slow kinetics result in a build-up of ferric near the mineral surface and consequently the reaction proceeds at much higher solution potentials (> 650 mV Ag/AgCl). At the same time elemental sulfur is generated, which can be oxidised to generate sulfuric acid, and pyrite oxidation can also proceed more favourably. As the rate of reaction is relatively slow, it proceeds homogeneously over the entire height of the column (or heap), once stage 1 has passed through. Bacterial growth and oxidation kinetics in chalcocite heaps must be investigated against this background. It is conceivable that reaction (3) is severely retarded by bacteria growing under adverse conditions, and thus becomes overall rate limiting over other factors. In the present study, column leach experiments using a laboratory culture are compared with those using a native culture growing in a high-TDS raffinate. The different growth rates are quantified and the effect in large columns and heaps is investigated with the aid of a comprehensive modelling tool.
66
Bioleaching Applications
2.
EXPERIMENTAL
2.1 Column Studies The ore used in all experiments originated from a typical copper mining and heap leach operation. Top particle size was ¾ inch and the minus 100 micron fines content was 10%. The sample ore had an average Cu grade of 1.45%, approximately 30% of which was acid-soluble (mostly brochantite) and the rest was chalcocite, with some minor occurrences of chalcopyrite. The pyrite content was approximately 2.5%. The experimental work was conducted in mini-columns, 400 mm tall and 100 mm in diameter, immersed in a water-bath with the temperature controlled at 25°C. In each experiment a number of these columns were run in series, with solution from one column collected in a small, sealed interim container and pumped from there into the next column by means of a peristaltic pump. The standard experimental irrigation rate was 5 L/m2-hr and all columns were well aerated with air enriched with 1% v/v CO2. Before charging to the columns, the ore was acid agglomerated in accordance with operating practice at the mine site. All columns were operated on a once-through basis. Two column experiments, denoted Z5 and Z8, were run in the context of this study. Z5 had 8 columns in series with a total height of 2.5 m. The experiment was inoculated with a laboratory bacterial culture derived from a blend of pure cultures of Acidithiobacillus ferrooxidans, Acidithiobacillus thiooxidans and Leptospirillum ferrooxidans. This had been continuously maintained on a minus 40-mesh fines fraction of the original ore sample in standard K9 medium, in an incubator shaking at 150 rpm at 30°C, with weekly transfers. The feed solution consisted of 7.5 g/L H2SO4, 0.7 g/L Fe(III) as sulfate and 1.3 g/L Fe(II) as sulfate. This solution reflects the pH, potential and Fe concentrations reported from the mine. Column experiment Z8 consisted of four columns in series with a total height of 1.2 m. This experiment was inoculated with a native culture obtained from the mine site, which was well adapted to the high TDS raffinate solution, and the original raffinate sampled run-of-production at the mine site. The culture was maintained in the same way as the laboratory culture, but in the original raffinate augmented with K9 culture medium. The composition of the original raffinate is given in Table 1. Of note are the extremely high concentrations of Al and Mg, which exceed toxicity limits reported in the literature [1, 4, 5], and the elevated chloride levels. Table 1. Analysed composition of the mine site raffinate (only components >10 mg/L listed) Element Al Ca Co Cu Fe Mg Mn
Concentration [mg/L] 12,200 467 16.2 216 2,460 10,100 669
Element P K Na Zn Cl– F– NO2–
Concentration [mg/L] 221 29.0 1,670 376 1,300 80.1 28.1
NO3– o-PO4 SO4
Concentration [mg/L] 105.9 532 116,880
pH E (mV vs SHE)
1.24 640
Element
Figure 1 shows the copper extraction vs. time achieved in both tests. Although Z5 was halted to allow early assay of solids, the expected extraction trend (dashed line) - based on 67
Bioleaching Applications
observations from similar experiments - has been added for comparison. In both experiments the first 30% of copper extraction represents the rinsing of acid-soluble copper dissolved upon acid agglomeration. The next 30% of copper extraction corresponds roughly to stage 1 leaching, and it is in this phase where the two experiments differ significantly in terms of extraction rate: In Z5 this reaction is nearly complete within 20 days, whereas Z8 requires almost 40, although being only half as tall. Leaching beyond 60% extraction continues in Z8 at a rate similar to that expected for Z5 (although not measured in this particular experiment). 100%
Cu extraction
80%
60% 40% Exp. Z5
20%
Exp. Z8 0% 0
10
20
30
40
50
60
70
days on stream
Figure 1. Copper extraction in experiments Z5 and Z8. The dashed line indicates trends observed from similar experiments
Figures 2 shows the effluent pH and the progression of solution potentials along the height of the columns. With regard to Z5 these two figures illustrate that chalcocite stage 1 leaching progresses in a narrow zone down the column, consuming all available acid and ferric, and thus maintaining the solution potential at very low levels. 800
3
Pot ential (vs . Ag/AgCl)
Exp. Z5 Exp. Z8
pH
2.5
2
1.5
1
700
Z5 - 0.3 m Z5 - 0.9 m Z5 - 1.8 m
600
500
Z5 Z8 Z8 Z8
- 2.4 - 0.3 - 0.6 - 0.9
m m m m
400
Z8 - 1.2 m
300 0
10
20
30
40
days on s tream
50
60
70
0
10
20
30
40
50
60
70
days on s tream
Figure 2. pH measured in the effluent and solution potentials measured over the length of the column from experiments Z5 and Z8. The dashed lines indicate trends observed from similar experiments
In the wake of this zone (stage 2 leaching) the potential is rising rapidly to levels around 700 mV (vs. Ag/AgCl), and once the zone breaks through (not quite achieved in Z5 before shut-down, but observed in similar experiments (dashed line)), the solution pH gradually drops to feed levels. In Z8 the zone is still observed, but it progresses much more slowly, and the transition from stage 1 to stage 2 leaching is much more gradual. Only some of the available acid is consumed, but never depleted, and the pH remains consequently more or less stable with only a minor peak around day 50, when the low potential front breaks through. Thus it is clear that stage 1 leaching proceeds through Z8 68
Bioleaching Applications
within a broad band rather than a narrow zone, and it is not controlled by acid supply as opposed to Z5. The cause for these discrepancies becomes clear from Figure 3. Plotting the number of bacteria counted in solution as a function of time and depth in the heap shows counts lower by an order of magnitude in Z8 (native bacteria in high-TDS raffinate) as compared to Z5 (laboratory culture in artificial raffinate). Both sets of data show the same progression trends, however, with numbers moving down the columns in a "growing wave". The propagation rate of this wave corresponds to that of the high potential wave in the wake of the chalcocite stage 1 leach front. Therefore, it is postulated that the slower copper extraction rate in Z8 is linked to the slower propagation of bacteria through the column at much smaller numbers. The rate of copper leaching is hence controlled by bacterial growth kinetics rather than acid supply.
1.0E+08
1.E+0 8
1.E+0 7 Cell Count per mL
Cell Count per mL
1.0E+07
1.0E+06
1.0E+05
1.E+0 6
1.E+05 1.0E+04 m 0. 3
m 0. 6
m 0. 9
m 1. 2
Column Depth
m 1. 5
m 1. 8
m 2. 1
m 2. 4
we ek 3 we ek 2 we ek 1
1.E+04 0 .3
m
0 .6
m
0.9
m
1.2
m
we e we k 8 e we k 7 ek we 6 ek 5 we ek 4
Column Depth
Figure 3. Bacterial counts over the length of the column in experiments Z5 and Z8 2.2 Bacterial Growth Curves The growth characteristics of the cultures used in the column experiments were investigated in a series of shake-flask experiments. Two sets of seven flasks containing 75 mL of culture medium (standard K9 for the laboratory culture and original raffinate augmented with K9 growth medium for the native culture, each containing 2 g of ore fines) were prepared. At time zero, 25 mL of the respective mature culture (grown for 5 days in the same medium in a shaker incubator at 30°C, 150 rpm) was introduced into each medium and placed in the shaker. Both sets were prepared in duplicate, and a blind test was also prepared. Flasks were removed after 20 minutes (for the time-zero sample), 1, 2, 3, 5, 7 and 9 days, and left to settle for 5 minutes, after which a 5-mL sample was drawn from the supernatant, and from which the bacterial culture was enumerated. The resulting growth curves are shown in Figure 4. Clearly the laboratory culture grows much more rapidly and to larger numbers than the native culture. Interesting is the initial "overshoot" in the lab culture, before numbers settle to lower levels. This is thought to correspond to the stage-wise leaching of chalcocite, similar to the trends observed in the column experiments. The growth curves were analysed to extract initial growth rate and final cell yield, as reflected in Table 3. This suggests that the native culture grew about 6 times more slowly than the laboratory culture, and at one fifth of the yield. 69
Bioleaching Applications
Figure 4. Growth curves for the laboratory culture in the basic culture medium and the native culture in the original raffinate Table 2. Growth rate and final yield of growth experiments Culture
Growth rate (h–1)
Doubling time (h)
Final Yield (cells/mL)
Laboratory
0.109
6.36
1.10 × 108
Native
0.019
36.3
2.27 × 107
3.
MODELLING STUDY The authors have developed a comprehensive heap leach modelling tool (HeapSim) [2, 6], which combines an advection-diffusion model to account for transport of solutes through the heap (or column) to mineral sites in the ore, with a multi-reaction model at the mineral site. Gas adsorption, microbial growth and oxidation, and mineral leaching are represented through appropriate kinetic terms. A comprehensive description of the model is beyond the scope of this paper, but the equations describing bacterial growth and oxidation, as well as some key parameters, are discussed below. The bacterial growth model essentially follows Monod type kinetics [5]: dX = Xk g Π dt (4)
where X denotes the bacterial population per unit volume (i.e. [cells/mL]), kg is the growth rate (also commonly referred to as ⎧max, [h-1]), and Π denotes the product of a number of terms describing substrate and kinetic limitations (involving Fe2+, Fe3+, O2, acid, T, etc.). Some forms of the Π term have been reviewed by Nemati et al. [7]. The rate of ferrous oxidation is related to bacterial growth by a simple yield coefficient, Y [cells/mol Fe2+], thus: rFe2+ = −
1 dX Y dt
(5) The HeapSim code has been calibrated to model the column data generated in experiment Z5 and Z8 on the basis of as much independent bench and literature data as possible. With respect to bio-oxidation, the bacterial growth rate kg was obtained directly from the growth curve experiments described above (values in Table 2). The yield coefficient Y (which does not correspond to the maximum cell yield given in Table 2) was obtained for Z5 (and a number of other column studies using the laboratory culture, not 70
Bioleaching Applications
detailed here) by trial and error as 2 × 1012 cells/mol Fe2+. Based on the growth experiments, it was decided to take the yield coefficient for Z8 five times smaller than for Z5, i.e., as 0.4 × 1012 cells/mol Fe2+. The simulated copper extraction curves for both Z5 and Z8 and their closeness to the experimental data are shown in Figure 5. It should be stressed that the set of parameters for these simulations was identical except for the values of kg, Y, and column height. For Z5 the fit to the experimental data is very close, while the fit of Z8 is reasonable, considering that only the bacterial growth parameters were modified. These results thus lend some credence to the correctness of the modelling approach. 100%
Copper extraction
80% 60% 40%
Z5 Z5 Z8 Z8
20%
Data Modelled Data Modelled
0% 0
10
20
30
40
50
60
70
Time on stream [d]
Figure 5. Modelled extraction curves and experimental data for Z5 and Z8
Figure 6 shows the progression of the bacterial population in solution for both simulations. Both follow the rising wave pattern observed in the experiments, and the peak levels of bacterial counts correspond closely to those observed in the experiments. The simulated and measured peak levels between Z5 and Z8 are at a ratio of approximately 5, confirming that the selection of a yield coefficient for Z8 five times smaller than for Z5 was correct. Different from the measured data, however, the simulated curves do not progress as sharp fronts. In the model, bacterial transport is modelled as advectiondiffusion with concomitant Langmuir-type (physical) adsorption. The experimental data suggests, however, that bacteria do not migrate beyond the chalcocite stage 1 zone, but accumulate where there is good substrate (i.e. ferrous and acid) supply. Thus bacterial attachment appears to be part of the growth process rather than merely a physical phenomenon. This is not reflected in the model at present. 1.E+08
Bact erial count per mL
Bact erial count per mL
1.E+08
1.E+07
1.E+06
1.E+05
1.E+07
1.E+06
1.E+05
1.E+04
1.E+04 0.0
0.5
1.0
1.5
2.0
Col umn Dept h [m]
2.5
3.0
0.0
0.2
0.4
0.6
0.8
1.0
1.2
1.4
Col umn Dept h [m]
Figure 6. Modelled trends for bacterial propagation in experiments Z5 (left) and Z8
The simulations support the hypothesis that the slower extraction rate in Z8 can be explained by the much lower bacterial growth rate and yield. However, as was stated in the introduction, chalcocite stage 1 leaching in columns is normally limited by acid supply. It is of interest, therefore, to investigate the conditions under which transition from 71
Bioleaching Applications
time for 98% stage 1 conversion [d]
acid-limited to bacterial growth-limited leaching occurs. A number of simulations of Z8 for various values of kg (0.015 to 0.17 h–1, corresponding to doubling times of 4 to 48 h) and Y (0.4 to 10×1012 cells/mol Fe2+) were run, and the predicted times required to achieve 98% chalcocite stage 1 conversion are plotted against growth rate constant in Figure 7. From this it becomes clear that for growth rates above about 0.07 h–1 little improvement in conversion time is achieved, indicating acid-limited conditions. This holds true even for growth rates considerably below 0.07 h–1, provided that the cell yield is sufficiently low. The conditions of the native culture in Z8, however, fall clearly outside acid-limited conditions. It must be stressed that this applies strictly only to a short column scenario as given in experiment Z8. 60 y = 0.4
50
y =2
40
y = 10
30
bacterial growth limited
20
acid limited
10 0 0
0.05
0.1
0.15
0.2
growth rate const. [1/hr]
Figure 7. Time taken for 98% chalcocite stage 1 conversion as a function of various growth rates and yield coefficients, modelled for a Z8-type experiment (1.2 m column). The dashed line indicates the transition from acid-limited to growth-limited conditions
It is now of interest to investigate the effect that growth-limited cultures like native culture used in the present study would have on the rate of extraction in a full-scale heap. Retaining the calibration achieved for Z5, the HeapSim code was run to simulate two 10m heap scenarios, one with the parameters of the laboratory culture, and one with those of the Zaldívar culture. In addition, two further simulations of 10-m columns, 10 cm in diameter, were also run. Changing from a column simulation to a heap simulation involves changing the model parameter, which describes the length of stagnant pores in the bed, by as much as an order of magnitude [2]. Copper extractions are plotted against time in Figure 8. While the retarded bacteria have quite a noticeable effect on the laboratory columns, the effect is marginal in the heap scenario. As discussed above, the native culture does indeed become overall rate limiting in a narrow-bore column scenario. In unconfined heaps, however, diffusion of acid from the flowing solution into large stagnant zones (through which no solution flow occurs) governs the rate of acid supply to the reaction sites. This process is so slow that even under growth-limiting conditions the rate of bacterial oxidation is still not limiting the overall process.
72
Bioleaching Applications 100%
Cu extraction
80%
60%
40% Column fast bacteria Column slow bacteria
20%
Heap fast bacteria Heap slow bacteria
0% 0
60
120
180
240
300
360
Time on stream [d]
Figure 8. Copper extraction vs. time simulated for 10 m columns and heaps with and without bacterial rate limiting parameters 4.
CONCLUSIONS Chalcocite bioleaching proceeds in two stages, the first of which requires rapid biooxidation of ferrous to ferric, which is typically limited by the availability of acid at the reaction site rather than bacterial growth. In the present study a native culture growing in a high-TDS raffinate was compared against a laboratory consortium. It was found that in laboratory columns the native culture displayed severely restricted growth behaviour, thus governing the overall rate of leaching. Simulations with the HeapSim modelling tool, calibrated on the basis of bench-scale data, emulated the measured extraction data fairly well and thus confirmed the observed trends in bacterial numbers and propagation. In fullscale heaps, however, acid diffusion through large stagnant zones would still most likely govern the overall rate of copper extraction, despite restricted bacterial growth.
ACKNOWLEDGEMENTS The authors wish to thank Placer Dome Technical Services Limited for their generous support. REFERENCES
1. C. Garcia G., J. Binvignat T., Jaime Roco R., and J. Campos B., Randol at Vancouver ’98 Copper Hydromet Roundtable, 1998, pp 249-254. 2. J. Petersen and D.G. Dixon in Hydrometallurgy 2003, Proceedings of the 5th International Symposium Honoring Professor Ian M. Ritchie, Volume 1, C.Young, A. Alfantazi, C. Anderson, A. James, D. Dreisinger, B. Harris (eds.), TMS Publishers, Warrendale, PA, 2003, pp 351-364. 3. S.A. Bolorunduro, “Kinetics of leaching of chalcocite in acid ferric sulfate media: chemical and bacterial leaching”, M.A.Sc. Thesis, University of British Columbia, 1999. 4. O.H. Touvinen, S.I. Niemelä and H.G. Gyllenberg, Antonie van Leeuwenhoek 37, 1971, pp 489-496. 5. G. Rossi, Biohydrometallurgy, McGraw-Hill, Hamburg, 1990. 6. J. Petersen and D.G. Dixon in P.R. Taylor (Ed.) EPD Congress 2002, TMS Publishers, Warrendale, PA, 2002, pp 757-771. 73
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7. M. Nemati, S.T.L. Harrison, G.S. Hansford and C. Webb, Biochemical Engineering Journal 1, 1998, pp 171-190.
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Bacterial leaching studies of a Portuguese flotation tailing M.C. Costaa, N. Carvalhoa, N. Iglesiasb, I. Palenciab a
Faculdade de Engenharia de Recursos Naturais, Universidade do Algarve, Campus de Gambelas, 8005-139 Faro, Portugal b Departamento de Ingeniería Química, Facultad de Química, Universidad de Sevilla, 41071 Sevilla, Spain
Abstract Two bioleaching processes - a direct bioleaching and an indirect bioleaching with chemical and biological separate stages - have been applied to the flotation tailings of a chalcopyrite ore. In the present paper the influence of experimental parameters is investigated and the obtained results are described and compared in order to test the efficiency of these processes for the recovery of copper and zinc. Keywords: bacterial leaching, copper, zinc, tailing 1.
INTRODUCTION Neves Corvo mine – one of the major mines of the Iberian pyrite Belt, located in Southern Portugal – produces high-grade copper (and tin) sulphide ores which are submitted to flotation processes in order to obtain copper and zinc concentrates for sale to international smelters. The content of copper and zinc in the flotation residue (0.85 and 0.64%, respectively) and the environmental problems caused by its accumulation justify the development of a biohydrometallurgical process, which could be at the same time an economical viable approach for its treatment and for the minimization of its environmental impact. Bacterial leaching can be a potential treatment for this type of residue. Bacterial leaching is an environmentally safe and flexible alternative. It is nowadays well established that the most widely used bacterium – Thiobacillus ferrooxidans – is able to oxidize Fe2+, S°, as well as other reduced inorganic sulphur compounds. Two mechanisms have been established: the direct mechanism that requires physical contact between bacteria and the particles of the metal sulphide (recently it has been suggested that this mechanism should be renamed to “contact” mechanism [1]) and the indirect mechanism, according to which the bacteria oxidize ferrous ion to the ferric state, thereby regenerating the ferric ion required for chemical oxidation of the sulphide mineral. According to the new integral model for bioleaching, metal sulphides are degraded by a chemical attack of iron (III) ions and/or protons on the crystal lattice [2]. The mechanism of degradation is determined by the mineral structure: pyrite via a thiosulphate mechanism with acid production, and chalcopyrite, sphalerite and galena via a polysulfide mechanism. 75
Bioleaching Applications
The use of bioleaching for copper recovery is usually limited because chalcopyrite is very refractory to oxidation in acid media and so presents problems to bioleaching [3]. Moreover, when bioleaching is carried out in a single reactor several phenomena take place that limit to a great extent the rate of the ferrous biooxidation. Firstly, the mineral particles exert an important abrasive effect on the bacteria with a partial breakdown of them, which has two negative effects on the process kinetics: the active bacterial population decreases and the resulting organic matter would reduce or inhibit the bacterial growth [4]. The slow kinetics results in residence times of several days, even weeks, which can limit its application. In addition, the values of pH, temperature and the use of catalysts are conditioned to those values compatible with bacterial growth. Several approaches, such as the use of thermoplilic microorganisms [3] and/or the addition of catalysts [3-6] have been attempted in order to overcome these problems. Copper concentrates can be effectively bioleached by performing chemical and biological oxidation in separate steps and using silver as a catalyst in the chemical oxidation. For copper-zinc concentrates both metals can be recovered to a great extent by conducting the ferric leaching in two stages, the first one without silver to recover zinc and the second one with silver as a catalyst to extract copper [5]. However, until now this approach has not been applied to flotation tailings in which the content of both copper and zinc are low and they are finely disseminated in the sulphide matrix. The main concern relative to the application of the indirect bioleaching with chemical and biological separate steps to the treatment of flotation tailings is that the kinetics and extent of the leaching reactions and the amount of catalyst required are economically competitive with the direct bioleaching of the residue. 2.
EXPERIMENTAL
2.1 Materials The tailings used in this work were produced by the copper/zinc selective flotation plant of the Neves Corvo mine in Southern Portugal. Its chemical composition is shown in Table 1. Table 1. Chemical composition of the tailings Element
Content
Element
Content
Cu (%)
0.85
As (ppm)
6438
Zn (%)
0.64
Sb (ppm)
654
Fe (%)
22.9
Bi (ppm)
114
S (%)
23.7
Hg (ppm)
21
Pb (%)
0.21
Sn (%)
0.27
The direct bioleaching experiments were performed using a mixed culture of Thiobacillus ferrooxidans and other related bacteria isolated from Aljustrel mines drainage waters (Portugal) and routinely maintained at 34ºC on a modified Silverman and Lundgren 9K nutrient medium at pH=2.0. Chemicals were of reagent grade and all solutions were made up with distilled water.
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2.2 Procedure 2.2.1 Direct bioleaching experiments In general, bioleaching experiments were carried out in 250 cm3 conical flasks with 90 cm3 of 9K nutrient medium (without Fe(II)) at pH=2.0, 5 g of tailings (previously washed with ethanol to remove traces of flotation reagents that may inhibit bacterial growth) and 10 cm3 of inoculum. The incubated flasks were on a thermostatted bath shaker at 34ºC and at a constant speed of 280 min-1. The pH was maintained at a constant value by adding 3 M H2SO4. Suspended solids were allowed to settle and liquid samples were drawn daily for copper, zinc and total iron analysis. pH and redox potential were measured by a Pt electrode and an Ag/AgCl electrode as reference. Experiments with an ethanol solution containing 2% (v/v) of thymol instead of inoculum were used as controls. The effect of variables such as pulp density, air supply and particles size was investigated. A stirred tank bioleaching experiment was also carried out in a one dm3 glass jacketed vessel. One dm3 of the bioleaching solution was placed in the reactor vessel and heated to the working temperature (34ºC). The experiment was initiated by the addition of 50 g of dried solid. The agitation was obtained with a mechanical stirrer at 500 rpm. Air supply was provided (100 dm3.min-1). 2.2.2 Ferric sulfate leaching experiments Experiments were carried out in 250 cm3 conical flasks with 150 cm3 of ferric sulfate solution. The flasks were continuously agitated at 280 min-1 on an orbital shaker supplied with a forced air circulation thermostat. Ferric sulfate solutions were first heated to the desired temperature and the reaction was initiated by adding a dried mineral sample. In all tests, the water losses due to evaporation were taken into account during recovery calculations. At the end of the experiment, the slurry was filtered using 0.45 µm Millipore filters and the residue was washed with distilled water, dried and stored in a desiccator. The leach liquor was analyzed for copper, zinc, total iron and ferrous iron. In catalytic tests, the leaching medium consisted of ferric sulfate solutions with silver. An aliquot of a solution of silver sulfate in aqueous sulfuric acid at pH 1.40 containing 300 ppm of silver was added to the ferric sulfate solution. The amount of catalyst is expressed as milligrams of Ag+/gram of concentrate. Unless otherwise stated, the experimental conditions were: initial pH of solution 1.40, ferric iron concentration 12 g.dm-3 and duration of the test 8 hours [7]. Because all the experiments were carried out at low ferric iron concentration (12 g.dm-3) in batch systems, the studied pulp solids concentration have to be low. In a continuous operation the pulp solids concentration of leaching might be higher than the values considered in this study. The effect of the variables such as ferric iron concentration, temperature and amount of catalyst was investigated. Copper and zinc soluble in acid media were determined by performing a test in sulphuric medium at pH 1.40 without ferric iron. Copper, zinc and iron in leaching and bioleaching solutions were analysed by flame atomic absorption spectroscopy (AAS). Ferrous iron concentration was determined by standard potassium dichromate solution in an automatic titrator.
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3.
RESULTS AND DISCUSSION Pyrite was identified by X-ray diffraction (XRD) as the predominant mineral in the solid material. Considering that the residue is coming from the flotation of chalcopyrite and sphalerite bearing ores, the presence of those minerals, although in much lower contents (and therefore not identified by XRD), should be taken into account. The presence of some oxides should be considered as well. 3.1 Direct bioleaching experiments Bioleaching tests were performed in order to investigate the possibility of using a biohydrometallurgical treatment for the recovery of copper and zinc from the flotation tailing. Therefore, preliminary studies were carried out taking into account the influence of some relevant experimental parameters on metals extraction. 3.1.1 Influence of pulp solids concentration Pulp solid concentration is an important parameter that must be considered in bioleaching. Thus, in order to check its influence, bioleaching experiments were carried out in 250-cm3 conical flasks using 5, 10, 15 and 20 g of solid material in 100-cm3 solution. The evolution of the redox potential in the bioleaching solution with 5% (w/v) pulp solids concentration as a function of time denotes a considerable increase which is consistent with the ferrous ion oxidation to ferric ion and thus, it confirms bacterial growth in the presence of the tailing. The amounts of copper and zinc recovered as a function of pulp solids concentrations are shown in Figure 1.
Figure 1. Effect of pulp solids concentration on copper and zinc extraction. Conditions: 34ºC, pH=2.0
This set of experiments shows that copper and zinc recoveries depend on the pulp solids concentration. A fast initial bioleaching rate is observed in all cases, after which it becomes very slow. The best copper and zinc recoveries, 36% and 77%, were achieved with the lower solids concentration. Although the production capacity increases with the use of high solids concentrations, when the solid concentration exceeds a value between 5 78
Bioleaching Applications
to 10%, the rate of metals leaching decreases probably due to important shearing stresses and lower dissolved oxygen concentration. In addition, when high solids concentration are used, the concentration of metals is high which may inhibit bacterial growth. In fact, experiments carried out with solids concentration higher than 5% showed a slight increase in the redox potential probably due to lower bacteria activity, which lead to lower metals recovery. Iron concentration also increased with time. At the end of the experiment performed with 5% (w/v) about 60% of iron was extracted probably from chalcopyrite and pyrite dissolution. The iron recovered in the rest of conditions varied considerably and was always lower than 30%. In the control experiments copper and zinc recoveries slightly increased with time and varied from 18% to 22% for copper and from 25% to 31% for zinc. These results show the important role of bacteria in the dissolution of those metals. In order to improve metals recovery other experiments with air supply and with previous grinding were performed. 3.1.2 Influence of air supply Figure 2 compares copper and zinc extraction in the absence and in the presence of air supply (50 dm3.min-1). Air supply does not have considerable influence on metals extraction: a slight increase was observed for copper and a small negative effect was detected for zinc.
Figure 2. Effect of air supply on copper and zinc extraction. Conditions: solids concentration 5% (w/v), 34ºC, pH=2.0 3.1.3 Influence of particle size The tailing was submitted to a granulometric separation after grinding. Five different fractions were separated: < 0.106 mm, 0.106-0.200 mm, 0.200-0.500 mm, 0.500-1.00 mm, 1.00-2.00 mm. The amount of copper, zinc and iron in each fraction was determined and the corresponding values were used for metals recovery calculations (Figure 3). As expected, grinding has a positive influence on zinc recovery. More than 95% of zinc was extracted in all experiments regardless of the particle size. On the other hand, copper recovery does not seem to be affected by grinding. The results also suggest that grinding has a positive effect on the rate of metals recovery, particularly for zinc since more than 95% of metal was dissolved in the first 250 hours from the fractions with particle size lower than 0.5 mm. 79
Bioleaching Applications
50
Zn extraction (%)
Cu extraction (%)
100
40 30 20 10
80
[1.0-2.0]mm [0.5-1.0]mm
60
[0.2-0.5]mm [0.1-0.2]mm
40
< 0.1 mm
20 0
0 0
250
500
750
0
1000
250
500
750
1000
Time (h)
Time (h)
Figure 3. Effect of particle size on copper and zinc extraction. Conditions: pulp solids concentration 5% (w/v), 34ºC, pH=2.0 3.1.4 Stirred tank bioleaching tests Considering the usual limitations of the shake flask technique (mainly because of continuously changing conditions which lead to a difficult control of experimental variables) stirred tank bioleaching experiments in similar conditions were carried out. Copper and zinc extractions are presented in Figure 4. A more efficient aeration and a complete mixing of suspended solids lead to a higher zinc recovery: 92% at 500 hours against only 77% at 623 hours in the shake flask experiment. An improvement of the copper leaching rate was also observed, but only a small increase in its maximum extraction (42% against 35%) was detected.
Extraction (%)
100 80 60
Cu
40
Zn
20 0 0
500
1000
Tim e (h)
Figure 4. Copper and zinc extraction (stirred tank bioleaching). Conditions: pulp solids concentration 5% (w/v), 34ºC, pH=2.0
In these tests the amount of iron recovered was 30% at 400 hours. After that time a decrease in iron concentration was observed probably due to jarosite precipitation which usually takes place in bioleaching of sulphide materials [8]. The high zinc recoveries obtained in the bioleaching experiments were probably due to the presence of chalcopyrite and pyrite. Sphalerite chemical and biological dissolution is reported to be improved in the presence of chalcopyrite and pyrite [9], due to galvanic interactions. Contrarily, the low results of copper dissolution obtained in all tests are probably due to incomplete oxidation of chalcopyrite by bacteria (bioleaching rates of Cuoxides are reported to be faster [10]). This phenomenon is well known and has been attributed to the formation of elemental sulphur on the mineral surface. 80
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3.2 Ferric sulphate leaching experiments The leaching of the flotation tailing was initially studied at 70ºC at two different pH values, 1.25 and 1.40. Results from these tests are shown in Figure 5 in which copper and zinc soluble in both acid media (without ferric iron) are also shown. Both copper and zinc soluble in acid medium are very high, 26.5% to 28.0% of copper and 45.4% of zinc. The presence of ferric iron does not influence copper extraction, even it decreases slightly, and has a noticeable effect on zinc extraction, that markedly increases. Figure 6 shows the ferrous iron concentration and the final pH of the leaching liquor. In the absence of ferric iron there is a net consumption of acid, as the pH increases, and a small ferrous iron production (0.76 and 0.68 g/dm3 for initial pH values of 1.25 and 1.40 respectively). The presence of 12 g/dm3 of Fe (III) leads to a net production of acid, as the pH decreases, and to a high production of ferrous iron (6.11 and 6.03 g/dm3 for initial pH values of 1.25 and 1.40 respectively). These results, together with those shown Figure 5, indicate that an acid solution with 12 g/dm3 of ferric as ferric sulphate is not able to dissolve the non-acidsoluble copper of the flotation tailing (the majority being as chalcopyrite and other copper sulfides) and it is effective for the dissolution of the sphalerite present. The high ferrous iron production together with the net production of acid in ferric leaching suggests that some reductant component whose dissolution produces acid is being dissolved. This component could be pyrite. 100 90.9
90
82.8
80
Extraction (%)
70 %Cu, pHo=1,25
60
%Cu, pHo=1,40
50
45.4
45.4
%Zn, pHo=1,25 %Zn, pHo=1,40
40 30
28.0
25.9
26.5
24.5
20 10 0 [Fe(III)]=0g/L
[Fe(III)]=12g/L
Figure 5. Effect of the pH on the copper and zinc extraction with and without ferric iron. Conditions: pulp solids concentration 2% (w/v), 70°C, 8 hours 7 6
[Fe(II)] (g/L)
5 4
Fe(II) pHf
3 2
1,52
1,83 1,22
1,33
1 0
[Fe(III)]=0g/L pHo=1,25
[Fe(III)]=0g/L pHo=1,40
[Fe(III)]=12g/L pHo=1,25
[Fe(III)]=12g/L pHo=1,40
Figure 6. Ferrous iron production and final pH in tests with and without ferric iron. Conditions: pulp solids concentration 2%(w/v), 70°C, initial pH 1.25 or 1.40, 8 hours 81
Bioleaching Applications
3.2.1 Effect of the ferric iron concentration The effect of the ferric iron concentration on the copper and zinc extractions was studied over the range 8-12 g/dm3 at 70ºC. The leaching results, shown in Figure 7, indicate that an increase of the ferric iron concentration in this range does not affect the copper extraction and has a positive effect on zinc extraction. 100 90 80 Extraction (%)
70 60
%Cu
50
%Zn
40 30 20 10 0 0
2
4
6
8
10
12
[Fe(III)] (g/L)
Figure 7. Effect of the initial ferric iron concentration on the copper and zinc extraction. Conditions: pulp solids concentration 2% (w/v), pH 1.40, 70°C, 8 hours 3.2.2 Effect of the pulp solids concentration The effect of the pulp solids concentration on the copper and zinc extractions was studied over the range 2% to 8% (w/v) at 70ºC. Figure 8 shows that as the pulp solids concentration increases both copper and zinc extractions decrease. This effect is not very important in the studied range of pulp density. The increase of the ferrous iron concentration as the pulp solids concentration increases suggests that the decrease in metal extraction observed at high pulp solids concentration could be due to the depletion of ferric iron in those conditions. 100
12
90 10
70
8
60 50
6
40 4
30 20
[Fe(II)] (g/L)
Extraction (%)
80
%Cu %Zn [Fe(II)]
2
10 0
0 0
1
2
3
4
5
6
7
8
9
Pulp solids concentration (w/v %)
Figure 8. Effect of the pulp solids concentration on the copper and zinc extraction and on the ferrous iron production. Conditions: pH 1.40, 70°C, 12 g/L Fe3+, 8 hours 82
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3.2.3 Effect of the amount of catalyst The effect of the amount of silver on the copper and zinc extraction was studied over the range 0-1 mg silver/g of flotation residue at 70ºC and 2% (w/v) of pulp solids concentration. The leaching results, shown in Figure 9, indicate a marked effect of the presence of catalyst on the copper extraction. There is a noticeable increase in copper extraction as silver amount increases from 0.2 to 1 mg/g of flotation residue. Zinc extraction decreases as silver amount increase. These results are in agreement with previous results about the catalytic effect of silver on copper and zinc extraction from a mineral that contains both metals as sulphides [4]. 100 90
Extraction (%)
80 70 60
% Cu
50
% Zn
40 30 20 10 0 0
0,1
0,2
0,3
0,4
0,5
0,6
0,7
0,8
0,9
1
Ag(I) (mg Ag/g residue)
Figure 9. Effect of the amount of catalyst on the copper and zinc extraction. Conditions: pulp solids concentration 2% (w/v), pH 1.40, 70°C, 12 g/L Fe3+, 8 hours 3.2.4 Effect of temperature Figure 10 shows copper and zinc extractions in the absence and in the presence of ferric iron from tests carried out at 34ºC and 70ºC. As it was observed at 70ºC, the presence of ferric iron does not have influence on the copper extraction at 34ºC and has a positive influence on the zinc extraction. Temperature of 34ºC was chosen since it is the temperature of the direct bioleaching experiments and with the aim of comparing the efficiency of chemical and bacterial leaching. The increase of temperature from 34ºC to 70ºC influence both copper and zinc extraction as it can be observed in Figure 10. 100
90.9
90 80
Extraction (%)
70 %Cu 70°C
60
%Cu 34°C
50
45.4
%Zn 70°C %Zn 34°C
40 30 20
26.5
26.0
25.9 19,7 14.0
14.5
10 0 [Fe(III)]=0g/L
[Fe(III)]=12g/L
Figure 10. Effect of the temperature on the copper and zinc extraction with and without ferric iron. Conditions: pulp solids concentration 2% (w/v), pH 1.40, 8 hours 83
Bioleaching Applications
4.
CONCLUSIONS More than 90% of zinc can be recovered by both chemical and biological leaching of the flotation tailing, while only 40% of copper can be recovered by direct bioleaching and 20 to 30% by ferric sulfate leaching. The whole experimental conditions needed to reach those recoveries either by chemical or biological leaching (temperature, pulp solids concentration, leaching time, reagents consumption) should be further evaluated in order to establish the most favourable process. In general, the obtained results are in agreement with previous studies showing that sphalerite is more easily dissolved than chalcopyrite. However, other bioleaching tests could be performed to extend this investigation to the effect of other relevant experimental parameters (i.e. amount of inoculum, initial Fe(II) concentration, pH), which can eventually have a positive influence on copper recovery. Other approaches, such as the use of catalysts, extreme thermophiles or regrinding of the leach residue, should be considered carefully, taking into account the low copper and zinc grade of this type of material. REFERENCES
1. H. Tributsch, In Biohydrometallurgy and the environment towards the mining of the 21st century. Elsevier (1999) 51. 2. W. Sand, T. Gehrke, P.G. Jozsa, A. Schippers. In Biohydrometallurgy and the environment towards the mining of the 21st century. Elsevier (1999) 27. 3. E. Gómez, A. Ballester, M.L. Blásquez, F. González, Hydrometallurgy, 51 (1999) 37. 4. F. Carranza, I. Palencia, R. Romero, Hydrometallurgy, 44 (1997) 29. 5. R. Romero, I. Palencia, F. Carranza, Hydrometallurgy, 49 (1998) 75. 6. F. Carranza, N. Iglesias, Min. Eng., 11 (4) (1998) 385. 7. Palencia, I., Romero, R., Carranza, F., Hydrometallurgy, 48 (1998) 101. 8. J. Frenay, X. Ciechanowski, Cours de Biometallurgie, Dep. de Métallurgie et Traitment des Minérais, Université de Liége, 1996. 9. A.P. Mehta, L.E. Murr, Biotech. Bioeng. 24 (1982) 919 10. H.T. Olli, In Microbial mineral recovery. Mc Graw Hill (1990) 67.
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Bacterial tank leaching of zinc from flotation tailings V.V. Panina*, E.V. Adamova, L.N. Krylovaa, T.A. Pivovarovab, D.Yu. Voronina, G.I. Karavaikob a
Moscow State Institute of Steel and Alloys (Technological University), Leninsky Prospect, 4, Moscow, 119049, Russia b Institute of Microbiology, Russian Academy of Sciences, Prospect 60-let Octyabrya, 7/2, Moscow, 117312, Russia
Abstract The parameters of tank bacterial-chemical leaching of stale flotation tailings containing 5.6% Zn, 12.96% S and 11.6% Fe were investigated. The particle size of the tailings was 65% minus 44 µm. The main minerals were pyrite (25-30%), sphalerite (78%), pyrrhotite (7-10%), marcasite, chalcopyrite, galena (5-7%) etc. The gangue rock (55%) was predominantly present in the silicate form. Zinc was leached under continuous conditions at 28-30°C in agitated tanks with a pulp density of 16.7, 28.6 and 40.0% of solids. Acidithiobacillus ferrooxidans strain TFI was isolated and cultivated on zinc flotation tailings and adapted to zinc ions concentrations in the liquid phase of up to 40 g/L. Zinc extraction by leaching at a pulp density of 40% of solids with the return of 10% of the solution from the last tank reached 87.12% while the concentration of zinc ions in the leach solution was in the range of 31.4-32.4 g/L. The solids throughput was increased 8 fold, pulp flow rate 2.4 fold and overall leaching residence time was reduced 2.7 fold as compared to leaching at a pulp density of 16.7% of solids. Keywords: bioleaching, sphalerite, flotation tailings, acidithiobacillus, pulp density, process flowsheet 1.
INTRODUCTION Huge amounts of technogenic resources including flotation tailings are being accumulated worldwide in the areas adjacent to existent and closed mineral processing plants. The higher grades of ores processed earlier together with imperfection of the technology used are the main reasons why the stale flotation tailings are relatively rich with valuable components. In certain cases they are comparable by tenor with some of the ores mined nowadays and therefore can be reprocessed to extract metals. Generally the use of expensive pyrometallurgical and pressure leaching technologies to extract valuable components from low-grade technogenic resources is not economically viable. In such cases the technologies with lower operating costs come to the foreground. Primarily these are the cost-effective hydrometallurgical technologies and among them bacterial-chemical leaching should be considered [1-4]. *
Corresponding author: Tel./Fax: + 7-095-951-21-39, E-mail address:
[email protected] (V.V. Panin)
85
Bioleaching Applications
The aim of this work was to investigate and develop the technological parameters of bacterial-chemical leaching of zinc from the flotation tailings. These parameters include solids content, specific pulp flow rate and flowsheet configuration. 2.
MATERIALS AND METHODS Quantitative chemical analysis showed that the stale zinc-containing flotation tailings with a particle size of 65% minus 44 µm contained 5.6% Zn, 12.96% S and 11.6% Fe. The gangue rock (55%) was predominantly present as silica – 32-38%, mica (biotite, muscovite) – 10-12%, graphite – 5-7% and Fe silicates. The main minerals were pyrite (25-30%), sphalerite (7-8%), pyrrhotite (7-10%), marcasite, chalcopyrite, galena (5-7%), sulphosalts, Cu sulfate, secondary Fe minerals, magnetite and haematite (partially substituted by Fe hydroxides). More than 70% of valuable minerals were intergrown with other valuable and gangue minerals. Zinc in flotation tailings was present both in the sulfide form (sphalerite) and in the oxide form (sulfates, oxides, silicates and ferrites). The major amount of zinc was found to occur in particle size range of 20-44 µm. Bacterial-chemical leaching of zinc flotation tailings was performed under continuous conditions at 28-30°C using three agitated tanks connected in a series. The volume of each tank was 1600ml with the mixer speed of 400 rpm. Aeration rate during the experiments was 1L/min per 1L of pulp volume. Acidithiobacillus ferrooxidans strain TFI was routinely isolated and cultivated on zinc flotation tailings as the energy substrate. Adaptation efforts resulted in obtaining a strain tolerant to 40 g/L of zinc ions in solution. The cell concentration was evaluated by both Lowry protein measurement [5] and end-point tenfold dilutions method. Strain identification in liquid phase of leaching pulp was performed by analyzing chromosomal DNA structure using pulsed field gel electrophoresis method [6]. Cells activity in leaching pulp and solutions was controlled by manometric method (O2 consumption) and by measuring the rate of ferrous iron oxidation to ferric iron. Total iron concentration (ferrous plus ferric iron) in solution, as well as zinc concentration in solution and in leach residue (after acid decomposition), were determined by atomic absorption spectrophotometry (Perkin-Elmer mod. 3100). Separately, ferric and ferrous iron concentrations in solution were determined by complexometric titration. Redox potentials were measured using a platinum electrode (combined with a silver/silver chloride reference electrode) and converted to Eh values (relative to a hydrogen reference electrode). 3.
RESULTS It is common practice to leach and oxidize sulfide concentrates in the presence of bacteria at a pulp density of 16-20% of solids. In the present work zinc was bacterially leached at pulp densities of 16.7, 28.6 and 40% of solids. Experimental results are reported in Table 1. The cells concentration and activity (measured as O2 consumption) together with ferric iron and zinc concentrations in leach solution increased with an increase in the pulp density. Experimental mass balance of bacterial-chemical leaching of zinc from the stale flotation tailings at different pulp densities is shown in Table 2. The dynamics of zinc leaching at different pulp densities are shown in Fig. 1. High zinc concentrations in leach
86
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solutions were achieved by increasing the pulp density; zinc extraction was still at a high level (87.12%). Zinc content in solid leach residue was reduced from 5.6% to 0.75-0.87%. Table 1. Technological parameters of bacterial-chemical leaching of zinc from the flotation tailings at different solids content, % Factors Leach solution pH value Leach solution Eh value (mV) Ferric iron concentration (g/L) Zinc concentration (g/L) Cells respiratory activity (µL O2/mL⋅hour) Cells concentration (cells/mL) Flow rate (tank volume/hour) Pulp throughput (ml/day) Solids throughput (g/day) Single tank residence time (hour)
16.7% 1.5-2.0 650-750 5.5-6.0 8.7-10.0 20.0-44.0 108-109 0.013 500 100 96
Values 28.6% 1.4-2.0 700-750 10.0-14.0 13.2-19.0 36.0-40.0 109-1010 0.031 1200 480 40
40.0% 1.3-2.0 700-750 13.6-15.5 31.4-32.4 40.0-56.0 1010-1011 0.031 1200 804 35
Table 2. Mass balance of bacterial-chemical leaching of zinc from the flotation tailings at different solids content, % Tank
Leach residue yield (%)
Zinc concentration in solution (g/L)
Zinc content in leach residue (%)
Zinc extraction (%)
1.40 1.20 0.75 0.75
77.50 81.80 88.90 89.30
2.52 1.51 1.01
58.96 76.97 85.01
Pulp density 16.7% solids Feed preparation tank Leaching tank 1 Leaching tank 2 Leaching tank 3
90.0 85.0 83.0 80.0
8.68 9.16 9.96 10.00 Pulp density 28.6% solids
Feed preparation tank Leaching tank 1 Leaching tank 2 Leaching tank 3
95.0 90.0 86.0 83.0
13.21 17.40 19.04 Pulp density 40.0% solids
Feed preparation tank Leaching tank 1 Leaching tank 2 Leaching tank 3
95.0 89.8 85.7 82.8
31.36 32.39 32.37
1.02 0.89 0.87
83.64 86.38 87.12
Feed (flotation tailings)
100.0
⎯
5.6
100.0
High redox potentials and low pH values of the solution were observed during bacterial-chemical oxidation. At a pulp density of 16.7% of solids zinc extraction in the first tank was significantly higher than at a pulp density of 28.6% due to the extended residence time (96 hours vs. 40 hours). Increasing the pulp density to 28.6% with higher flow rate led to a decrease of zinc extraction in the first two tanks but in third tank extraction reached 85%. At a pulp density 87
Bioleaching Applications
of 40.0% of solids high zinc extraction and high flow rates in all tanks were achieved by returning 10% of the leach solution from the last tank to the first one. Returned liquid phase was characterized by high concentrations of active cells and ferric iron. It is possible at higher pulp densities to increase the flow rates and hence the overall leaching process throughput while keeping fast enough oxidation. Leaching at a pulp density of 40.0% of solids increased the solids throughput 8 fold, pulp flow rate 2.4 fold and decreased overall leaching residence time 2.7 fold as compared to the pulp density of 16.7% of solids. Mineralogical analysis of flotation tailings leach residues revealed significant phase transformations and the presence of coarse aggregates of up to 300 µm in size cemented by newly formed substance. Inside these aggregates relic mineral inclusions were found: sphalerite grains (10-15µm), pyrite, graphite and gangue. Only gangue minerals, undestructed pyrite and secondary Fe minerals were observed in "free" form. No "free" sphalerite grains were found in leach residue.
Figure 1. The dynamics of zinc leaching from flotation tailings at different pulp densities
The substance cementing the mineral aggregates was formed by incompletely destructed gangue rock fragments, whity-yellow and gray colored matter (elemental sulfur and bacterial metabolites) and unidentified porous material with sub-micron intermetallic inclusions. Mineralogical data on initial flotation tailings and on bacterial-chemical leach residues correlated the results of chemical analyses and gave evidence that "free" zinc sulfide grains and some other sulfide minerals were leached almost completely. Sphalerite bacterial-chemical oxidation mechanism can be described according to Fowler and Crundwell [7-9] by the following reaction:
ZnS + Fe2 ( SO4 ) 3 → ZnSO4 + 2 FeSO4 + S 0
(1) Iron oxidizing acidophilic bacteria regenerate ferric iron and oxidize elemental sulfur:
4 Fe 2+ + 4 H + + O2 → 4 Fe 3+ + 2 H 2 O
(2)
S 0 + H 2 O + 1.5O2 → H 2 SO4
(3)
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Bioleaching Applications
As a result of the reaction according to Eqn. 3, the probability of sphalerite passivation by elemental sulfur layer is reduced greatly; the addition of sulfuric acid is not necessary because the latter is generated as a product of sulfur oxidation process. Surface-active properties of bacterial metabolites in leaching pulps have a technological significance since they increase the solids settling rate during thickening and dewatering. Since iron is present in bacterial-chemical leach solutions in the oxidized ferric form the lime consumption for Fe hydroxide precipitation prior to zinc extraction from the solution will be lower. 4.
CONCLUSIONS It has been shown in this work that unlike the common practice of leaching and oxidizing sulfide concentrates in the presence of bacteria at pulp densities of 16-20% of solids it is possible to bacterially leach zinc at a pulp density of 40% of solids with high technological results. Raising the pulp density in tank bacterial-chemical zinc leaching up to 40% of solids with the return of 10% solution from the last tank increased solids throughput 8 fold, pulp flow rate 2.4 fold and allows to reduce 2.7 fold the overall leaching residence time as compared to the leaching at a pulp density of 16.7% of solids. ACKNOWLEDGEMENTS This research was carried out with the support of Federal Scientific and Technical Program "Investigations and Developments on Priority Directions of Science and Technology" (years 2002-2006) on the subject "Biotechnology in mining and processing of mineral resources". REFERENCES
1. Sheveleva L.D., Abakumov V.V., Korkin B.I., Bishev L.Z. and Karavaiko G.I., 1995. Development of new technology for reprocessing of concentrating mill tailings. Tsvetnye Metally, 12, 23-26. 2. Panin V.V., Adamov E.V., Karavaiko G.I., Khamidullina F.G. and Voronin D.Yu., 1999. Use of the Bacterial Leaching Technology in Processing of Refractory CopperZinc Ores. Tsvetnye Metally, 5, 9-11. 3. Karavaiko G.I., Sedelnikova G.V., Aslanukov R.Ya., Savari E.E., Panin V.V. and Adamov E.V., 2000. Biohydrometallurgy of Gold and Silver. Tsvetnye Metally, 8, 2026. 4. Pol’kin S.I., Adamov E.V., Panin V.V. Technology of Bacterial Leaching of NonFerrous and Rare Metals. Nedra, Moscow, 1982, 288 p. 5. Lowry O.H., Rosenbrough N.J., Farr A.L. and Randell R.J., 1951. Protein Measurement with the Folin Phenol Reagent. Journal of Biological Chemistry, 193, 265-275 6. Kondratyeva T.F., Muntyan L.N. and Karavaiko G.I., 1995. Zinc- and ArsenicResistant Strains of Thiobacillus ferrooxidans have Increased Copy Numbers of Chromosomal Resistance Genes. Microbiology, 141 (5), 1157-1162. 7. Fowler T.A. and Crundwell F.K., 1998. Leaching of Zinc Sulfide by Thiobacillus ferrooxidans: Experiments with a Controlled Redox Potential Indicate No Direct Bacterial Mechanism. Applied and Environmental Microbiology, 64 (10), 3570-3575. 8. Fowler T.A. and Crundwell F.K., 1999. Leaching of Zinc Sulfide by Thiobacillus ferrooxidans: Bacterial Oxidation of the Sulfur Product Layer Increases the Rate of 89
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Zinc Sulfide Dissolution at High Concentrations of Ferrous Ions. Applied and Environmental Microbiology, 65 (12), 5285-5292. 9. Driessens Y.P.M., Fowler T.A. and Crundwell F.K. A comparison of the bacterial and chemical leaching of sphalerite at the same solution conditions. In: Biohydrometallurgy and the Environment toward the Mining of the 21st century. R Amils and A. Ballester (eds.) Elsevier, 1999. Part A, pp.201-208.
90
15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Behaviour of elemental sulphur in the biohydrometallurgical processing of refractory gold-sulfide concentrates of various mineral types Sedelnikova G.V.*, Savari E.E. Central Research Institute of Geological Prospecting for Base and Precious Metals (TsNIGRI), Varshavskoe sh. 129 b, 117545 Moscow, Russia Abstract The kinetics of oxidation of gold-sulfide concentrates of various mineral types with the use of Acidithiobacillus ferrooxidans monoculture and mixed culture of mesophilic bacteria has been studied. It is shown that the oxidation of the main sulfide minerals arsenopyrite, pyrite and pyrrhotite is more efficient when applying mixed cultures. The oxidation of sulfide concentrates is accompanied by release of elemental sulfur. Most amount of elemental sulfur is derived in the course of bacterial oxidation of pyrrhotitepyrite-arsenopyrite concentrates. With actually complete oxidation of pyrrhotite, arsenopyrite and some pyrite, the process of elemental sulfur biooxidation is not terminated. Even additional bacterial leaching of pyrrhotite concentrate within 7 days does not completely oxidize elemental sulfur. As elemental sulfur has a negative effect on the process of gold recovery by cyanidation of biooxidation residues and leads to high consumption of sodium cyanide, some ways of additional oxidation of elemental sulfur by aeration in lime environment or electrolytic treatment were investigated. The research outcomes were taken into account in design and construction of the first commercial plant in Russia at the Olimpiada gold deposit.
Keywords: gold-sulfide concentrates, biohydrometallurgy, elemental sulfur 1.
INTRODUCTION Since 1986, the bacterial leaching of refractory gold ore and concentrates has been employed for gold recovery at a commercial scale. At present, more than 10 commercial plants are operating in the world: in Australia, Ghana, South Africa, Brazil, Peru, China etc. In Russia, the first plant was commissioned in 1997 at the Olimpiada gold deposit, the Krasnoyarsk territory, its daily capacity being 1 t concentrate with further increasing up to 400 t/day in 2001. The bacterial leaching is carried out around the year under severe climatic conditions, at winter temperatures of -25 to -45°C.
* Corresponding author: Fax/ Phone (095) 113-68-22, E-mail:
[email protected]
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The known BIOX® technology developed by Gencor, first tested at Fairview and later on applied at the other plants uses a mixed culture of mesophilic bacteria and biooxidation plants are operating within the range of 40° to 45°C. The other technology - Bac Tech process employs moderately thermophiliс bacteria at about 50º C at the Youanmi commercial plant in Australia [1]. Taking into account cold climate and refractory ore reserves in Russia which are to be developed, this paper considers the process of biooxidation of refractory gold-sulfide concentrates of various mineral types with the use of mesophilic bacteria at 28-32°C. While elaborating technology of treating the pyrrhotite - pyrite - arsenopyrite concentrates from the Olimpiada deposit, we faced some problems caused by release of a great amount of elemental sulfur as a result of biochemical oxidation of sulfides and first of all of pyrrhotite. It is known that S° [2] and S2, S2O32-, SO32- sulfur-bearing anions [3-6] produce (CSN)- and cause high cyanide consumption in the course of cyanidation of biooxidation residues. At the plants using biooxidation technology for the treatment of refractory gold concentrate the cyanide consumption attains 30 kg/t of concentrate [3]. Therefore, several methods of reducing cyanide consumption are applied, such as: increase of pulp density [4, 6], pre-leaching lime aeration [7, 8], pre-treatment in sodium hydroxide solution [9], electrolytic treatment [10]. This paper describes the results of study of the elemental sulfur behavior in the process of bacterial oxidation of various mineral types of concentrates and recommends pre-aeration in the lime environment and electrolytic treatment of pulp prior to cyanidation in order to reduce cyanide consumption and to increase gold recovery from the biooxidation residues. 2.
MATERIALS AND METHODS
2.1 Sulfide concentrates Concentrate samples were produced by mineral processing of refractory gold ore of Russian deposits. Four samples of refractory gold - sulfide concentrates were studied under laboratory conditions, such as: gravity concentrate (1) from the Nezhdaninka deposit, Republic of Sakha -Yakutia, gravity - flotation concentrate (2) from the Albazin deposit, the Khabarovsk region, and two flotation concentrates (3) and (4), accordingly, from Nezhdaninka, Republic of Sakha – Yakutia, and Olimpiada, the Krasnoyarsk region. The main gold-bearing sulfide minerals in concentrates are arsenopyrite and pyrite. The concentrate (4) contains also pyrrhotite. The quantitative predominance of one or another mineral in decreasing order allows to distinguish three main mineral types of refractory gold-sulfide concentrates: arsenopyrite - pyrite concentrates (1) and (2); pyrite arsenopyrite concentrate (3); pyrrhotite - pyrite - arsenopyrite concentrate (4). Gold grade in concentrates is 21.6 - 150 g/t, silver 2.3 - 160 g/t. Total sulfur actually occurs in the sulfide form amounting to 5.68-28.8%; the content of elemental sulfur in pyrrhotite-free concentrates (1 -3) is insignificant (0.11-0.15%) while its grade in pyrrhotite-bearing concentrate (4) is higher by an order – 1.3%. Arsenic occurs, mainly, in the form of sulfide arsenic amounting to 4.5-18.9%. The concentrates also contain nonsulfide constituents, such as (%): 9.15 – 51.6 SiO2, 0 – 8.16 CO2, 0.54 – 4.5 C organic. The mineral composition of concentrates (Table 1) is represented by various contents of the main gold-bearing sulfide minerals - arsenopyrite, pyrite and pyrrhotite as well as sulfides grading 17 to 74.8%. Pyrrhotite is abundant (39%) only in the concentrate (4); 92
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antimonite (1.7%) is also contained, mainly, in the concentrate (4). The grain size of concentrate samples is within the range of 80-90% of the class – 0.044 mm. Table 1. Mineral composition of concentrates Mineral type of concentrates, grade % Sulfides
FeS2 FeAsS FeS Sb2S3 Total sulfide content
Arsenopyrite-pyrite (1) 33.7 41.0 single grains 0.1 74.8
(2) 7.2 9.8 not found not found 17.0
Pyrite-arsenopyrite (3) 24.3 10.4 not found single grains 34.8
Pyrrhotite-pyritearsenopyrite (4) 12.0 11.8 39.0 1.7 64.5
2.2 Microorganisms The research on bacterial oxidation of the concentrate (1) employed mesophilic bacteria: the Acidithiobacillus (A.) ferrooxidans monoculture with concentration of 1,0 x 107cells/ml and mixed culture of A. ferrooxidans and A. thiooxidans with corresponding content of 2.5 x 107 and 2.5 x 104 cells/ml. (Table 2). Table 2. Caracterization of the applied culture* Mineral type of concentrates
Bacteria species, quantity of cells per milliliter of solution
Arsenopyrite-pyrite: (1)
Monoculture – A.ferrooxidans 1.0 x 1010
(1)
Mixed culture: A.ferrooxidans – 2.5 x 107, A.thiooxidans – 2.5 x 104
(2)
Mixed culture: A.ferrooxidans – 1.6 x 109, A.thiooxidans – 2.5 x 104
Pyrite-arsenopyrite (3)
Mixed culture: A.ferrooxidans – 3.5 x 109, A.thiooxidans – 5 x 103
Pyrrhotite-pyrite-arsenopyrite (4)
Mixed culture: A.ferrooxidans – 4.5 x 109, A.thiooxidans – 4.5 x 108 Leptospirillum ferrooxidans – 2.2 x 104, Sulfobacillus thermosulfudooxidans – 2.2 x 106
* The microorganism populations were studied by G.I.Karavaiko, T.F.Kondrateva and T.A.Pivovarova, specialists from the Institute of Microbiology, RAS.
The oxidation of concentrates (2, 3) was performed with the use of the mixed culture of A.ferrooxidans and A. thiooxidans bacteria with cell concentrations of 7.05x108 – 1.0x109 and 2.5x104 cells/ml correspondingly. The pyrrhotite-bearing concentrate (4) was tested with the use of the mixed culture of A. ferrooxidans – 4.4 x109, A. thiooxidans – 4.5 x108, Leptospirilllum ferrooxidans – 2.2 x104 and moderately thermophilic Sulfobacillus thermosulfidooxidans bacteria – 2.2 x106 cells/ml. Bacterial culture available in the TsNIGRI’s biotechnological laboratory which were previously adapted to pyrite-arsenopyrite substratum were used in researches. Adaptation and growth of the biomass was implemented for each particular concentrate (1-4) by transferring bacteria to the 9K environment with portioned addition of concentrates up to the ratio of S(solid):L(liquid) = 1:50-1:20. The biomass was further re-disseminated and 93
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the content of solids increased up to S:L=1:5. The process of bacterial adaptation continued with checking pH= 1.5-2.2, Eh = 500-680 mV, concentration of iron oxide and protoxide, biomass activity measured by oxygen consumption on the Warburg device and the rate of iron protoxide oxidation. 2.3 Biooxidation tests The process of bacterial oxidation of concentrates was studied on the pilot laboratory unit under the conditions of continuous bacteria cultivating. The unit includes 5 reactors of capacity 2 litres each, furnished by devices for mechanical mixing, dispersion and additional 1000 cm³ /l per min air supply. Bacterial oxidation was performed at S:L = 1:5, pH=1.5-2.2 and temperature of 28 - 32º C. The pH value was maintained by addition of sulfuric acid or lime slurry depending on the material composition of the concentrate. The required temperature of the process was provided with the help of thermoelectric heaters placed directly in the reactors. The main parameters were constantly monitored, such as: pH, t°C, the biomass oxidizing activity; bacterial solutions were analyzed for Fe+3, Fe+2 and As. Once a day, 45 ml pulp samples were taken to make analyses for solid phase – sulfide As, sulfide and elemental S and sulfide Fe. 3.
TEST RESULTS
3.1 Comparison of the efficiency of FeAsS - FeS2 concentrate bioleaching with the use of A. ferrooxidans and a mixed culture of A. ferrooxidans and A. thiooxidans In order to select the most effective microorganisms for biooxidation of refractory concentrates containing arsenopyrite and pyrite at 28-32°C, the comparative tests were made on bacterial leaching of arsenopyrite - pyrite concentrate (1) with the use of A. ferrooxidans monoculture and a mixed culture of A. ferrooxidans and A. thiooxidans. Performance of biooxidation was estimated from the analyses of the contents of sulfide and elemental S and sulfide As in the residues of biooxidation of the concentrate (1) within 5 days (Fig. 1A), and also indexes of the completeness of arsenopyrite and pyrite oxidation (Fig. 1B). The kinetic curves show higher rate of sulfide components biooxidation in case of the mixed A. ferrooxidans and A. thiooxidans culture as contrasted to the single A. ferrooxidans (curves 1 and 1°). By the end of the 5-th day, residual content of sulfide S was, accordingly, 8.12 and 10.2%, sulfide As 0.1 and 1.0% and elemental S - 0.42 and 1.41%. When using the mixed bacterial culture, more complete oxidation of arsenopyrite – 99.4% and pyrite 71.4% is attained as contrasted to the monoculture (correspondingly, 94.7% and 62%). Joint use of A. ferrooxidans and A. thiooxidans was recommended as most effective tool for further investigation of bacterial oxidation of other mineral types of sulfide concentrates containing arsenopyrite and pyrite. 3.2 Kinetics of biooxidation Fig. 1 shows kinetic curves corresponding to biooxidation of sulfide components of three mineral types of concentrates at 28-32°C with the use of mixed bacteria, the performance of the latter being shown in Table 2. The kinetic curves of bacterial oxidation of the concentrates (1-3) containing only two main minerals - pyrite and arsenopyrite differ from kinetic curves of biooxidation of the concentrate (4) which additionally contains pyrrhotite. 94
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95
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1-4 residues of biooxidation of the concentrates (1 – 4) with mixed culture of bacteria. 1° residue of biooxidation of the concentrate (1) with A.ferrooxidans.
Figure 1. Biooxidation of sulfur, sulfide arsenic, elemental sulfur, pyrite and arsenopyrite during bioleaching of gold sulfide concentrates
The sulfide minerals are known to have different crystalline structures and different stability in sulphate solutions. With pH= 2.5 a row of sulfide stability is as follows: pyrrhotite (0.45B), arsenopyrite (0.50B), pyrite (0.55 – 0.6B). Pyrrhotite, as most electrochemically sensitive mineral, is oxidized, first of all, chemically and microbiologically according to the reactions: (1) FeS + Fe2(SO4)3 = 3 FeSO4 + S° (2) FeS + 0,5 O2 + H2SO4 = FeSO4 + S° + H2O bacteria (3) 2FeS + 4,5 O2 + H2SO4 = Fe 2(SO4)3 + 2 H3AsO4 In the course of biochemical oxidation of pyrrhotite Fe2+, Fe3+ sulfates and S° are created, which are further oxidized by bacteria: Fe2+up to Fe3+, S° up to SO4. In the first two days of concentrate biooxidation a considerable growth of S° grade in the residues of biooxidation of pyrrhotite-bearing concentrate (4) is observed - from 1.3 up to 6.85% while in the course of oxidation of the pyrrhotite-free concentrates (1-3) less amount of S° is produced – 0.1-1.3%. This testifies about predominant oxidation of pyrrhotite (90-95%) and lower oxidation rate of arsenopyrite (19-29%) and, especially, pyrite when leaching the FeS - FeS2 - FeAsS concentrate (4). In all other pyrrhotite-free concentrates, the arsenopyrite oxidation rate in the first two days is 2.5 - 3 times higher and it is oxidized for 78-91%. The lower rate of arsenopyrite oxidation in pyrrhotite-bearing concentrate as compared to pyrrhotite-free concentrates is, probably, explained by variations in electrochemical characteristics of the environment: increase of pH from 1.8 up to 2.2 and decrease of Eh value from 700 up to 500 mV due to Fe2+ accumulation and reduced 96
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soluble sulfur compounds resulted from chemical oxidation of pyrrhotite and, partially, arsenopyrite that is evidenced by diminishing concentration of A.ferrooxidans bacteria from 2.5 x 107 up to 2.5 x 104 cells / ml in the first two days of oxidation of the pyrrhotitebearing concentrate (4). Then S° content decreases from 6.8 up to 3.9% due to microbiological oxidation by A. Thiooxidans bacteria the amount of which increases from 2.5 x 107 to 2.5 x 109 cells / ml. The Sulfacillus thermosulfidooxidans bacteria living inside the system also facilitate S° oxidation. At the same time, the rate of arsenopyrite oxidation increases that is evidenced by diminishing concentration of sulfide As from 4.4 up to 0.2-0.4%, accordingly, after 2 and 5 days of bioleaching, and by growing quantity of A.ferrooxidans bacteria from 2.5 x 104 up to 6 x 107 cells /ml. For 5 days of the Fe S - FeS2 - FeAsS concentrate (4) bioleaching, arsenopyrite is oxidized for 98%, pyrrhotite - 99%, pyrite 35%. The residual content of sulfide S accounts for 6.1% as compared to the initial 28.8% value. The completeness of sulfide oxidation in the FeAsS - FeS2 and FeS2 - FeAsS concentrates is governed by a total grade of sulfides and their quantitative ratio. All kinetic curves of oxidation of pyrite, arsenopyrite, sulfide S and As, and also elemental S in the concentrates (1-3) are ranked as 1, 3, 2. The content of sulfides in concentrates decreases in the same order, %: (1) – 74.8, (3) – 34.8, (2) – 17.0. Hence, the less is the content of sulfides, the more completely they are oxidized. This dependence is most pronounced for pyrite as an example. For 5 days, the highest degree of pyrite biooxidation (94.8%) is attained in the concentrate (2) which contains the least amount of sulfides - 17% (7.2 FeS2). The worst oxidation of pyrite (71.1%) occurs in the concentrate (1) that contains the greatest quantity of sulfides - 74.8% (33.7% FeS2). The duration of bacterial leaching of refractory concentrates is governed by their composition and, first of all, by sulfide content. The concentrate (2) containing the least quantity of sulfides (17%) is oxidized almost completely within 3-4 days. The oxidation of the concentrates (1,3,4) containing 34.8-74.8% sulfides requires 5-6 days of bioleaching with pH = 1.5-2.2, S:L = 1:5, t = 28-32ºC with the use of mixed culture of mesophilic bacteria. The analysis of kinetic curves of Sº biooxidation demonstrates that the process of Sº oxidation is not terminated after 5 days of concentrate leaching, except for the concentrate (2) with the lowest sulfide grade (17%). A great amount of Sº (3.9%) is retained in the pyrrhotite-bearing concentrate (4) while 0.42 and 0.24% Sº remains in the concentrates (1) and (3). For microbiologic oxidation of Sº the leaching was prolonged from 5 to 12 days (Fig. 1). This allowed to decrease the content of elemental sulfur from 3.9 only to 1.5% while the content of sulfide sulfur dropped from 6.1 up to 5%. 3.3 Oxidation of Sº by hydrometallurgy before cyanidation As a rule, the biooxidation residues have a complex composition and contain constituents which consume oxygen and cyanide, and deteriorate the process of gold cyanidation. It is shown in publication [8] that the residue of biooxidation of pyrite – arsenopyrite ore contain S2-, Fe (OH) compounds, residual FeAsS, jarosite and Sº that are the consumers of oxygen and sodium cyanide. In order to diminish their action, the pulp was treated, prior to cyanidation, with lime at рН=10.5 within 24-40 hours, then sodium cyanide was added to perform leaching within 24 hours. We studied an effect of several methods of treating the biooxidation residues before cyanidation on the decrease of the content of elemental sulfur, such as: 3-4 repeated water 97
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washing, neutralization, aeration in limy environment, electrochemical treatment. The hydraulic washing of biooxidation residues does not affect the Sº content. The neutralization just insignificantly decreases the Sº content - from 3.9 up to 3.75%. The aeration of the pulp in lime environment causes a partial oxidation of Sº - from 3.75 up to 2.71%. Most effective is the electrolytic treatment of pulp within 1-2 hours at pH=11-12 (Fig. 2A).
Figure 2. Effect of treatment of biooxidation residues before cyanidation on content of sulfur (A), gold extraction and cyanide consumption (B)
In the course of electrolysis of aqueous solutions the gaseous oxygen is released; its solubility in solution increases up to 20-30 mg/l. The gas saturation of pulp with finely dispersed bubbles increases 100-1000 times as compared to aeration. The process of Sº oxidation is enhanced and its content decreases from 3.75 up to 2.17% for 1-2 hours of treating. A mechanism of pre-aeration in lime environment or electrolytic processing can be represented by two stages: 1. Oxidation and precipitation of oxygen and cyanide absorbers prior to cyanidation: –Fe2+→ Fe3+→ Fe (OH)3, S2-→ S° → S2O3 2- → SO42- CaO→ CaSO4 2.
Lime passivation of sorption-active surface of newly precipitated iron hydroxides and decrease of a sorption capacity of products of bacterial oxidation in relation to cyanide complexes of precious metals. The applying of aeration and electrolytic processing of pulp allows to stabilize the pulp ionic composition prior to cyanidation. As the duration of preparation increases, the concentration of oxygen and cyanide absorbants decreases, mg/l: from 52.5 up to 10.2 S2-; from 35 up to 0 S2O32-; from 6.4 up to 0 Fe2+. A chemical value of oxygen absorption decreases from 480 up to 320 mg/l and remains constant during the whole cyanidation process. The stabilizing of ion composition of pulp allows to reduce 3.3 times the CNSion concentration and 1.7 times the SO42- ions. The neutralization of harmful constituents allows to improve gold dissolution and to reduce cyanide consumption. The preliminary aeration within 24 hours allows to increase gold recovery from the residue of biooxidation of the concentrate (4) from 90 up to 96% 98
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and to reduce the cyanide consumption from 16.0 up to 7.2 kg /t. The electrolytic processing of pulp before cyanidation increases gold recovery up to 97% and reduces the cyanide consumption from 16 up to 7 kg /t (Fig. 2B). Table 3 shows that biohydrometallurgical processing of refractory gold-sulfide concentrates is an effective technology of sulfide oxidation and recovery of finely disseminated precious metals. Gold recovery from the studied mineral types of concentrates with the use of biohydrometallurgical technology attains 92-98% as compared to low-effective parameters of gold recovery by cyanidation of source concentrates – 12.9-63.4%. Table 3. Effect of biooxidation on gold recovery from concentrates Mineral type of concentrate Processing method
Gold recovery by cyanidation, % Gold recovery by biohydrometallurgy, % Duration of biooxidation, days
(1) 63.4
(2) 12.9
(3) 37.5
Pyrrhotite-pyritearsenopyrite (4) 51.2
92.1 5
98.0 4
92.0 5
97.0 5
Arsenopyrite-pyrite
Pyrite-arsenipyrite
4.
CONCLUSION S° is formed during biooxidation of all mineral types of concentrates containing arsenopyrite, pyrite and pyrrhotite. The presence of pyrrhotite enables creating of the most amount of S°. The process of its biooxidation is slower than sulfides oxidation. S° is saved in biooxidation residues and has a negative effect on the process of cyanidation, leads to high consumption of cyanide because of (CNS) creating. Using of aeration in lime environment or electrochemical treatment enables to decrease cyanide consumption as much as 2.3 times and increase extraction of gold from 90 to 96-97%. REFERENCES
1. Bierly I.A., International Congress, Biotecnology, 2002 Moscou, Russia, (2002) 457. 2. Shrader, V.J. and S.X. Su, International Biohydrometallurgy Symposium '97 (1997) M3.3.1. 3. Lodeishikov V.V., Technology of gold and silver recovery from refractory ores. Irgiredmet JSC. Irkutsk, (1999). 4. Jones L. and Hacki R.P., International Biohydrometallurgy Symposium '99 (1999) 337. 5. Rossi, G., Biohydrometallurgy, McGraw-Hill, Hamburg, 1990. 6. Komnitsas C. and Pooley E.D., Minerals Engineering, 3 (1990) 295. 7. Hacki, R. P. and L.Jones, International Biohydrometallurgy Symposium '97 (1997) M14.2.1. 8. Livsesey, E., P. Norman and D.R.Livesey, International Biohydrometallurgy Symposium '83 (1983). 9. Xiang, L. and J. Ke, Transactions of Nfsos., 4(4) 1 (1994) 42. 10. Sedelnikova G.V. and A.V.Narseev, Proceedings of TsNIGRI, 233 (1989) 26.
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Beneficiation of phosphatic ores from Hirapur, India A. D. Agate 7, Narmada Apts., United Western Society, Pune - 411 052, India Abstract The availability of phosphorous in tropical soils is generally low due to immediate precipitation and fixation of applied phosphorous to soil. In India, 98% of the soil contains insufficient amounts of available phosphorous to support maximum plant growth and hence, application of phosphorous is very essential. High grade rock phosphate is in short supply. Therefore, it becomes necessary to import di-ammonium phosphate from abroad. Hence, it was thought to beneficiate the low grade rock phosphate available at Hirapur in India. Unfortunately, it contains high amounts of iron and silica (3-5%). Hence, microbial beneficiation was attempted using Thiobacillus ferrooxidans to remove extra iron by ferrous iron oxidation and by removal of both iron and silica by sieving and cleaning the ore using various physical treatments. This resulted in a cleaner ore, with nearly 70% iron and 25% silica impurity removed. The factors for such a beneficiation process were standardized at laboratory level (optimum pH = 3, temperature = 30°C, 1% pyrite added as nutrient). These factors are used for scaling up the process at field site at one t.p.d. level. This is an important first step in production of indigenous phosphate source to eliminate import - dependence in agriculture.
Keywords: phosphatic ores, Thiobacillus ferrooxidans, beneficiate 1.
INTRODUCTION The availability of phosphorus in tropical soils is generally low due to immediate precipitation and fixation of the applied phosphorus. In India, 98% of the soils contain insufficient amounts of available phosphorus to support maximum plant growth. Application of phosphatic fertilizers, therefore, becomes very essential (1, 2). According to several workers (3, 4), about 85 to 90% of the added phosphorus in the fertilizer gets fixed in the soil within 24 to 48 hours of its application and becomes unavailable to the plants. Thus, efficiency of utilization of phosphatic fertilizers in soils ranges from 10 to 15% only. The problem of phosphorus fixation was found to be very acute in neutral to alkaline soils in India, where the phosphorus is precipitated as calcium phosphate upon application of superphosphate. This results in non-availability of applied phosphorus in soils, which is a considerable loss from the point of view of productivity and economy to the nation. Manufacturing of phosphatic fertilizers in India is faced with serious problems, as it requires the use of non-renewable resources, such as high grade rock-phosphate and sulphur, which are in short supply and are being depleted progressively; thereby, becoming costly. 101
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The magnatic or igneous rocks are the principle sources of phosphorus on our planet and the igneous rocks rich in phosphorus are industrially exploited. In India, the Mussourie rock phosphate is currently exploited, as it has 8.6% phosphorus or 20% P2O5. At Hirapur, M.P. India, in addition to phosphorus it contains high amounts of iron along with silica. Therefore, it becomes difficult to remove iron and silica by conventional chemical and physical means. Hence, the aim of the present study was to screen and select microorgamisms, giving maximum phosphate solubilizing activity to remove the phosphorus part or removing the impurities of iron and silica from the phosphatic ore of Hirapur, India. 2.
MATERIALS AND METHODS
2.1 Screening and selection of microorganisms 2.1.1 Phosphate solubilizing activity Our previous studies had indicated 10 species of bacteria, two of yeasts and four mycelial fungi to have a good phosphate dissolving activity. (5). These microorganisms were isolated from nine different ecosystems containing phosphates or having come into contact with phosphates. A chemical analysis of the rock phosphate from Hirapur is shown in Table 1. When these microorganisms were screened qualitatively by using Pikovskaya's agar medium or quantitatively in Pikovskaya's liquid medium, (6) it was found that the culture of Arthrobacter species gave maximum solubilization of 27.1% in 14 days (Table 2). This culture was used to find out the amount of phosphorus released from rock phosphate of Hirapur. 2.1.2 Removal of ferrous iron and silica from phosphate ore Such a beneficiation was attempted using Thiobacillus species on the phosphate ore crushed to -30 mesh size. From the different thiobacilli tried, Thiobacillus ferrooxidans strain B-101 from MACS Culture Collection MCM (7) was able to remove 86.5% ferrous iron from the ore along with silica. The phosphorus and silica content was analyzed by standard techniques (8) before and after the treatment and the analysis is shown in Table3. 2.1.3 Optimization of the beneficiation technique Since it was observed that instead of only solubilisation of phosphate, the beneficiation achieved by Thiobacillus species yielded good results, the optimisation was attempted only with Thiobacillus species, which removed the impurities of both silica and iron from the ore. The growth of the culture was attempted under various conditions of pH, temperature and nutrients to optimize the parameters and the results are reported in Table 4. 3.
RESULTS It can be seen from Table 1 that the ore contains, a high amount of iron and silica. Normal beneficiation operations, such as sieving and cleaning of ore using different techniques have very little effect in removal of these impurities.
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Table 1. Chemical analysis of samples of phosphatic rocks from Hirapur, India Element CaO P2O5 SiO2 MgO A12O3 Fe2O3 and FeO
Average Occurrence 65 - 63% 25 - 20% 5 - 3% 2 - 0.5% 1 - 0.5% 5-3 %
It is also seen from Table 2 that even though the Arthrobacter species is able to solubilize phosphorus to some extent, the net effect is that most of the phosphorus in the ore remains insoluble, probably due to impurities associated with the ore. Therefore, the second aspect of the studies, i.e. beneficiating phosphatic rocks by removing silica and iron was considered promising and was attempted. Table 2. Phosphatic solubilizing activity of Azotobacter sp., Arthrobacter sp. and Candida sp. in Pikovskaya's broth Cultures used Bacteria: Azotobacter sp. Arthrobacter sp. Yeast: Candida sp.
after 7 days
% phosphatic solubilization after 14 days
after 21 days
5.75 13.82
10.50 27.10
7.04 17.82
6.74
14.05
8.71
When the Thiobacillus strain B-101 was tried, it was found that on an average 86.5% of iron and 30.4% silica impurities were removed at laboratory level (Table 3). Therefore, it was thought that this method could be used for beneficiating the phosphatic rocks from Hirapur after optimization of various parameters for heap leaching on a large scale. Table 3. Beneficiation of phosphatic ore from Hirapur, India using a culture of Thiobacillus feroxidans (MCM B-101) Ore analysis Before treatment After treatment
Fe 4.5 0.6
SiO2 2.3 1.6
Upon optimization, it was found that the Thiobacillus culture can optimally grow at pH 3 and at 35° C. It becomes necessary to add a source of sulfur, such as pyrite as part of the nutrient supplement to allow the cultures to grow optimally. When the experiments are scaled up from the laboratory level to the heap leaching level, it was found that the efficiency of the process decreased slightly, when it was observed that 70% of the iron and 25% silica impurities were removed at that level. However, the process takes place in a continuous manner and therefore is advisable to use. The process parameters for treatment of 300 kg heap of ore are listed in Table 5. Based on the encouraging results obtained, it is proposed that the beneficiation of rock phosphate first by using microbiological means followed by physical and chemical treatment would be an important first step for producing clean phosphatic rocks for use as fertilizer in India. 103
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Table 4. Optimization of various parameters for growth of Thiobacillus ferrooxidans B-101 Parameter tested I. pH 2.0 3.0 4.0 5.0 II. Temperature 25°C 30°C 35°C 40°C III. Energy source Sulfur (1%) Ferrous sulfate (47%) Pyrite (1%)
Growth (cell density in 48 hrs) 4 x 107 1.3 x 108 2 x 107 0.5 x 107 6.6 x 107 1.3 x 107 4 x 108 1.6 x 108 3 x 107
Table 5. Beneficiation of phosphatic ore by heap leaching (Heap size = 300 kg) Optimum pH Optimum temperature Suitable Nutrient (Energy Source)
pH 3 30° C 1% Pyrite
Process Efficiency
Removal of Fe 70%
SiO2 25%
4.
DISCUSSION It was observed earlier in the microbial beneficiation of manganese ores, which was tried to remove phosphorus from the ore, that a combination of physical and chemical treatments, such as passing the crushed material after sieving through immersion magnetic separator (physical), followed by chemical precipitation in flotation cells, etc. had very little beneficial effect (9). Even though, some amount of iron and silica content was reduced, due to such treatments, it was not found effective in reducing the phosphorus content of the ore. However, a culture of Arthrobacter was observed to leach out 70% to 85% phosphorus from manganese ore (10). In the present study, it was observed that time 70% iron and 25% silica were removed from the phosphatic rocks in six days using a Thiobacillus culture. If physical and chemical beneficiation techniques are tried later on these 'cleaned' ores, it would provide an economically feasible operation to beneficiate phosphatic rocks, to be used as fertilizer. This is being tried at site on 1 t.p.d. plant now. REFERENCES
1. D.O. Norris, In: Tropical Pastures, W. Davis and C.L. Skidnore (eds.), Faber and Feber, London, (1968) 89-106. 104
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2. K.R. Sonar and G.K. Zende, J. Maha Agric. Univ., 9 (1984), 74-81. 3. S. Banik and B.K. Dey, Plant and Soil, 69 (1982), 353-364. 4. J.M. Vincent, In: Soil Nitrogen, W.V. Bartholomew and F.E. Clark (eds.), American Soc. Agronomy 10, (1965) 384-385. 5. S. Beheray and A.D. Agate, In: Bio-fertilizer Technology Transfer, A.V. Gangavane (ed.) Associated Publishing Co., New Delhi (1992), 193-197. 6. R.I. Pikovskaya, Mikrobiologiya, 17 (1948), 362-370. 7. S.M. Hideaki, Juncai, M. Satoru, S. Junko and T. Youka (eds.) World Directory of Collection of Cultures of Microorgamisms - Bacteria, Fungi and Yeasts, WFCC World Data Center on Microorganisms, Saitama (Japan), (1993), 130-131. 8. F.J. Welcher, Standard Methods of Chemical analysis, Part A, Vol II, Industrial and Natural Products and Non-instrumental Methods. 6th ed. D. Van Norstrand Co., Princeton, N. J. (1979). 9. A.D. Agate, In: Biogeotechnology of Metals, G. I. Karavaiko and S. N. Groudev (eds.) Center for International Projects, Moscow (1985), 349-395. 10. A D. Agate, In: Proceedings of the Symposium "Biological approach to problems in medicine, industry and agriculture", N.K. Notani and K. Sundaram (eds.), Bhabha Atomic Research Center, Dept. of Atomic Energy, Bombay (1974), 161-176.
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Biohydrometallurgy of antimony gold-bearing ores and concentrates P.M. Solozhenkin1 and V.P. Nebera2 a
Institute of Earth Resources Development RASc Kryukovsky Tupik, 4, Moscow, 111020, Russia;
[email protected] b Moscow State Geological Prospecting University, Miklucho - Maklaya, 23, Moscow, 117997, Russia;
[email protected] Abstract Experimental data on bacterial leaching of antimony-bearing ores of the Sarylakhsk and Sentochansk deposits (Yakutia) and ores of Tajikistan deposits. Bacterial accumulated culture Desulfovibrio desulfuricans have been employed as leaching and chemical agents. Encouraging results have been obtained in isolation antimony into the solution (recovery 96-98%). Thiobacillus ferrooxidans used for flotation selection of Hg minerals from antimonite and for transformation of antimonite into commercial product of antimony trioxide. 1.
INTRODUCTION Antimony gold-bearing ores (Sb2S3-Au) are wide spread in Russia, People's Republic of China, Republic of South Africa and Bolivia. In the Russian Federation, the main sources of antimony production are sarylakhsk flotation to produce antimony gold-bearing concentrates, obtained by enriching ores of the Sarylakhsk and Sentochansk deposits, Republic Sacha (Yakutiya) [1]. The current processing of Au-Sb concentrates is accomplished using pyrometallurgy. The gold is concentrated at the bottom of Au-containing antimony alloy ingots; the Au content ranges from 2 to 4% [2]. Only commercial trioxide of antimony is a salable product. When a hydrometallurgical alkaline leach method of production is applied, antimony is leached with the extraction of gold in cake. Soluble antimony is then subjected to cathodic electrodeposition to recover antimony [3, 4]. The existing methods for processing gold-antimony concentrates do not solve he problem of their complex use, since these processes do not make it possible to concentrate gold to products suitable for refining. Not all of Russia’s antimony needs are currently met. In this regard there are needs for geological prospecting of new deposits in Yakutiya, the Khabarovsk region, the Baikal region, in the Arctic and in Central Ural and new efficient technology developments for processing.
1 2
Prof. Dr. Sci., Academician Natur. Sci. (RAEN) and Tajik Acad Sci. Prof. Dr. Sci., Academician Natur. Sci. (RAEN) and Mining Sci. (RAGN)
107
Bioleaching Applications
A widely available cyanidation process is not applicable for gold extraction from Sb2S3-Au ores and concentrates, because of large cyanide consumption due to Sb reaction with NaCN and the locking of Au in the sulfide matrix. Use of thiourea allows to extract gold from gold-bearing stibnite, however, this development is still under laboratory investigation [5]. The problems of antimony leaching can be solved by applying an oxidizing acid leach. Other investigators have used ferric chloride to leach antimonite (stibnite) obtaining antimony trioxide [6-9]. Several microbial processes, including oxidizing and reducing type reactions, which can alter the physio-chemical, sorptive and flotation properties have been evaluated on various antimony minerals [10-12]. Some of these microbial processes are promising. When the surfaces of mercury and antimony sulfides are biomodified, selective separation is possible. It may also be possible to develop new microbial-produced solvents for Sb2S3 that would yield soluble antimony chlorides. Also, conventional bacterial leaching is promising. Kenzhalov et al. [17] from the Institute of Metallurgy and Mineral Processing, Republics Kazakhstan, have selected a heterothroph, Pseudomonos aureofaciens, from a Kazakhstan mineral deposit. They have studied in detail the influence of this heterothroph on antimonite, defining of the kinetics of the reaction, the order of reaction, the velocity constant, and the maximum velocity of product formation –(Mmax and the Michael’s constant, Km). Rapid dissolution of antimony from antimonite was obtained in the presence of the bacteria (0.11x10-4); without the bacteria the rate was 0.29x10-5. Bacterial affinity to antimony was quite high (Km = 0.07-0.37). Dissolution of sulphur surpassed that of antimony. Sulphate-reducing microorganisms (SRB) are wide-spread in nature. They utilize sulphate-sulphur as an electron acceptor. The most representative organism of the SRBs is Desulfovibrio desulfuricans. Intact cells of SRB rapidly reduce sulphate, while simple organic compounds or molecular hydrogen as follows: (1) Corg + SO42- = S2- + 2CO2 22(2) 4H2 + SO4 = S + 4H2O 2+ ∆F = -46410 kal (3) SO4 + H + 4H2 = HS + 4H2O A number of enzymes participate in the above reactions. Many studies have been conducted using SRBs in flotation of oxidized antimony and lead ores, as a desorbent and depressor in flotation of concentrates, for selection of lead from zinc minerals, and for the separation of molybdenite from chalcopyrite [10]. However researchers pay little attention to utilizing SRBs to produce H2S reagents for ore leaching and especially for leaching antimony-bearing materials. Lyubavina and coworkers (personal communication) have perfected a nutrient medium for the large-scale production of SRBs; this medium is based on the use of linter dust, a waste product of cotton-seed processing, as a carbon source for the organisms. Only alkaline leaching of antimony sulfides has achieved industrial application. This hydrometallurgical process results in high selectivity for the noble-metal groups, which remains in a leaching cake. Mel’nikov and coworkers found that dissolving antimony sulphide and antimony oxide in sodium sulphide and caustic soda results in sodium thioantimonite and thioantimonate formation according to the following reactions [2]: H2S + 2NaOH = Na2S + 2H2O (4) (5) Sb2Sb3 + 3H2S + 6NaOH = 2Na3SbS3 (6) Sb2S3 + 4NaOH = Na3SbS3+ NaSbO2+ 2H2O 108
Bioleaching Applications
Sb2O3+ 5Na2S + ÇÍ2Î = NaSb2S3+ 6NaOH (7) The objectives of this study were to: 1. Evaluate the use of sulphate-reducing bacteria (SRBs) as agents to produce H2S for the leaching of gold-bearing antimony sulfide minerals, and 2. Assess antimonite transformation into Sb2O3 using Thiobacillus ferrooxidans. 2.
EXPERIMENTAL
2.1 Materials and methods Antimonite (an antimony-sulphide mineral) ore and concentrates were investigated. A flotation sulphide concentrate, containing 39.55% antimony, was studied using 250-ml conical flasks at 28°C. An apparatus, specially designed for bacterial leaching, was also employed. Ores and concentrates for study were sterilized by boiling. In some experiments the solutions were filtered from insoluble residue. pH, the number of bacterial cells and antimony content were measured. Some antimony compounds were identified using spectral analyses with a quantometer VRA-2. For research purposes a SRBs were isolated from the Tyrny-Auz molybdenumwolfram deposit. Postgate medium [12] was used to cultivate the SRBs under laboratory conditions. Using a direct observation method a maximum of 220x106 cells.ml-1 of SRBs were found out to grow on the fourth day of culturing. During this period of time 400 mg.l-1 of soluble H2S was obtained; that was 10 times less than the maximum solubility of H2S in water. SRBs a ready-to-employ reagent solution of H2S for the leaching process after four days of growth. 2.2 The ores Ore from the Joint-Stock Co. Gold of Sacha (Yacutiya) was used to produce a flotation concentrate of antimony. This concentrate, produced by the Indigir-Gold plant, contained, %: 55 Sb; 22 S; 0.4 As; 30 g.t-1 Au and 3 g.t-1 Ag. An antimonite concentrate procured from the Adichanskiy operation at the Sentochanskaya Au-Sb deposit contained, %: 37 Sb; 14 S; 0.16 As; 60 g.t-1 Au, and 25 g.t-1 Ag. Thiobacillus ferrooxidans were cultured from the Bakyrchik and Olimpiadinsk deposits. The experimental testwork was performed in 250-ml Erlenmeyer flasks using 9K medium. 3.
RESULTS AND DISCUSSION
3.1 Sulfate-reducing bacteria as antimonite solvents Recovery of gold from refractory ores requires a pretreatment to liberate the gold particles from the host mineral. The antimony forms stable compound with NaCN during the cyanidation process. Pretreatment is usually an oxidation step. As an alternative, chemical or bioleaching can be applied to liberate the gold particles from the sulfur matrix. Emphases deserves an operation of sulfide-alkaline leaching as a way of selective separation of stibium from gold-bearing concentrates. Known developped by Irgiredmet (Irkutsk) technology of metallurgical conversion of rich gravity concentrates of Sarylachsk dressing plant (Au – 1050g.t-1; Sb - 65 %) [18]. The scheme includes sulfidealkaline leaching with the following electrolytic extraction of stibium from solutions. Three-phase recleaning of stibium leaching tailings from concentration tables, melting of secondary gravity concentrates (contents Au – 48 kg.t-1) on the dore metal and cyanidation 109
Bioleaching Applications
of tailig from final gravity concentration, received total extraction in corresponding commodity products Au - 98 %, Sb - 96,5%. The authors demonstrated the technical feasibility recovery of Sb by Na2S and NaOH leaching, the successive gold solubilisation by conventional cyanidation process and the recovery of Sb and Au from the respective leach solutions by electrowinnig [4]. Chemical basic leaching of pure stibnite by Na2S and NaOH under different experimental conditions at 40°C has been studied in order to optimise the reagents concentrations for the antimony dissolution process. Response surface methodology has been used to find the best experimental concentrations to maximize the Sb extraction yield. 98-100% of antimony recovery was obtained by using 1g Na2S and 1g NaOH per gram of pure stibnite. Treatment of a gold-bearing stibnite ore with cyanide yielded only 4% Au extraction; however, after seven days of bioleaching 85.5% Au recovery was attained (Table 1) [13]. Tests were conducted at laboratory scale utilising a refractory stibnite ore [14]. The gold content of the sample was 32 g.t-1. Bacterial cultures utilised in the biological test consisted predominantly of Thiobacillus. Table 1. Gold recovery by cyanidation, with and without biooxidation [14] Time, hours 1 12 24
Recovery Au, % 7 days bioleaching 52.2 63.5 85.5
No bioleaching 1.2 1.5 4.0
14 days bioleaching 53.2 64.1 86.0
At laboratory scale was investigated the best conditions for alkaline leaching of a refractory gold-bearing Sb2S3 (13.25 Sb2S3; 30 g.t-1 Au) coming from South America [19]. The solution was constituted by sodium sulfide and sodium hydroxide. Main parameters studied were: Na2S concentration, NaOH concentration, pulp density and temperature. It was reasonable to check a possibility of leaching stibium by bacteria. The selective flotation and separation of cinnabar from antimonite minerals using T. ferrooxidans [15] is illustrated in Figure 1. Bacterial conditioning of 5 h did not affect cinnabar flotation (recovery 89.6%), while the antimonite recovery by flotation decreased from 89 to 6.2%; this led to almost complete selection of the minerals. The results presented in Figure 1 indicate the superiority of biological separation compared to chemical separation. 100
Recovery, %
80 60 Sb
40
Hg
20 0 1
2
3
4
5
Time, h
Figure 1. Changes in flotation recovery (%) of Sb2S3 and HgS minerals in processing by T. ferrooxidans from time, h 110
Bioleaching Applications
As a result of bacterial oxidation, antimonite is converted to antimony trioxide, the mineral senarmontite. T. ferrooxidans oxidizes the surface of antimonite crystals while cinnabar remains intact. As a result, HgS is floated and extracted into the concentrate, while Sb2S3 is coated by a fine film of oxides (Sb2O3) and removed in the tailings. Partly regenerated culture can be recycled in this process. When the bacterially-oxidized products of the antimonite (stibnite) concentrate were analyzed by x-ray diffraction, no Sb2S3 was observed. However, XRD analysis of the original concentrate revealed intensive lines, belonging to antimonite [16]. Biooxidation also reduced the content of other elements in the stibnite concentrate. Antimony leaching was done by somewhat different technique. Antimony-containing portion was mixed up with SRB of different hydrogen sulphide concentration and with caustic soda solution. The pulp was heated up to 90°C and it was stirred with S:L = 1:16 ratio. Antimony solubility is most effective at 120 g.l-1 caustic soda concentration and maximum hydrogen sulphide concentration in SRB. Special experiments established that the time necessary for leaching of antimony is 1 or 1.5 h. To increase antimony transition into the solution, it is necessary to increase contact time with SRB during leaching. However, when the time of contact was increased, it was necessary to control antimony ions in the solution whose optimal concentration slowed down the leaching process. Therefore, antimony leaching was done in two stages with gradual addition of reagents. The first stage lasted 1 h, after that the solution was decanted, then once again necessary reagents were added and contacted for 0.5 h. Effective antimony leaching was observed under maintained optimal conditions. Isolation of antimony into the solution under those conditions was about 96.5-98.0%. Results obtained in antimony leaching from antimony-bearing raw materials are given in Table 2. These data show the extent of effective application of bacteria as compared to sodium sulphide in antimony leaching. Furthermore, the suggested technology of antimony leaching has a number of advantages. The presence of intact alkali in electrolyte leads to considerable increase of electric conductivity in the solution and promotes better results of the subsequent electrolysis. Table 2. Effect of sulphate-reducing bacteria (SRB) and Na2S on leaching Sb from antimony-bearing materials* Antimonite Stage I . -1
Stage II
SRB, mg l
8.7
17.5
43.7
52.5
8.7
7.5
43.7
52.5
Sb recovery, %
80.5
83.2
94.9
91.8
95.6
93.6
96.79
98.04
Antimony concentrate Stage I
Stage II
SRB, mg.l-1
25.4
70.9
86.7
139.1
25.4
70.9
87.7
139.1
Sb recovery, %
41.3
46.7
56.8
77.7
92.9
92.9
94.3
96.5
Antimonite . -1
Na2S, mg l
Sb recovery, %
60
20
180
240
300
360
51.4
78.8
92.2
98.0
98.9
98.9
. -1
*NaOH 120 g l , leaching time 1 h, temperature 90°C
111
Bioleaching Applications
Thermodynamic calculations show that hydrogene sulphide oxidation occurs at a lesser rate, than that of sodium sulphide. The concentration of S2- and SH- ions in the pulp containing SRB is weaker than that in the pulp, containing sodium sulphide. Traditional technology of antimony leaching requires higher concentrations of sodium sulphide, than in the case of SRB leaching, the latter reducing yield on the electric current. Given cake is processed by usual methods: - cyanidation, since contents of stibium does not render influences upon the leaching of gold; - presence in cake sulfur (more than 14%) must be sodium neutralized; - by gravity concentration methods in centrifugal devices of Knelson type with following separation of concentrate in ferro-magnetic liquid to receive rich Au-containing concentrate, ready for affinage. 3.2 Transformation of antimonite into antimony trioxide by Th. ferrooxidans Information on antimonite biooxidation is of certain interest [14]. According to the results of investigations carried out by Irgiredmet, Sb2S3 oxidation with Th. ferrooxidans was described in [20, 21]. Antimonite biooxidation realized in relatively pliant regime to improve the technological characteristics of cyanided material due to the release of gold associated with Sb2S3 and transformation of antimony to the less active chemical form [16]. It was established that bacterial oxidation of gold-arsenic concentrates by Th. ferrooxidans occurred within 100-120 h, the high degree of sulfide oxidation achieved (%) 96-98 of arsenopyrite, 97-98 of pyrrhotite, 92-95 of antimonite, and 65-84 of pyrite [17]. Under biochemical leaching was observed greater amount of the oxidized forms of stibium and also its oxides of high valences regardless of initial contents of stibium minerals. After bacterial influence an intensity of lines D 5.05 D 5.66 A, belonging to Sb2S3. Thionic bacteria oxidized animonite sulfur and formed antimony trioxide of cubic syngony of senarmontite type (the heat of formation of ∆H278 = 165.5 kJ) according to the following reactions: (8) 2Sb2S3 + 12O2 + 6H2O = Sb4O6 + 6H2SO4 (9) 3Sb2S3 + 3Fe2(SΟ4)3 + 6H2O = Sb4O6 + 12FeSΟ4 + 3S (10) S + H2Ο + 3/2Ο2 = H2SO4 (11) 2FeSΟ4 + l/2Ο2 + H2SO4 = Fe2(SΟ4)3 + H2Ο (12) Sb4O6+ H2SO4 = 2Sb2(SΟ4)2 + 6H2O The rhombic form of Sb2O3 is obtained in hydrolysis of antimony-chloride solutions. Table 3 presents the data of influence exerted by bacterial processing on the change in phase content of minerals under different conditions of the experiments. The action of diluted H2SO4 on antimony sulfate results in hydrolysis with formation of antimonyl sulfate: (13) Sb2(SΟ4)3 + 2H2Ο = (SbO)2SΟ4 + 2H2SO4 Senarmontite processed into antimony trioxide by the reactions: (14) Sb4O6 + 12HCl = 4SbCl3 (15) 2Sb2(SO4)2 + 6HC1 = 2SbC13 + H2SO4 (16) SbCl3 + H2O = SbOCl + 2HCl 112
Bioleaching Applications
SbOCl + 2NH4OH = Sb2O3 + 2NH4Cl + H2O
(17)
Table 3. Transformation minerals by Th. ferrooxidans Minerals Senarmontite Antimonite Pyrite Quartz
Initial content, % 0 71.7-81.5 13.3-6.8 15.0-11.7
No. 1 94.6 26.0 0.1 2.7
No 1' 93.9 1.5 — 46.0
Contents after treatment, % Samples numbers No 2 No 2' No.3 No 3' 87.9 88.3 72.0 69.3 7.5 37 8.7 10.8 1.1 — 6.2 6.3 35.0 80.0 12.9 13.6
No 4 40.9 79.0 50.0 46.1
No 4' 44.7 68.0 64.0 44.7
Figure 2. Recommended flow sheet of processing Au-Sb concentrates
Selected stibium trioxide, the contents not below 99.5% Sb2O3 and recovery of stibium in the product - 90.5 %. On the contents of allowing admixtures stibium trioxide corresponds to analytical grade (Tech.Cond. 6-09-3267-84) and exceeds requirements for the lavsan production (Tech.Cond. 6-09-2897-77). Offered technological conversion scheme for Sb-Au concentrates settles problems of their complex using and allows to get high-quality stibium trioxide and gold-containing product. One of the ways of oxidation of stibium minerals is a hydrogen peroxide oxidation, prodused by heterothrofic bacteria [17]. In the reactionary ambience, oppressing development of bacteria, occurs improvement of respiratory activity that brings about hypersynthesis of H2O2. Senarmontite disolves in the alkaline water-glycerin medium containing 250-300 g.l-1 glycerin, 50-60 g.l-1 NaOH, and 100-20 g.l-1 Sb with the complex formation: 113
Bioleaching Applications
Sb2O3 + 2NaOH + 2C3H8O3 = 2NaOSbO2C3H5OH + 3H2O (18) Leaching was carried out for 30-60 min without electrolyte heating and with mechanical air mixing. During electrolysis of water-glycerin solutions, antimony was separated from the complex anion on the cathode: (19) SbO3C3H5OH- + H2O + 3e = Sb + 2OH- +C3H5O2OH2Thus, the technological scheme is proposed for producing antimony trioxide or cathode antimony (Fig. 2). Realization of the technology will make it possible to begin development of the Sentachansk deposit (Yakutia). The bacterial transformation of Sb2S3 into Sb2O3 from gold-antimony concentrate favours production of the material suitable for cyanidation. In this case, several advantages can be expected: 1. An increase in market cost of antimony in concentrate due to its production in the form of trioxide (up to 83.53%), decrease in volumes of ore-mass transportation, and reduction in arsenic content in bacterial processing. 2. Reduction in time for gold and silver return to metallurgical conversion. 3. Removal of arsenic from concentrates, which will favour improvement of their quality and release of gold-bearing minerals. 4. Creation of small-scale production of antimony trioxide and diluted sulfuric acid suitable for ore processing and other purposes directly in the deposit. The bacterial transformation of antimony sulfide in trioxide from gold-antimony concentrates produced material applicable for the cyanidation. Herewith possible expect a number of essential advantages. 4.
CONCLUSIONS 1. The most advanced results in bioleaching of minerals were echieved with thiobacteria on sulphides of iron and some non-ferrous metals. These data showed the extent of effective application of bacterial strains as compared to sodium sulphide in antimony leaching. Furthermore, the suggested technology of antimony leaching has a number of advantages. Antimony recovery, as well as transformation of sulfates into Sb2O3, discussed here, are of great interest. 2. Increased stibium market value in the concentrate as Sb2O3 (up to 83.53%) in contrast with Sb2S3 (71.69%) and reduced volumes of transportation, reduction of arsenic contents in the process of biological conversion. 3. Increased return of gold and silver to metallurgical processing. 4. Removal of arsenic from concentrates promotes both: a raise of concentrate quality and opening gold-containing minerals.
REFERENCES
1. E.A. Kozlovskiy, Mineral and Raw Material Problems in Russia on the Eve of XXI Century (State and Prediction). Antimony. Mosk. Gos. Gorn. Universitet, Moscow, Russia (1999). 2. S.M. Mel'nikov (ed.), Antimony. Metallurgia, Moscow, Russia (1977). 3. Zhao Tian-cong, The Metallirgy of Antimony, Central South University of Technology Press, China (1988). 4. S .Ubaldini, F. Veglio, P. Fornari, C. Abbruzzese, Hydrometallurgy, 57 (2000) 187. 114
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5. M.A. Meretukov and A.M. Orlov, Metallurgy of Noble Metalls, Foreign Experience, Metallurgia, Moscow, Russia (1991). 6. P.M. Solozhenkin, S.V. Usova, T.N. Aknazarova, and R. R. Fazylova, Tzvet. Metally, 1 (1994) 23-26. 7. P.M. Solozhenkin, Proceedings of the XIX IMPC, (1995) 223-226. 8. P.M. Solozhenkin, V.P. Nebera, and I.G. Abdulmanov, Proceedings of the XX IMPC, 4, Aachen (1997) 227-237. 9. P.M. Solozhenkin, V.P. Nebera, and I.G. Abdulmanov, Proceedings of the 7th International Symposium on Mineral Processing, Balkema, Rotterdam, Brookfield (1998) 495-500 10. P.M. Solozhenkin and V.P. Nebera, Proceeding of the IV International Conference on Clean Technologies for the Mining Industry, 1, University of Concepcion, Chile (1998) 399-407. 11. P.M. Solozhenkin, N.N. Lyalikova-Medvedeva, Journal of Mining Science, Vol. 37, No. 5 (2001) 534-541. 12. P.M. Solozhenkin, N.N. Lyalikova-Medvedeva, Physical technical problems of mining, No. 5 (2001) 95. 13. C. Abbruzzese, S Ubaldini, F. Veglio, Acta Metallurgica Slovaca, Special Issue, 4, 1 (1998) 19-24. 14. S. Ubaldini, F. Veglio, L. Toro, C. Abbruzzese, Minerals Engineering, Vol. 13, No. 14-15 (2000) 1641. 15. V.P. Nebera , P.M. Solozhenkin, N.N. Lyalikova-Medvedeva, Proceeding of the 7th International Conference on Mining, Petroleum and Metallurgical Engineering (MPM’7-Assiut) V. II Metallurgy & Mineral Processing, Assiut University, Egypt (2001), 295-303. 16. G.I. Karavaiko, G.V. Sedel'nikov, R.Ya. Aslanukov, U.U. Savari, V.V. Panin, E.V. Adamov, and E.F. Kondrat'eva, Tsvet. Metally, No. 8 (2000) 20-26. 17. B.K. Kendzhalov, G.V. Semenchenko, T.O. Ospanbekov, Ch.G. Absalyamov, M.M. Ignat’ev, Strain of bacteria Pseudomonas aureofaciens T 10 IMI, desolving gold from sulfide ores and concentrates, Kazakstan Patent PP¹ 11409 (2002). 18. F. Veglio, S. Ubaldini, EJMP&EP, Vol. 1, No. 2, (2001) 103-112. 19. A.F. Panchenko, V.V. Lodeyshchikov, V.Ya. Bivaltcev, Development of productive forces of Sybiria, Krasnoyarsk. V. 1, Pt. II (1985) 354-358. 20. V.V. Lodeishchikov, State of Investigations and Practical Developments in the Field of Bioliydrometallurgical Processing of Rebellious Gold-Bearing Ores and Concentrates: Review. Irgiredmet, Irkutsk, Russia (1993). 21. V.V. Lodeishchikov, A.F. Panchenko, 0.D. Khmel'nitskaya and L.P. Semenova, Ore beneficiation, Collected Works, Issue 5, Irkutsk, Russia (1994).
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Bioleaching of Argentinean sulfide ores using pure and mixed cultures Frizán V.♦, Giaveno A.♦, Chiacchiarini P.♦, Donati E.♠ ♦
Facultad de Ingeniería, Universidad Nacional del Comahue, Buenos Aires 1400 (8300) Neuquén, Argentina* ♠ Centro de Investigación y Desarrollo de Fermentaciones Industriales (CINDEFI-CONICET), Facultad de Ciencias Exactas, Universidad Nacional de La Plata, 47 y 115 (1900) La Plata, Argentina Tel/Fax +54 221 4833794. E-mail:
[email protected] Abstract The objective of this work was to evaluate the efficiency of the bioleaching of three different Argentinean sulfide ores using pure and mixed cultures of Acidithiobacillus ferrooxidans and Acidithiobacillus thiooxidans. The samples used were obtained from La Silvita, La Resbalosa and Mallín Quemado ores. Their main constituents are quartz galena, sphalerite, and chalcopyrite and their chemical composition includes 8.98% Zn, 0.046% Cu, 5.86% Pb and 0.55% Mn (La Silvita sample), 17% Zn, 0.064% Cu, 2.37% Pb and 0.56% Mn (La Resbalosa sample) and 7.95% Zn, 0.022% Cu, 64.7% Pb and 1.5% Mn (Mallín Quemado sample). Bioleaching experiments were carried out in glass columns with the percolation of the medium through the minerals. Solubilized metals (zinc, copper, manganese and total soluble iron) were determined using an atomic absorption spectrophotometer while iron(II) was measured by titration. Solid residues recovered by filtration were analyzed by means of X-ray diffraction. Metal recoveries from La Silvita and La Resbalosa were significantly enhanced by inoculation. The highest extraction was reached using Acidithiobacillus ferrooxidans.
Keywords: bioleaching, sulfide ores, zinc, acidithiobacillus 1.
INTRODUCTION Bioleaching is applied to ores which cannot be treated economically by conventional processes like flotation and roasting [1]. Acidithiobacillus ferrooxidans and Acidithiobacillus thiooxidans bacteria are capable of acting directly or indirectly on metallic sulfides, oxidizing the sulfides to sulfate and thus releasing the metals in those cases in which the respective sulfates are soluble. For this reason, these microorganisms were traditionally used in bioleaching processes of low-grade ore on a commercial scale [2].
* The authors wish to acknowledge the financial support provided by Agencia Nacional de Promocion Cientifica y Tecnologica (PICT99) and Universidad Nacional del Comahue, both of them from Argentina.
117
Bioleaching Applications
Although there are several minerals amenable to bioleaching or biooxidation on a commercial scale, only two metals, copper and gold, are currently recovered using this technology. Nowadays, important efforts are being done to adapt heap biooxidation technology to treat sphalerite concentrates [3]. On the other hand, in western Patagonia (Argentina) there are some reservoirs containing important amounts of base metal sulfides, which could be suitable to biohydrometallurgy commercial application. Therefore, it is interesting to study the application of this technique at these ores thoroughly. The objective of this work was to evaluate the efficiency of the bioleaching process of three different complex Zn-Mn-Fe-Pb sulfide ores from Neuquén (PatagoniaArgentina) using pure and mixed cultures of A. ferrooxidans and A. thiooxidans. 2.
MATERIALS AND METHODS
2.1 Microorganisms and media Pure and mixed cultures of A. ferrooxidans (DSM 11477) and A. thiooxidans (DSM 11478) were used throughout this study. The first strain was cultivated routinely in 9K medium of initial pH 1.80 [4]. The cells were harvested when the culture had consumed 90% of the iron (II) available. The culture was filtered through blue ribbon filter paper to retain the jarosite deposits and then through a filter Millipore of 0.22 microns in order to retain the cells. Cells were washed at least twice with iron-free 9K medium and finally resuspended in the same medium pH 1.8. The second strain was cultivated routinely in iron-free 9K medium with sulfur as energy source. The cells were collected when the pH descended below 1.0. The procedure for the preparation of the inoculum was similar to the one applied for the other strain. These suspensions (with bacterial populations of approximately 2x108 cells/ml) were used as inoculum at the 10%v/v. 2.2 Mineral The samples used were obtained from La Silvita, La Resbalosa and Mallín Quemado ores (Province of Neuquén, Patagonia Argentina). Chemical Analysis and Mineral Composition of the samples are given in Table 1. Table 1. Chemical Analysis and Mineral Composition Mineral
Fe (%)
Zn (%)
Pb (%)
Mn (%)
Cu (%)
Mineral Composition
La Silvita
15.03
8.98
5.86
0.54
0.046
ZnS 5.5%, PbS 3%, FeS2 12%, CuFeS2 0.5%
La Resbalosa
9.34
17.01
2.37
0.56
0.064
ZnS 12%, PbS 2%, FeS2 8%, CuFeS2 1%
Mallín Quemado
2.94
0.795
64.75
1.55
0.022
PbS 96%
2.3 Detection of indigenous bacteria In order to detect sulfur or iron oxidizing bacteria two tests in 250 ml shake flasks on a rotatory shaker at 180 rpm and 30 ± 0.5 ºC were carried out. Each flask, containing the ore (pulp density of 10%), was respectively filled with 100 ml of the appropriate liquid medium (9K iron free pH=3.0 supplied with 1% w/v of sulfur powder or 9K medium pH=1.8).
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2.4 Experiments in columns Tests were conducted in 50 mm inside diameter glass columns with 260 mm high at constant temperature of 30ºC ± 0.5ºC. Columns have a perforated plate in the bottom and a layer of glass wool to support the ore sample. Thirty grams of each ore sample with particle size ranging between 10 and 16 mesh were used in every column under flood conditions. Each column was charged with 300 milliliters of free-iron 9K medium at initial pH 1.8. Percolating solution was recirculated from the open top of the column to the bottom by continuous airflow rate at 1.2 VVM. In order to reach a pH condition compatible with the bacterial growth it was necessary to achieve an acid stabilization before the inoculation. Therefore, the pH value was initially adjusted at 1.8 by adding drops of sulfuric acid solution (4.8 N) and then it was not controlled again. The amount of sulfuric acid added was used to calculate the initial acid consumption. After the pH stabilization, the inoculation was done. Sterile control columns were prepared replacing the same volume of inoculum by a solution of 2% timol in methanol. Additionally, uninoculated control columns were prepared to check indigenous bacterial activity. Sterile distilled water was added to compensate evaporation. Samples from every column were taken at regular intervals. 2.5 Analytical methods Copper, total iron, manganese and zinc in solution were determined by atomic absorption spectrophotometry. Iron (II) concentration was determined by permanganimetry. Bacterial populations in solution were determined using a PetroffHausser camera in a microscope with a contrast phase attachment. This determination was not representative of the bacterial growth because the cells attached to the mineral were not determined. Both the redox potential and the pH were measured with specific electrodes Sulfuric acid production was analyzed by titration with sodium hydroxide solution. Solid residues were recovered by filtration and analyzed by X-ray diffraction (XRD) in Rigaku DII-Max equipment. 3.
RESULTS AND DISCUSSION
3.1 Detection of indigenous bacteria After 25 days, ferrous iron was completely oxidized, but no free cells in media were observed at microscope. On the other hand, the pH values in the systems supplemented with sulfur did not decrease. Moreover the pH remained almost constant in the case of La Silvita (pH 3.5) whereas it increased in La Resbalosa and Mallín Quemado (pH 4.9). Since it was not possible to detect free bacterial cells, ferrous iron was oxidized more slowly than the usual by A. ferrooxidans and cells were not able to oxidize sulfur as energy source, it is possible to suggest that the indigenous bacterial activity was due to the presence of Leptospirillum ferrooxidans-like microorganisms. As it is known, these bacteria grow preferentially adhered to surfaces and their physiology is consistent with the result observed [5]. 3.2 Experiments in columns The experiments took one hundred days. After that period, the solid residues were removed from the columns and analyzed by XRD. The main mineralogical species originally present in the ores remained until the end of all leaching process indicating that 119
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the oxidation of sulfides associated with the dissolution of metals was not complete. Additionally, after the leaching processes it was possible to identify sulfur and jarosite from La Silvita and La Resbalosa ores both in the biotic and abiotic systems. Galene present in the ore treated was oxidized to anglesite only in the biotic systems showing that bioleaching processes were important to improve this phase transformation. The initial sulfuric acid consumption before the inoculation was: 39.2 g H2SO4/kg ore to La Silvita, 40.5 g H2SO4/kg ore to La Resbalosa and 43.5 g H2SO4/kg ore to Mallín Quemado, which indicates the presence of slightly alkaline gangue content such as carbonate minerals. 3.3 La Silvita Figure 1 shows the percentages of zinc and manganese solubilized from La Silvita ore during the test. The highest zinc extraction (75%) was obtained in the column with A. ferrooxidans. In addition, zinc solubilization in the column with mixed culture (48%) was significantly higher than the one with A. thiooxidans (22%). The performance of the last column was similar to the uninoculated one (21%) throughout the experiment. The zinc extraction in the sterile column only reached 7%. These results suggest that A. ferrooxidans plays a key role in the bioleaching process for La Silvita ore contributing to release the zinc from the sphalerite. Meanwhile A. thiooxidans as pure or mixed cultures was not so efficient to improve the Zn solubilization. This was probably due to the mineralogical species present in this natural ore, since the extraction obtained in this work was lower than those reported using synthetic sulfide [6]. In contrast with the zinc solubilization, the amount of manganese released from La Silvita ore was higher when A. thiooxidans was present in the cultures. Figure 1 shows that the behavior displayed by this pure culture was very similar to the mixed one, reaching in both cases a manganese extraction close to 100%. Since the manganese solubilization was higher in presence of A. thiooxidans or mixed cultures and considering that these systems reached pH values lower than the other systems (Fig. 2), manganese could be present as a mineralogical species easily leachable by acid. This hypothesis could not be confirmed by XRD analysis probably due to the low amount of manganese present in this ore. In Figure 2, the pH evolution can be observed. During the first twenty-five days, all columns showed an increase of pH values indicating that the solubilization of some basic species present in the mineral continued beyond the inoculation. After that period, pH values decreased in all biotic system around 1.7 while the pH value increased to reach a value of 2.6 in the sterile system. These results could indicate that sulfur obtained from the redox dissolution of sulfides and identified by XRD analysis of leached residues, was partially oxidized by bacterial action and contributed to the acid production. Therefore, the whole bioleaching process had a positive acid balance. Additionally, in Figure 2, the redox potential evolution can be observed. Eh rapidly increased during the first days of the experience in columns inoculated with A. ferrooxidans reaching values over than 550 mV. Then, Eh values oscillated around this threshold until the end of the experiment. The uninoculated system took more than twenty days to reach the same final value. In the sterile column Eh remained constant close to the initial value.
120
Bioleaching Applications Af At Af/At Uninoculated Sterile
80
60
60 40
2+
40
2+
Zn Solubilization (%)
80
100
Mn Solubilization (%)
100
20 0
0
20
40
60
80 100
20 0
0
20
40
60
80
100
Time (days)
2.6 2.4
Eh (mV)
Figure 1. Comparison of zinc and manganese extraction from La Silvita ore using pure and mixed cultures of A. ferrooxidans (Af) and A. thiooxidans (At) with uninoculated and sterile control systems 600 500 400
0
20 40 60 80 100
Time (days)
Af At Af/At Uninoculated Sterile
pH
2.2 2.0 1.8 1.6 0
20
40
60
80
100
Time (days)
Figure 2. Evolution Eh and pH values from the La Silvita under different conditions: A. ferrooxidans (Af) cultures, uninoculated system and sterile control
Figure 3 shows the evolution of total soluble iron and ferrous iron concentrations in columns inoculated with A. ferrooxidans, uninoculated and sterile controls. The iron evolution was completely in agreement with Eh values shown in Figure 2. Ferrous iron was rapidly oxidized in A. ferrooxidans column and slightly slower in the uninoculated one. On the other hand, the ferrous iron remained reduced in the sterile controls reaching 2.5 g/l. Additionally a significant amount of iron was released from La Silvita ore during the leaching process mainly in A. ferrooxidans column (12 g/l). 121
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Fe(II) (g/l)
12 10
FeTotal(g/l)
1 0 0
8
Af Uninoculated Sterile
2
20 40 60 80 100
Time (days)
6 4 2 0
0
20
40
60
80
100
Time (days)
Figure 3. Evolution of total iron and ferrous iron concentrations from La Silvita under different conditions: A. ferrooxidans (Af) cultures, uninoculated system and sterile control
These results show: i) A. ferrooxidans cultures remained active along the experience and the cells in suspension reached a value of 2.5x108cells/ml after one hundred days of operation. ii) This microorganism contributed to increase the solubilization of pyrite and other iron species. And iii) the bacterial activity detected in the uninoculated column was probably due to another ferrous iron oxidizing L. ferrooxidans-like bacteria. From Figures 2 and 3 the total iron dissolved with the pH evolution can be correlated either in the sterile or biotic systems. The amounts of iron dissolved were minimal when the values of pH were maximum. Correspondingly, an abundant amount of jarosite and other ferric oxyhydroxides covering solid residues was visually observed and then detected by X ray diffraction analysis. 3.4 La Resbalosa Figure 4 shows the percentages of zinc and manganese solubilized from La Resbalosa ore during the test. In the column with A. ferrooxidans the highest extraction of zinc was obtained, reaching 17.5%. In addition, zinc solubilization in the column with mixed culture (12.4%) was slightly higher than the one observed in the column inoculated with A. thiooxidans (9.7%) and in the uninoculated system (9.3%). The sterile control only removed 2.6% of zinc initially present in the ore. Figure 5 shows pH evolution and total iron solubilized throughout the experience. When the pH values increased, the iron in solution decreased. In A. ferrooxidans inoculated column the solubilization of different basic mineralogical species present in the ore caused an increase of pH values of over 3.3. After that, an important precipitation onto the mineral surface was observed. Moreover, the zinc extraction and the bacterial counts (1x108 bacteria/ml) were lower than those obtained from La Silvita. It was probably associated with diffusional barriers due to the large amount of brown amorphous ferric precipitated, which avoided further sphalerite dissolution. The inoculated systems with A. thiooxidans, as pure or mixed cultures reached the highest manganese solubilization (Fig. 4). Although La Resbalosa and La Silvita ores 122
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showed a similar behavior, in the first ore the percentage of manganese extraction was only 70%. This was probably due to the fact that pH values reached in La Resbalosa systems were higher than those obtained in La Silvita. Af At Af/At Uninoculated Sterile
20
70 Mn Solubilization (%)
10
5
0
60 50
2+
2+
Zn Solubilization (%)
15
80
0
20
40
60
80 100
40 30
0
20
40
60
80
100
Time (days)
Figure 4. Comparison of zinc and manganese extraction from La Resbalosa using pure and mixed cultures of A. ferrooxidans (Af) and A. thiooxidans (At) with uninoculated and sterile control systems 4
pH
1200
Fe(total) Concentration (mg/l)
1000
Af At Af/At Uninoculated Sterile
3 2
800
0
20 40
600
60 80 100
Time (days)
400 200 0
0
20
40
60
80
100
Time (days)
Figure 5. Evolution of total iron and pH from La Resbalosa using pure and mixed cultures of A. ferrooxidans (Af) and A. thiooxidans (At) with uninoculated and sterile control systems 3.5 Mallín Quemado Figure 6 shows the evolution of zinc solubilization from Mallin Quemado ores. In contrast to the bioleaching experiences using La Silvita and La Resbalosa ores, Mallin Quemado reached higher Zn extraction when A. thiooxidans cultures were present. On the 123
Bioleaching Applications
other hand, the Zn solubilization was very similar in the A. ferrooxidans inoculated, uninoculated and sterile systems, and the metal extractions were not relevant in all these flasks. pH values at the end of the test were 2.0 in A. thiooxidans pure and mixed cultures and 2.3 in the other biotic systems while the sterile system reached a pH of 3.5. Eh values were increasing according to ferrous iron oxidation (data not shown). Af At Af/At Uninoculated Sterile
18
24
14
Mn Solubilization (%)
12 10 8
2+
2+
Zn Solubilization (%)
16
28
6 4
20 16 12 8
0
20
40
60
80 100
0
20
40
60
80
100
Time (days)
Figure 6. Comparison of zinc and manganese extraction from Mallín Quemado ore during leaching experiences using pure and mixed cultures of A. ferrooxidans (Af) and A. thiooxidans (At) with uninoculated and sterile control systems
The manganese extraction was higher in the systems inoculated with A thiooxidans cells. The different percentages of manganese removed could be attributed to the final pH in each system. The final percentages of Zn and Mn solubilized from Mallín Quemado ore were lower than those obtained using the other minerals. Bacterial counts were also very low (0.6 x 108 bacteria/ml) when this high-grade galena ore was tested. These results could indicate that the great lead percentage in the sample may inhibit the bacterial growth. Meanwhile the low solubility of lead (II) reduced the possibility of toxicity when it is present in lowgrade ores as La Silvita and La Resbalosa [7]. 4.
CONCLUSIONS The bioleaching treatment of La Silvita complex sulfide ore by bioleaching process appears to be technically feasible, since the zinc solubilization was increased ten fold when A. ferrooxidans was used as inoculum. Although La Resbalosa tests showed that the bioleaching processes were amenable, they could be improved if a rigorous pH control is done to avoid great oxides precipitation onto the mineral surface. Mallin Quemado tests did not show significant zinc extraction but it was possible to oxidize galene. Further tests should be carried out to analyze the extension of this transformation and the effects of lead toxicity on the bacterial grow. In summary, the results analyzed here are good enough to consider the metal biohydrometallurgical extraction as a promissory method to be applied for these regional ores to be treated using bacterial heap leaching technique. Moreover, future studies
124
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involving isolation and identification of indigenous bacteria should be done in order to improve the bioleaching treatment of these Argentinean ores. REFERENCES
1. W. Sand, T. Gehrke, R. Hallmann, K Rohde, B. Sobotke and S. Wentzien. Biohydrometallurgical Technologies, E. Torma, J.E. Wey and V.L. Lakshmanan (eds). The Minerals, Metals & Materials Society (1993) 15. 2. D.E. Rawlings. Biomining: Theory, Microbes and Industrial Processes, SpringerVerlag, Berlin, 1997. 3. T.J. Harvey, W. Van Der Merwe K. and K. Afewu, The applicacion of the GeoBiotics GEOCOAT biooxidation technology for the tratment of sphalerite at Kumba resources´ Rosh Pinah mine. Minerals Engineering15 (2002) 823. 4. M.P. Silverman and D.G. Lundgren, J. Bacteriol., 77 (1959) 642. 5. W. Sand, K. Rohde, B. Sobotke and C. Zenneck. Evaluation of Leptospirillum ferrooxidans for leaching. Appl. Environ. Microbiol. 58 (1992) 85. 6. M. Pistorio, G. Curuchet, E. Donati and P. Tedesco. Direct zinc sulphide bioleaching by Thiobacillus ferrooxidans and Thiobacillus thiooxidans. Biotechnol. Lett. 16 (1994) 419. 7. J. Barrett, M.N. Hughes, G.I. Karavaiko and P.A. Spencer. Metal extraction by bacterial oxidation of minerals. Ellis Horwood Limited, 1993, 28.
125
15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Bioleaching of complex gold-lead ores Z. Ulberg, V. Podolska, A. Yermolenko, L. Yakubenko and N. Pertsov Ovcharenko Institute for Biocolloidal Chemistry, NAS of Ukraine, 42 Vernadsky blvd., 03142 Kyiv, Ukraine Abstract The present work is relevant to bioleaching of galena from gold-lead ore using Acidithiobacillus such as Thiobacillus ferrooxidans and Thiobacillus thiooxidans in the presence of magnetite as well as under action of DC electric field, which resulted in the acceleration of sulfide dissolution from the natural and synthetic galena. Various electrochemical mechanisms in the leaching process were considered. The electrokinetic properties of the thiobacteria under the conditions of galena microbial leaching were also studied.
Keywords: galena; bioleaching; magnetite; galvanic couple; zeta-potential 1.
INTRODUCTION In recent years, the use of microorganisms for metal solubilization from ores has increasingly attracted the attention of hydrometallurgists and biotechnologists. Microbial leaching is characterized by low cost, and its realization continues to create fewer problems compared to corresponding hydrometallurgical or pyrometallurgical processes. The bioleaching method is based on the ability of some microorganisms to oxidize Fe (II) ions or reduced sulfur compounds. As a result of the accumulation of sulfuric acid in the biosuspension, the decrease in pH and the metal solubilization from sulfides takes place. The cultures Thiobacillus ferrooxidans and Leptospirillum ferrooxidans are most-used [1]. Galena is the main industrially available source of lead. Hydrometallurgical processes for lead leaching have been studied by a number of researchers [2]. Quite a few studies have been dedicated to the use of microorganisms for lead leaching from lead-bearing ores [3, 4]. In gravitation and flotation separation of ores where the gold-bearing minerals are associated with the sulfide minerals of non-ferrous metals one obtains concentrates enriched with these metals, namely, lead. The Muzhiyev gold-mining deposit (Eastern Carpathians, Ukraine) belongs to such an ore deposits. A gravity concentrate with a galena upwards of 40% is produced at the operation. To avoid the formation of matte (melted sulfide), which is capable of dissolving a significant part of the gold during melting, the cleaner concentrate needs to be subjected to either an oxidizing roasting or some other process to remove the sulfur, arsenic, and antimony. Roasting, however, volatilizes sulfur and arsenic. The pyrometallurgical method for processing sulfide concentrates, therefore, has limits as to its application both economically and from the standpoint of environmental safety. Microbial oxidation of sulfide minerals (concentrates) is proposed as a viable alternative to the pyrometallurgical method. 127
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Tomizuka and Yagisawa [4] examined the role of bacteria and proposed a base scheme for galena leaching in the presence of T. ferrooxidans; the leaching mechanism includes the following reactions: 2PbS + 2H2SO4 + O2 → 2PbSO4 + 2H2O + 2S
(1)
(2) 2S + 2H2O + 3O2 → 2H2SO4 According to these authors, at low pH values sulfur can serve as the sole energy source for the autotrophic microorganisms. The elemental sulfur is released as a result of the electrochemical reaction with oxygen (reaction 1), and further, sulfur serves as a substrate for thiobacteria which produce sulfuric acid (reaction 2). The aim of this investigation was to establish whether it is possible to improve the kinetics and the process efficiency of galena microbial leaching from the gold-bearing ore of the Muzhiyev deposit by the addition of natural iron-bearing raw material (for example, magnetite). Attention was also paid to the electrosurface properties of thiobacteria and galena in the microbial leaching process. 2.
MATERIALS AND METHODS The middlings of gravity separation (MGS) of the gold-bearing ore from the Muzhiyev deposit, which was produced in a Knelson centrifugal concentrator, was investigated. The phase composition of this sample was determined by four main components: galena PbS (∼42%), barite BaSO4 (∼20%), pyrite FeS2 (∼6%), and quartz. The sample also contained individual gold grains. MGS was very similar to the concentrate and differed from it only by low gold content. The sample was ground -0.063 mm. The mixed culture, containing mainly T. ferrooxidans as well as T. thiooxidans, was used for the microbial leaching. The culture was cultured from the samples of the same deposit and was adapted to a 10% (w/v) galena pulp density for 10 weeks. The accumulated culture was grown in basal 9K medium. The experiments were conducted in 0.5-L shaker flasks with 10 g of MGS, 2.5-15 g of magnetite, 50 ml of cell biosuspension (inoculum)) and 100 ml of 9K medium without iron at pH 2.1. The stirring speed was 150 rpm. The magnetite sample contained 64.5% of Fe3O4; the moisture content of the product was 10.5%. The efficiency of the microbial leaching (%) was estimated from the sulfidesulfur content (% w/w) of the dried mineral sample before and after treatment by bacteria; the iodine method of analysis was used. For this purpose the sample was dissolved in HCl; released hydrogen sulfide was absorbed by ammonia solution of zinc sulfate. The precipitate formed was dissolved in the mixture of HCl and titrated iodine solution; the sulfur quantity was determined on the iodine excess in the solution. The Fe3O4 dissolution was assessed on the Fe2+ and Fe-total concentration by colorimetric method with ortophenanthroline. The PbS electrode for rest potential investigations was prepared in the following way. The side surface of a circular graphite electrode with a cross-sectional area 0.3 cm2 was embedded in epoxy insulating glue. The working face surface was first polished and then rubbed with a finely dispersed powder of the synthetic PbS. Excess powder was removed with distilled water. Each electrode was placed in a vessel with 9K medium and bacterial inoculum. Stirring in the vessels with the biosuspension was accomplished with a magnetic-stirrer. The rest potential measurement was made with each electrode using an Ag-AgCl reference electrode. Zero time was considered to be the "control" without the microorganism action (Table 2). 128
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The experiments on the influence of an electric field on the rate of leaching were carried out in the following way. Suspensions containing either MGS or synthetic PbS, the mixed thiobacteria culture, and 9K medium with 44.5 g/L iron (II) sulfate at pH 2.1 were prepared. All vessels containing the suspension were incubated on the shaker. Two vessels, one containing synthetic PbS and one containing MGS, served as controls; no electric field treatment was applied to the control vessels. The graphite electrodes were introduced into the other two vessels, and periodically an electric field with a voltage of 0.5 V/cm (anode potential was +0.22V C.S.E.) was applied. The duration of the electric field treatment was 5 min. followed by a pause of 10 min. The treatment by the discontinuous DC field was conducted in vessels with non-separated cathode and anode chambers. The electrophoretic mobility of T. ferrooxidans M1 was measured by the microelectrophoresis method in different buffers: in the citrate buffer of McIlven and in universal buffer mixture composed of phosphoric, acetic, and boric acids; these measurements were recalculated into the electrokinetic potential (ζ) using the Smolouchowsky formula. All the said experiments were carried out at least by duplicate. 3.
RESULTS AND DISCUSSION
3.1 Galena oxidation in mixed biomineral suspension As seen from data given in Table 1 the MGS sample placed in the iron-free 9K medium was slowly oxidised. The galena oxidation reaction in water is described by this equation:
PbS + 4H2O → PbSO4 + 8H+ + e (3) The formation of insoluble anglesite (PbSO4) promotes the reaction shift to the right. However, the formation of an insoluble film on the surface of the mineral particles decreased the kinetics of the oxidation process. As a result, about 21% of the sulfides was oxidized from the original mineral sample in iron-free 9K medium in nine days. The second sample, which contained thiobacteria in the iron-free 9K medium with the MGS sample, was significantly oxidized. In three days about 73% of the sulfide was oxidized; in 9 days 82.5% of sulfide was oxidized. As can be seen, thiobacteria significantly improved both the oxidation kinetics and degradation of the mineral. Chemically produced sulfur (equation 1) served as the substrate for the T. thiooxidans, which metabolized it to sulfuric acid (equation 2). The data given in Table 1 for the sulfide oxidation (item 2) indicate that galena can serve as the energy source for the mixed culture resulting in degradation of the galena-containing ore. That is a reason why the degree of the sulfide destruction in the second sample considerably exceeded that of the first sample that did not contain bacteria. The data on the change in rest potential of lead sulfide with an increase in bioleaching time is testimony to the changes in the sulfide surface composition and the electrochemical properties of the surface. Table 2 gives the values for the electrode made from natural galena [5] and the electrode made from the synthesized PbS from day 0 through 14 days of leaching with the thiobacteria. The interaction between the mineral and thiobacteria was accompanied by the increase in rest potential with respect to that measured in fresh nutrient medium. The positive potential increases with the bioleaching time. Comparing the galena mineral and PbS electrodes, it is seen that pure product was oxidized more intensively because it had higher positive potential values. Note, however, 129
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that the rest potentials for both the mineral and powder electrodes were not large. The reason for this phenomenon was discussed earlier; it is the formation of reaction products on the electrode surface diminishes oxidation. Table 1. The sulfide oxidation and Fe3O4 dissolution the mixed biomineral suspension No. and sample composition
S2S2- in the oxidised ore Sulfide oxidation (%) before (%) leaching In 3 days In 9 days In 3 days In 9 days (%)
Fe2+ /Fe3+
Fe3O4 dissol. (%)
1. MGS
5.6
5.2
4.4
7.2
21.5
-
-
2. MGS + bacteria
5.6
1.5
0.8
73.2
85.7
-
-
3. MGS + 2,5 g of Fe3O4
4.5
3.8
2.8
25.6
37.8
9.8
51
4. MGS + bacteria + 2,5 g of Fe3O4
4.5
0.6
0.5
86.7
89.0
4.0
65
5. MGS + bacteria + 5 g of Fe3O4
3.7
0.15
0.13
96.0
96.5
3.0
71
6. MGS + bacteria + 10 g of Fe3O4
2.8
0.31
0.30
88.9
89.3
6.6
50.2
7. MGS + bacteria + 15 g of Fe3O4
2.2
-
0.3
-
86.6
7.1
59
Table 2. Stationary potential of PbS-electrode in the 9K medium with microorganisms Bioleaching time, days 0 2 7 9 14
Rest potential of natural galena [5], mV (C.S.E.) 5 75 110
Rest potential of PbS-electrode, mV (C.S.E.) 27 68 97 140 159
Now we shall address the experiments in which magnetite was introduced simultaneously with thiobacteria. It is known that iron (II) as well as the reduced sulfur species can serve as an energy source for thiobacteria; this property is widely used for the microbial leaching of ores that contain sulfides with iron (e.g. pyrite, arsenopyrite, chalcopyrite, and pyrrotite) [6]. The majority of thiobacteria can use these substrates in their metabolism. The role of the added magnetite may iron (II) generation from the dissolution of Fe3O4. The autotrophic bacteria, in turn, form the oxidizing medium by generating the Fe (III) ions. Galvanic interactions among different minerals are known and sometimes used to promote the leaching of sulfide minerals [7-8]. The electrochemical concept of many galvanic leaching systems is the basis of these processes [9]. In our case the role of the added magnetite in promoting galena leaching from ore may be due to coupling the electrochemical reactions of magnetite reduction and galena oxidation. The cathodic reaction under acidic conditions could be: Fe3O4 + 6H+ + 2e → FeO + 2Fe2+ + 3H2O 130
(4)
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The anodic reaction of galena oxidation to sulfur and then to sulfate could proceed according to polysulfide mechanism [10]: PbS + 4H2O → PbSO4 + 8H+ + 8e (5) However, the run of such galvanic pair can be retarded due to the formation of sulfur film on surface of galena particles and to low Fe3+ concentration in solution. Indeed, the sulfur leaching from galena in the magnetite presence was higher as compared in the absence of this addition, but in the whole remained still low (item 3, Table 1). The availability of the autotrophic bacteria and their metabolites promote galena oxidation by providing the Fe (III) generation and increasing the oxidizing conditions in a bioleaching system in the following way: Fe2+ bacteria Fe3+ + e, (6) as well as accordingly equation 2. Natural magnetite in 0.5 M sulfuric acid has a wide range, 0.5-1.4 V (S.H.E.), in which complete passivation occurs. Increasing the current during the cathodic scans at more negative potentials of 0.5 to –0.1 V may be necessary for the process described by equation 4. The mineral suspension, which contained sulfide ore, magnetite and thiobacteria culture had a positive Eh (0.2-0.4 V). Therefore, oxidative destruction of galena as well as microbial sulfuric acid production are possible. Superimposing several processes can have a number of consequences. For example, if the cell concentration is not high and these mechanisms have competitive character, then the appearance of the new substrate (Fe2+) for autotrophic bacteria can result in the decrease of the sulfuric acid production by microorganisms. On the other hand, the appearance of the new substrate can result in the increase of the cell concentration and then an acceleration of the sulfur oxidation rate. The appearance of the Fe3+ also must be accompanied by the acceleration of the solubilization process and greater dissolution of sulfide. The results of Table 1 (items 4-7) show that after magnetite introduction in the suspension, which included MGS, bacterial inoculum, and iron-free 9K medium, galena dissolution increased considerably. In three days sulfide oxidation reached ∼86-90%; this level of oxidation was not achieved until day 9 in the suspension without magnetite (sample 2). A number of experiments with different magnetite content, ranging from 2.5 to 15.0 g in the suspension, have been carried out. The best result was achieved in tests with 4 and 5 g of magnetite. In these tests, the sulfide oxidation level reached ∼96% after three days; little change in the oxidation level was observed with additional incubation. The sulfide content in the solid phase, the total concentration of the dissolved iron, and the Fe2+ concentration in incubation solution were all measured allowing assessment of magnetite degradation. As noted in sample 5 (Table 1), the magnetite was subjected to the highest solubilization level (∼71%) with a corresponding high degree of galena oxidation. The maximum concentration of oxidized iron Fe3+ was also achieved in this test. The results obtained lead to the conclusion that the addition of 2.5-5.0 g of magnetite for up to 10 g of MGS positively influenced galena leaching from ore. Addition of this mixed iron oxide promoted oxidizing conditions because of iron sulfate generation. The introduction of a larger quantity of magnetite also positively influenced the kinetics and efficiency of the sulfide leaching in relation to the control. However, with greater magnetite addition the Fe3O4 degradation declined as well as the oxidized iron concentration. Hence, it is possible to conclude that the increase in the magnetite 131
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concentration in the suspension is not expedient; this would lead to a decrease in the redox-potential due to the cathodic process activation and will mean non-productive additive use. 3.2 Galena electro-bioleaching We also performed experiments on the influence of an electric field on the rate of the lead sulfide and galena-bearing ore bioleaching. Table 3 shows sulfide analysis results for the four experimental tests. Within two days those tests in which the electric field treatment was applied exhibited sulfide oxidation of 98.9% for the synthetic PbS and 90% oxidation for the MGS. After four days all tests, including the controls, were analyzed for residual sulfide. The difference was notable. In the controls the degree of synthetic galena oxidation was 87% and for MGS the sulfide oxidation was 78%; this is considerably less sulfide oxidation that was observed for the corresponding samples subjected to the field action (Table 3). Table 3. Microbial galena oxidation in field 0.5 V/cm Duration, days 2 4
Degree of the sulfide oxidation, % Without field In field MGS PbS MGS 90 78 87 99
PbS 98,9 99,5
We attempt to explain the results. Our previous experiments with iron alum (NH4Fe(SO4)2·12H2O) showed that during the DC electric field treatment in the electrochemical cell with non-separated electrodes, a decrease in Fe3+ ion concentration took place due to the cathodic processes. With an increase in the cathode potential the concentration of reduced iron rose. Thus, under the action of the DC electric field in the inoculated 9K medium, Fe3+ was continuously reduced, providing a substrate for microbial oxidation of the Fe2+ to Fe3+ (equation 6) [11]. The additional substrate stimulated bacterial growth. This, in turn, enhanced sulfide oxidation, resulting in increased dissolution of the lead sulfide and increased sulfuric acid formation. It appears that in the present experiments the system of coupling the electrochemical reactions is similar to that with magnetite, which was described earlier (section 3.1) in this paper. In the tests that employ non-separated electrodes, it is important not to apply too great of an applied potential in the negative direction; such an action could increase the Fe2+ concentration too much for the level of microbial activity. It is important to maintain a low voltage field for the biosuspension. The dynamics of the oxidation-reduction potential change indicates that at the beginning of the field action on the biosuspension (about 18 hours) the Fe3+ ions were regenerated in excess and the Eh potential increased to the extent expected by thiobacteria activity under the field action. It is likely that the acceleration of the galena oxidation in the biomineral suspension is related to the action of the discontinuous DC electric field. 3.3 Electrosurface properties Some investigations were undertaken to determine if the electrosurface properties of thiobacteria play any role in the interaction of the bacteria with galena and if these properties can be regulated in enhance the bioleaching efficiency [12]. The surface chemical and electrokinetic properties of galena were studied in detail by Pugh [13]; this study focused on establishing optimal conditions for the flotation separation of sulfide ores. Main attention was paid to the mineral properties under neutral and alkaline pH 132
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values corresponding to the flotation conditions. Only a few studies have been dedicated to the electrical and hydrophilic/hydrophobic properties of the thiobacteria surface [1416]. The correlation between the oxidation-reduction potential and Zeta-potential of the bacteria during leaching was determined: the minimum ζ-potential values of bacteria corresponded to the maximum values of the Eh. One peculiarity connected with the presence of lead sulfide in the cultivation medium was noted. The ζ-potential value of the bacterial cells grown on 9K medium was higher by 1-2 mV compared to the cells grown in the same medium with MGS added. This was possibly due to the presence of highly dispersed particles of PbSO4 or PbCO3, which do not possess a high electrokinetic potential [17]. During the electrocoagulation of these minerals with the cells the decrease in the electrokinetic potential of the latter may have taken place. The interaction between a mineral and a cell depends on the pH value. Fig. 1 shows the Zeta-potential dependence with pH for galena as observed by Pugh [13] and our data for thiobacteria. In an acidic, oxidizing medium (NaNO3) the synthetic galena was positively charged (curve 4); under the same acidic, oxidizing conditions the natural Swedish galena was negatively charged (curve 5) and the value of the electrokinetic potential at pH 2 to 2.5 was very close to the ζ values of the cells T. ferrooxidans M1 (curve 1-3). The difference in the surface properties of various samples of galena is associated with its surface oxidation products including PbSO4 and PbCO3. As seen in Fig. 1, in the interval of pH 2.5-3.75 the Zeta-potential values closely coincide in both buffer solutions (curves 1-3). In the citrate buffer the maximum Zetapotential value was achieved at pH 3.75, however, the charge was not changed even at pH 1.75. It is likely that for the given culture the isoelectric point (IEP) does not exist. For two of six T. ferrooxidans strains studied by Skvarla and Kupka [13] the IEP was also not achieved. It is likely that the pK value observed at 2.60-2.70 represents the mixed value of two pKs related to the dissociated phosphate groups of phosphatidic acid and carboxyl groups of gluconic acid in the composition of the lipo-polysaccharide cell wall [18]. Consequently, the cellular envelope of T. ferrooxidans M1 bacteria, grown in the medium containing iron as an energy source, consists of acidic material determined by phosphate and carboxyl groups. The number of these groups changes depending on the growth phase, the bacterial oxidative activity, and the intensity of the exchange processes on the membrane. Electro-bioleaching experiments may be a proof of this. In the tests described in Section 3.2 the bacterial cells were separated from the solutions, two times washed off in 5 N H2SO4, re- suspended in McIlven buffer, and after the mentioned preparation they were subjected to ζ-potentials measurement. As seen in Fig. 2, the cells separated from the suspension subjected to the DC field action had higher negative ζ-potential values compared to non-DC field-treated cells. The ζ-potential of the cells not subjected to the electric treatment during the initial incubation period decreased slightly at first and then increased. The cells grown under the Fe2+ ion electrochemical regeneration and electrostimulation conditions showed a stable increase of ζ-potential as the cells aged. On the 9th day of the incubation (approximately 216 hours), the electrokinetic potential increased nearly two times in both the DC field-treated and non-treated cultures. The electrokinetic potential of the thiobacteria changed with increases of electrolyte concentration at a constant pH. Changes in the electrokinetic potential could not be attributed to compression of the electric double layer with increase of the electrolyte concentration. As observed in Fig. 3 (curve 1), with an increase in the Fe (II) concentration from 10-6 to 10-2 M the negative potential increased by 5 mV, indicating the 133
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higher affinity between T. ferrooxidans cells and iron. However, the cells proved practically indifferent to added calcium ions (curve 3); an insignificant decrease in the ζpotential value was observed when Ca(NO3)2 was added over a broad concentration range. Lead increased the negative charge of the cell surface at high Pb(NO3)2 concentrations (10-3 – 10-2 M). This same increase in the negative charge was not reached in pulp during galena leaching due to the extremely low solubility of the lead sulfate that was formed (curve 2). The aforementioned property of the T. ferrooxidans culture may indicate that the bonding locations of Fe (II) on the cell surface are not accessible to other cations and that its oxidation mechanism is protected from metals in the surrounding medium [19]. Figure 1. Zeta-potential versus pH plots of T. ferrooxidans M1 (1-3) and galena (4-6). The experimental conditions: 1 – growing in 9K medium with MGS (measurements in universal buffer mixture); 2 – growing in 9K medium, (measurements in McIlven buffer); 3 – growing in 9K medium with MGS (measurements in McIlven buffer); 4 – PbS in 2x10-3 M NaNO3 solution; 5 – natural Swedish galena in 2x10-3 M NaNO3 solution; 6 – natural Swedish galena in 2x10-3 M NaNO3 solution plus 1x10-5 M Pb(NO3)2. Curves 4-6 according to Pugh and Bergström [13]
Figure 2. Zeta-potential versus incubation time plot of T. ferrooxidans M1 bacteria in McIlven buffer. Previously incubated under the 0.5 V/cm DC field treatment (1); and without the DC field treatment (2)
134
Figure 3. Zeta-potential versus electrolyte concentration plot of T. ferrooxidans M1 bacteria dispersed in FeSO4 (1), Ca(NO3)2 (2), and Pb(NO3)2 (3). The ζ-potential value without electrolyte addition was –13.5 mV
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5.
CONCLUSIONS 1. The addition of magnetite to lead sulfide in the ratio of 1-2:4 in the microbial leaching system enhanced the efficiency and kinetics of sulfide degradation. 2. The thiobacteria substantially changed the galena surface, surface composition and increased the mineral rest potential during the leaching study. 3. Application of a discontinuous DC electric field with low voltage application to the galena bioleaching system improved the kinetics of the sulfide degradation. 4. The thiobacteria cell and the natural galena surfaces possess the same electrokinetic potential signs but low in value, which promotes their heterocoagulation. The electrokinetic potential of the thiobacteria changes during galena leaching and under the action of the DC electric field.
REFERENCES
1. G.S. Hansford and T.Vargas, Hydrometallurgy, 59 (2001) 135 2. M.C. Fuerstenau, C.O. Neto, B.V. Elango, and K.N. Han, Met. Trans. B, 18 B (1987) 25. 3. A.E. Torma and K.N. Subramanian, Int. J. Miner. Process., 1 (1974) 125. 4. N. Tomizuka and M. Yagisava, in: L.E. Murr, A.E. Torma and J.A. Brierley (Eds.), Metallurgical Application of Bacterial Leaching and Related Microbial Phenomena, Academic Press, New York, 1978, pp. 321 - 344. 5. J.L. González-Chaves, F. González, A. Ballester, and M.L. Blazquez, Minerals and Metallurgical Processing, 17 (2000) 116. 6. H.L. Erlich, Hydrometallurgy, 59 (2001) 127. 7. R.K. Paramguru and B.B. Nayak, Metals and Materials Processes, 6 (1996) 23. 8. R.K. Paramguru and B.B. Nayak, Electrochem. Soc., 143 (1996) 3987. 9. N. Jyothi, K.N. Sudha and K.A. Natarajan, Int. J. Miner. Process., 127 (1989) 189. 10. A. Schippers and W. Sand, Appl. Environ. Mocrobiol., 65 (1999) 319. 12. V.I. Podolska, A.I. Ermolenko, L.N. Yakubenko, and Z.R. Ulberg, Colloid Journal, (in press). 13. R. Pugh and L.Bergström, Colloids Surfaces, 19 (1986) 1. 14. J. Skvarla and D. Kupka, Biotecnol. Techniques, 12 (1996) 911. 15. J. Skvarla, D. Kupka, A. Naveshakova and A. Skvarlova, Folia Microbiol., 47 (2002) 218. 16. M. Misra, K. Bukka and S. Chen, Minerals Engineering, 9 (1996) 157. 17. J. Bebie, M. Schoonen and D. Strongin, Geochim. Cosmochim., 62 (1998) 633. 18. W. Sand, T. Gehrke, P. Jozsa and A. Schippers, Hydrometallurgy, 59 (2001) 159. 19. O. Tuovinen and D. Kelly, Z. Allg. Mikrobiol., 12 (1972) 311.
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Bioleaching of electronic scrap material by Aspergillus niger W.K.Ten and Y.P.Ting* Department of Chemical and Environmental Engineering, National University of Singapore, Kent Ridge Crescent, Singapore 119260 Abstract This work reports on the bioleaching of electronic scrap material (ESM) from a local waste recycling company. The most abundant elements present in the ESM dust were oxygen and silicon, followed by various base metals with concentration exceeding 10,000 mg/kg. Precious metals gold, silver and palladium were found in trace amount (<1,000 mg/kg). The fine ESM particles showed a heterogeneous matrix and a low specific area. Bioleaching experiments were carried out using Aspergillus niger, at various ESM pulp densities (0.1%-2.0% w/v). Using a two-step bioleaching process, the fungus was able to grow at up to 1.0% w/v, with the optimum pulp density at 0.1% w/v, under which about 35% Sn, 65% Pb, Zn, Al, and Mn and more than 70% Fe, Ni, and Cu were mobilized. It was also observed that various organic acids produced by the fungus during the leaching process paralleled the leaching of the heavy metals. Higher pulp density led to a decrease in acids produced due to the inhibitory effect of the toxic metals. The use of spent medium resulted in higher metal leaching efficiency than the two-step bioleaching at all pulp density (with the exception of Fe and Al at 0.1% w/v), due presumably to higher concentration of citric acid and lower concentration of oxalic acid in the spent medium which enhanced metals dissolution. Metals solubilization in spent medium leaching was not attributed to extracellular enzymes, but was mainly due to the action of the organic acids. Chemical leaching confirmed that citric, gluconic and oxalic acids were the responsible leaching agents in the bioleaching processes in removing heavy metals from the ESM. Compared with chemical leaching at 0.1% w/v, A. niger achieved similar leaching efficiency for Al, and was more efficient in the extraction of Fe, Sn and Au.
Keywords: Bioleaching, electronic scrap materials, Aspergillus niger 1.
INTRODUCTION Bioleaching may be described as an interaction between metals and microorganisms that causes the solubilization of metals. This process is based on the ability of microorganisms to transform solid compounds, and result in soluble and extractable elements which could be recovered [1]. Microbial leaching of metals may have been practiced as early as the 15th century, but the role of microorganisms in the leaching process was more clearly defined only from 1947 when bacterial catalyzation of iron oxidation and sulfuric acid formation in mine waters was demonstrated [2]. Microbial *
Corrresponding author:
[email protected]
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leaching technologies have gained more attention and have been used on an industrial scale for the recovery of copper, gold, uranium and zinc from low-grade ores or from lowgrade mineral resources [2]. Electronic waste, a new emerging and fast-growing waste stream, could be considered as an "artificial ore" due to the presence of intrinsically valuable heavy metals [2]. Recycling of the waste confers two possible benefits: removal of the potentially hazardous substances present in the waste, as well as the recovery of valuable metals in the waste. The use of bioleaching for the treatment of solid wastes may provide a more economic and environmentally friendly alternative to conventional technologies in the recover of the elements. A variety of microorganisms are known to be capable of carrying out the metal mobilization processes. This includes chemolithoautotrophic bacteria, heterotrophic bacteria, chemolithoautotrophic archea and heterotrophic fungi. Among the heterotrophic fungi, the genera Aspergillus and Penicillium are the most important microorganisms used in bioleaching [3]. Of these, Aspergillus niger and Penicillium simplicissimum are probably the most widely used. Metal leaching by heterotrophic microorganisms generally involves an indirect process with microbial production of organic acids, amino acids and other metabolites. Four mechanisms have been identified [4]: (i) acidolysis (ii) complexolysis (iii) redoxolysis and (iv) bioaccumulation. Metal solubilization in fungal bioleaching can also occur via the production of enzymes, e.g. phosphatases, which can solubilize metal phosphates [5]. The purpose of this study was to determine some aspects of the physical and chemical properties of electronic scrap material (ESM) and investigate the use of heterotrophic filamentous fungi A. niger in the leaching of heavy metals from the ESM under different pulp densities and bioleaching conditions. Commercial inorganic acids were also used to leach ESM and the results were compared with bioleaching. 2.
MATERIALS AND METHODS
2.1 ESM The ESM sample was collected from a local waste recycling company specializing in electronic waste. The ESM collected, in dust-like form, was generated during shredding and other separation processes in the mechanical recycling of the waste. The as-received ESM was screened with a series of testing sieves and the fraction < 0.212 mm was used in the experiments. 2.2 Characterization of ESM The composition of the ESM was determined through acid digestion/ICP-AES. The sample was acid digested using aqua regia/hydrogen fluoride/hydrogen peroxide in a microwave oven following the method of Das et al. [6]. The result was compared with the elemental determination through Energy-Dispersive X-ray Fluorescence Spectrometer (XRF) and Scanning Electron Microscope-Energy Dispersive X-ray (SEM-EDX). pH buffering capacity of the ESM was measured according to the method of Chandler et al. [7] and Crawford [8]: individual 5g samples of ESM were mixed in 30 ml nitric acid of varying strength and agitated for 48 hours before monitoring of the pH. The particle size distribution of the ESM was determined using Particle Size Analyzer (Coulter LS 230). The specific gravity of ESM was determined using specific gravity bottle (Bibby). Brunauer-Emmett-Teller (BET) multipoint method (Quantachrome, Nova 3000) was used 138
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to obtain the specific surface area of the ESM. Image of the ESM was obtained using a Scanning Electron Microscope (Jeol JSM-5600LV). 2.3 Microorganisms and spore inoculum preparation A. niger used in this study was supplied by Dr H. Brandl (University of Zürich, Switzerland). The fungus was maintained on potato dextrose agar (PDA) in petri dishes, incubated at 30°C for 7-10 days. Spores were harvested from the PDA surface by suspension in sterile deionised water. A haemocytometer was used to enumerate the number of spores. 1 ml of spore suspension (approximately 1 x 107 spores) was then added to 100 ml of culture medium with the following composition (g/l): sucrose (100), NaNO3 (1.5), KH2PO4 (0.5), MgSO4.7H2O (0.025), KCl (0.025), yeast extract (1.6). No pH adjustment was made for the medium. The cultures were incubated in a water bath with rotary shaker at 30°C and 120 rpm. 2.4 Bioleaching procedure Bioleaching experiments were performed at 0.1%, 0.5%, 1.0% and 2.0% w/v of ESM under two-step bioleaching and spent medium leaching. In two-step bioleaching, the fungus was incubated for two days before the ESM was added into the culture. In spent medium leaching, the ESM was added to the filtered cell-free spent medium wherein the fungus had been incubated for 18 days. Two control experiments (fresh medium and deionised water leaching) were also set up under identical incubation conditions. At intervals, samples were withdrawn from the flasks for the analyses of pH, sugars concentration, acids concentration, metals concentration and the biomass. The experiments were terminated after 18-26 days. 2.5 Chemical leaching Chemical leaching of ESM was carried out at a concentration of 0.1% w/v pulp density, with 100 ml of different molarities of individual sulphuric and nitric acids (30 mM, 50 mM and 100 mM) as well as a mixture of commercial citric, gluconic and oxalic acids at equal molarity with the acids biogenically-produced in bioleaching processes. Samples were withdrawn after 10 days incubation and analyzed for soluble metals. 2.6 Analytical methods The heavy metals in the solutions were quantified by Inductively Coupled Plasma Atomic Emission Spectroscopy (Perkin Elmer Optima 3000V). Glucose and sucrose were analyzed using a Biochemistry Analyzer (YSI 2700). The concentrations of the organic acids and fructose were determined by High Pressure Liquid Chromatography (Hewlett Packard 1100 Series). The separation was carried out with an Aminex HPX-87H cation exchanger column (Bio-rad); mobile phase: 5 mM H2SO4; injection volume: 20 µl; flow rate: 0.4 ml/min; temperature: 30°C and detected using UV at 210 nm and Refractive Index (RI) detector. For biomass determination, the culture broth with ESM was dried at 80°C for 24 hours, followed by ashing at 500°C for 4 hours; the biomass was determined gravimetrically. 3.
RESULTS AND DISCUSSION
3.1 Characterization of ESM Table 1 shows the elemental composition of ESM determined by acid digestion/ICPAES, XRF and SEM/EDX. The most abundant elements in ESM are oxygen and silicon, 139
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comprising 80,925 mg/kg and 64,625 mg/kg, respectively. It is known that refractory oxides typically constitute about 30% of electronic waste and the main component of refractory oxides is silica. Base metals, such as Fe, Cu, Pb, Al, Sn and Zn fall in the major element group (>6,000 mg/kg); most of these elements were found at a concentration higher than 10,000 mg/kg based on the result of digestion/ICP-AES and XRF analyses. Compared to the analyses of electronic scrap by Brandl et al. [9], the major elements found (Al, Cu, Pb, Sn and Zn) were similar with the ESM used in this work. Precious metals such as Ag, Au and Pd were found at lower concentration (<1,000 mg/kg). All subsequent metal analyses were performed using ICP-AES and calculations of metals leaching efficiencies were based on the digestion/ICP-AES results. The ESM was found to have a pH buffering capacity at a range from 4.6 to 1.2. This shows that the ESM is reasonably well buffered and can moderately resist changes in pH. Table 1. Elemental Analysis of ESM Element (mg/kg) Si
Acid digestion/ ICP-AES
Average 64625.00 RSD (%) 2.40 Cu Average 49600.00 RSD (%) 3.06 Fe Average 36000.00 RSD (%) 6.82 Al Average 27062.50 RSD (%) 3.14 Pb Average 22600.00 RSD (%) 1.43 Zn Average 21688.00 RSD (%) 2.18 nd: not detected; nt: not tested
XRF
SEM/EDX
100000.00 1.47 33000.00 0.42 19800.00 0.51 9000.00 4.25 20000.00 0.92 19500.00 0.66
16625.00 22.72 7200.00 37.46 9700.00 33.84 nd 6650.00 39.34 9875.00 15.61
Element (mg/kg) Sn Au Ni Mn Ag O
Acid digestion/ ICP-AES
XRF
Average 11212.50 60000.00 RSD (%) 6.91 6.79 Average 876.25 nd RSD (%) 4.38 Average 745.00 1200.00 RSD (%) 4.21 3.35 Average 575.00 1000.00 RSD (%) 7.68 7.02 Average 475.63 nd RSD (%) 3.18 Average nt nd RSD (%) -
SEM/EDX 10400.00 28.56 nd nd nd nd 80925.00 11.43
Using the Particle Size Analyzer, it was found that particles with the size range 10200 µm constitute the major portion of the ESM, with a mean at 64.43 µm. The fine ESM has a low specific surface area (1.95 m2/g) probably due to its non-porous structure. The SEM micrographs revealed a significantly heterogeneous matrix, with both rough and smooth surfaces, as well as particles of different shapes (Figure 1). Nevertheless, the ESM particles are typically smooth rods and spheres with many fine crystals and condensed flakes attached on its surface.
Figure 1. SEM micrographs of ESM (at 500 times magnification)
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3.2 Bioleaching of heavy metals from ESM Results showed that the fungus was able to grow at an ESM pulp density of up to 1.0% w/v in two-step bioleaching. The inhibition of growth at 2.0% w/v was due to increased concentration of toxic metals in the ESM. Figure 2 illustrates the growth of A. niger, at 0.1% w/v. Glucose and fructose were consumed simultaneously by the fungus, and the biomass increased over time and attained a maximum of 20.3 g/l. At higher pulp densities, slower rate of sugar consumption and biomass formation were observed (data not shown). For instance, a longer lag phase in biomass production (5-18 days) was noted at 0.5% and 1.0% w/v pulp densities. The toxicity of the metals present in the ESM was evident, even at a low concentration of 0.5% w/v.
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Figure 2. Growth of A. niger at 0.1% w/v ESM
During bioleaching, pH decreased from an initial value of 3.0 to 2.30 at the end of incubation, and revealed that the amount of protons produced exceeded the demand for the solubilization reaction. In spent medium leaching however, pH marginally increased along the incubation period for all the pulp densities tested (ranging from 2.52-2.71; see Figure 3). The increase in pH was due to the reaction between organic acids and the ESM that consumed protons during the leaching process. The result suggested that acidolysis played an important role in the leaching which converts the insoluble metal compounds to soluble metal salts. The control tests showed relatively constant pH over the acidic range at 5.05.6 during the leaching process. In the two-step bioleaching at 0.5% and 1.0% w/v, the pH remained constant as the fungus grew under these pulp densities (Figure 4). This may be attributed to the buffering capacity of the ESM. Table 2 summarizes the concentrations of the organic acids in the two-step and spent medium leaching. Compared with spent medium leaching, higher pulp density led to a lower acid production by the fungus. A significant decrease in citric acid was observed when pulp density increased from 0.1% to 0.5% w/v. This is due to the presence of increased manganese in the medium, which inhibited the biosynthesis of the acid even though low pH of the medium favors its production. The strong inhibition effect of manganese on citric acid synthesis has earlier been reported by Rőhr and Kubicek [10]. Similar to the two-step bioleaching, the 18-day spent medium contained citric acid as the dominant leaching agent, followed by small amounts of gluconic and oxalic acids. The low pH and deficiency of manganese in the medium led to an accumulation of citric acid in the pure fungal culture, as has been reported by others [6, 11, 12].
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4
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Figure 4. pH profiles in two-step bioleaching under various pulp densities
Figure 3. pH profiles in spent medium leaching under various pulp densities
Table 2. Organic acids in two-step bioleaching and spent medium leaching Organic acid (mM)
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1.0% w/v
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75.0
13.3
19.7
57.7
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31.2
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Tw o-step bioleaching Fresh medium leaching
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Figure 5. Metals leaching efficiency in various leaching processes at 0.1% and 0.5% w/v
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Figure 5 shows the leaching efficiency of various leaching processes at 0.1% and 0.5% w/v. The optimum pulp density for bioleaching for nearly all the heavy metals investigated was 0.1% w/v. Under two-step bioleaching at 0.1% w/v, 90-100% Fe, 6065% Al and Mn and 30-40% Sn removal were achieved. In general, metal extraction yields decreased as pulp density increased; an increase in pulp density to 0.5% w/v, led to a significant decrease in leaching efficiency. High leaching efficiencies (70-90%) for Ni, Cu and Pb were also obtained under the optimum pulp density. On the other hand, spent medium was able to leach Ni (75-95%), Cu, Pb, Zn and Mn (50-80%), Fe (5%-75%) and Sn (15-40%), depending on the pulp density. As a comparison, spent medium leaching generally gave higher metal extraction efficiencies than the two-step bioleaching at all pulp densities, with the exception of Fe and Al at 0.1% w/v. The spent medium contained higher concentrations of the organic acids than the two-step bioleaching at 0.5% w/v and 1.0% w/v (see Table 2). The lower pH of spent medium at these two pulp densities may have contributed to the higher leaching efficiency since a low pH is important in maintaining the availability of metal ions in the solution [13]. The higher concentration of oxalic acids in two-step bioleaching, compared with the spent medium, led to the formation of insoluble oxalate complexes and reduced metal leaching efficiencies. Fe and Al are exceptions, since the remarkable selectivity of oxalic acid for these metals and the solubility of the metal complexes may have contributed to its higher leaching yields [14, 15]. In the two-step bioleaching, the production of the organic acids and the concentration of the heavy metals increased with the growth of the fungus. Using Fe and Al at 0.1% w/v pulp density as examples, Figure 6 illustrates the parallel increase in metal leachability and the concentration of the organic acids, and shows that the leaching efficiency of the heterotrophic microorganisms depends on the extent of the production of organic metabolites. The organic acids have the dual effect on increasing metal solubilization by lowering the pH and complexing the metals into soluble organo-metallic complexes [16]. In spent medium leaching, the concentration of the heavy metals reached a plateau after a time period of 1-2 days of leaching. In order to determine if extracellular enzymes were involved in the spent medium leaching, a further experiment was conducted. ESM (0.5% w/v) was subjected to leaching by spent medium (without autoclaving), as well as autoclaved spent medium (at 121°C for 20 min.). As can be seen from Figure 7, similar metal extraction yields for both processes were noted for all the metals tested. This confirmed that the metal solubilization in spent medium leaching was not due to action by extracellular enzymes, but may be attributed to certain metabolites (principally organic acids) contained in the medium. 3.3 Comparison between chemical leaching and biological leaching As Figure 8 shows, the metal leaching efficiency of commercial organic acids was of the same order of magnitude as those obtained with biogenically produced organic acids of A. niger. The results confirm that citric, gluconic and oxalic acids were dominant leaching agents in the leaching processes, which effected heavy metals dissolution from the ESM. Table 3 shows the efficiency of bioleaching compared to chemical leaching using sulphuric and nitric acids. Both inorganic acids were able to completely solubilize Ni, Cu and Zn in the ESM. Lower Pb extraction by sulphuric acid compared to nitric acid was due to low solubility of the PbSO4 formed [17]. Chemical leaching showed a significantly higher leaching yield than bioleaching of A. niger in the case of Ni, Cu, Pb and Zn. Complete solubilization of these heavy metals were achieved in chemical leaching, compared to 40-90% extraction in biological leaching. The 143
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only exception is low Pb extraction by sulphuric acid (as explained earlier). A. niger was found to be more efficient in solubilising Fe (76-92%) due to the production of oxalic acid that favors the metal leaching. Together with Fe, higher extraction of Sn and Au were also achieved in bioleaching. The leaching efficiency of Al was generally similar for chemical and biological leaching. However, both leaching processes were unable to efficiently leach Ag from the ESM; a low extraction yield of less than 10% was obtained.
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Gluconate
Figure 6. Organic acids and metals leached in two-step bioleaching (0.1% w/v)
Native spent medium
Autoclaved spent medium
Figure 7. Metals leaching efficiency of native and autoclaved spent medium (0.5% w/v)
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Element Tw o-step bioleaching
Chemical leaching
Figure 8. Comparison of metals leaching efficiency between (a) spent medium and (b) two-step bioleaching and chemical leaching at equal molarity of mixture acids (0.1% w/v)
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Table 3. Metals leaching efficiency between bioleaching and chemical leaching (0.1% w/v) Metal Extraction (%) Two-step bioleaching
Spent medium leaching
Sulphuric acid (mM) 30
50
Nitric acid (mM)
100
30
50
100
Fe
91.9
76.4
63.8
72.1
84.0
16.9
25.3
34.9
Ni
72.1
94.3
100.0
100.0
100.0
100.0
100.0
100.0
Cu
78.2
75.4
100.0
100.0
100.0
100.0
100.0
100.0
Ag
0.0
7.4
4.7
4.1
2.6
15.8
9.5
5.1
Pb
65.5
48.2
31.5
32.5
25.9
100.0
100.0
100.0
Al
62.8
43.1
52.7
63.0
75.0
30.3
34.3
45.5
Zn
57.9
81.6
100.0
100.0
100.0
100.0
100.0
100.0
Mn
64.4
67.5
82.7
79.3
83.4
78.1
67.7
68.2
Sn
36.0
32.7
0.0
1.4
1.5
0.6
0.0
2.3
Au
12.9
15.4
0.0
0.0
1.0
0.3
1.3
0.9
4.
CONCLUSIONS The ESM used in this study is a fine particle with low specific surface area due to its non-porous structure. It contains high concentration of base metals and trace amount of precious metals. The optimum pulp density for the bioleaching was 0.1% w/v. Under this condition, 75-94% Fe, Ni and Cu and 45-67% Pb, Zn, Al and Mn were solubilized. Except for Fe and Al, spent medium leaching was more efficient than two-step bioleaching in removing heavy metals from ESM. The metal leaching efficiency of chemical and biological leaching varies according to the heavy metal being extracted, the type of acid used and the leaching process. Metals solubilization in spent medium leaching was not attributed to extracellular enzymes, but was mainly due to the action of the organic acids. Chemical leaching confirmed that citric, gluconic and oxalic acids were the responsible leaching agents in the bioleaching processes in removing heavy metals from the ESM.
ACKNOWLEDGEMENTS This work was funded by the National University of Singapore research grant R-279000-059-112. The authors thank Dr H. Brandl (University of Zürich, Switzerland) and Citiraya Industries Pte. Ltd. for providing the A. niger and the electronic scrap materials respectively. REFERENCES
1. Krebs, W., C. Brombacher, P.P. Bosshard, R. Bachofen and H. Brandl. Microbial Recovery of Metals from Solids, FEMS Microbiology Reviews, 20, pp. 605-617. 1997. 2. Brombacher, C., R. Bachofen and H. Brandl. Biohydrometallurgical Processing of Solids: a Patent Review, Appl Microbiol Biotechnol, 48, pp. 577-587. 1997. 3. Bosecker, K. Bioleaching: Metal Solubilization by Microorganisms, FEMS Microbiology Reviews, 20, pp. 591-604. 1997. 145
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4. Burgstaller, W. and F. Schinner. Leaching of Metals with Fungi-Minireview, Journal of Biotechnology, 27, pp. 91-116. 1993. 5. Morley, G.F., J.A. Sayer, S.C. Wilkinson, M.M. Gharieb and G.M. Gadd. Fungal Sequestration, Mobilization and Transformation of Metals and Metalloids. In Fungi and Environmental Change, ed by J.C. Frankland, N. Magan and G.M. Gadd, pp. 235256. New York: Cambridge University Press. 1996. 6. Das, A. K., R. Chakraborty, M.D.I. Guardia, M.L. Cervera and D. Goswami. ICP-MS Multielement Determination in Fly Ash after Microwave-Assisted Digestion of Samples, Talanta, 54, pp. 975-981. 2001. 7. Chandler, A.J., T.T. Eighmy, J. Hartlen, O. Hjelmar, D.S. Kosson, S.E. Sawell, H.A. Van der Sloot and J. Vehlow. Studies in Environmental Science 67: Municipal Solid Waste Incinerator Residues. pp.235-236. The Netherlands: Elsevier Science. 1997. 8. Crawford, J. A Model of pH and Redox Buffer Depletion in Waste Landfills. Ph.D Thesis. Royal Institute of Technology, Stockholm, Sweden.1999. 9. Brandl, H., R. Bosshard and M. Wegmann. Computer-Munching Microbes: Metal Leaching from Electronic Scrap by Bacteria and Fungi, Hydrometallurgy, 59, pp. 319326. 2001. 10. Rőhr, M. and C.P. Kubicek. Regulatory Aspects of Citric Acid Fermentation by Aspergillus niger, Process Biochemistry, pp. 34-37. 1981. 11. Bosshard, P.P., R. Bachofen and H. Brandl. Metal Leaching of Fly Ash from Municipal Waste Incineration by Aspergillus niger, Environ. Sci. Technol., 30, pp. 3066-3070. 1996. 12. Rőhr, M., C.P. Kubicek, and J. Kominek. Citric Acid. In Biotechnology, Vol. 3, ed by H. Dellweg, pp. 423-430. Basel: Verlag Chemie. 1983. 13. Gadd, G.M. and A.J. Griffiths. Microorganisms and Heavy Metal Toxicity, Microbial Ecology, 4, pp. 303-317. 1978. 14. Ambikadevi, V.R. and M. Lalithambika. Effect of Organic Acids on Ferric Iron Removal from Iron-stained Kaolinite, Applied Clay Science, 16, pp. 133-145. 2000. 15. White, C., J.A. Sayer and G.M. Gadd. Microbial Solubilization and Immobilization of Toxic Metals: Key Biogeochemical Processes for Treatment of Contamination, FEMS Microbiology Reviews, 20, pp. 503-516. 1997. 16. Bhattia, T.M, and T. Yasminb. Bioleaching of Uranium from Sandstone Ore by Aspergillus niger. In Biohydrometallurgy “Fundamentals, Technology and Sustainable Development”, Part B: Biosorption and Bioremediation, ed by V.S.T. Ciminelli and Jr. O. Garcia, pp. 651-660. New York: Elsevier. 2001. 17. Bock, M., K. Bosecker and R. Winterberg. Bioleaching of Soils Contaminated with Heavy Metals. In Biodeterioration and Biodegradation: Paper of the 10th International Biodeterioaration and Biodegradation Symposium, 1996, New York: VCH Verlagsgesellschaft, pp. 639-644.
146
15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Bioleaching of metallic sulphide concentrate in continuous stirred reactors at industrial scale – Experience and lessons D. Morina, P. d’Huguesa and M. Mugabib a
BRGM, 3 Av. Claude Guillemin, BP 6009 45060 Orléans Cedex 2, France b KCCL, Kasese road, PO Box 524, Kasese, Uganda
Abstract From laboratory to industrial size, scale-up of bioleaching of sulphide concentrate in continuous stirred reactors derives from the properties of the reactions and of the medium as measured at lab-scale. Literature is abundant in emphasising the influence of the main operating parameters of this process from small-scale testwork, and provides modelling approaches for simulation and performance predictions. Experience gained in real case situation shows that a very critical aspect is the control of the performances and how to evaluate the efficiency of the process on real time. Gas analysis is applicable to performance control and can be integrated to the process control system. Even if such a method gave relative values in the case mentioned in this paper because of the difficulties of gas sampling, it is a reliable way to follow up bacterial activity on real time, which can be improved to give accurate figures. The operating ranges of parameters like pH, air flowrates, temperature, etc. are reviewed and discussed by comparing information obtained from literature, laboratory testwork, and industrial practice. Eventually, indications are provided concerning the information that a practical model should give as relevant to process operation at industrial scale and questions that remain unanswered although they have a key influence on process performances and costs.
Keywords: bioleaching, sulphide concentrate, pyrite, process control 1.
INTRODUCTION Industrial-scale bioleaching has known a relatively rapid growth in the last twenty years with the start-up of about ten plants using bioheap leach, seven plants treating sulphide concentrates in agitated tanks [1] and countless pilot-scale operations. The process is reliable, efficient, simple, user-friendly and safer for the environment than other conventional techniques. Bioleaching is known to be a more cost-effective process to the extent that the very conservative decision-makers in mining and metallurgy have dared to invest in industrial-scale bioleach operations. However, successful operations do not mean that there are no technical difficulties or requirements for further improvement. Imaginative engineering addresses the main concept issues of design. However, the operators’ experience is invaluable in highlighting the pros and cons of a technique and it is a source of information that cannot be ignored. 147
Bioleaching Applications
In that context, this paper mainly aims at giving a practical examination of some of the critical technical points of the bioleach treatment of a sulphide concentrate in continuous stirred tank reactors (CSTR) at industrial scale. The experience of bioleach treatment at this scale presented in this paper refers to the Kasese cobalt producing plant operation in Uganda [2]. The production operation started in June 1999 and was shut down in August 2002 as a result of the persistently low cobalt prices. 2.
EXPERIMENTAL The Kasese Cobalt Company Limited (KCCL) bioleach plant was designed to treat 10.2 t per hour of a cobaltiferous sulphide concentrate. The concentrate contains average grades of 80% pyrite and 1.4% cobalt. Cobalt is finely disseminated in the pyrite crystalline matrix. Copper, and nickel also occur as sulphides at low grades; 0.14% and 0.12% respectively. The non-sulphide components are silicates. No carbonates were found in the concentrate. The design and operating characteristics of the bioleach unit are summarised as follows:
•
A primary stage with three tanks, and secondary and tertiary stages with one tank each. Launders and gravity transfer ensured pulp flowing from one stage to the next.
•
Every tank had the same total operating volume of 1,380 m3. Tanks were made of stainless steel 304L (BS).
•
Mixing and aeration in the tanks was provided by a rotating shaft system, equipped with two upper impellers and one bottom turbine. Air was injected under the turbine. The turbine was a disc with flat blades on the lower face. The size and layout of the blades were designed to provide the required air dispersion with the least energy consumption possible. About 60% of the energy for agitation is used by the turbine rotation. Contrary to what was observed on a small scale, the air injection does not reduce this value. This system was designed, scaled up and supplied by Robin Industries, France. BRGM tested and provided technical assistance to the design of this system called BROGIM®.
•
Air supply was provided by five blowers, which could supply up to 20,000 Nm3/h air to every tank. The operating airflow rates in the tanks were in the range of 10,000 to 15,000 Nm3/h in the primary stage reactors and between 5,000 and 10,000 Nm3/h in the secondary and tertiary stage reactors. Flowmeters on every air feed pipe provided airflow rates in Nm3/h.
•
Nominal temperature was 42°C. Heat removal was ensured by internal stainless-steel cooling coils connected to a cooling tower
•
pH was kept constant in every tank by a continuous addition of limestone slurry at a controlled rate. The nominal values were 1.4 to 1.5 and 1.6 to 1.8 respectively in the primary tanks and in the secondary/tertiary tanks.
•
Solids concentration in the feed was 20% (solids wt / pulp wt). The follow-up of the performances in the tanks was ensured by cobalt and sulphur assays in liquids and solids. Cobalt in solution was determined by atomic absorption after dilution in acidic solutions. Cobalt in solids was obtained from cobalt concentration in solution after acid digestion of the solids. Sulphate analysis by barium sulphate gravimetry in liquids and after hydrochloric acid digest for the solids provided the sulphide oxidation rate. The proportion of elemental sulphur has always been quantitatively negligible. 148
Bioleaching Applications
Another way of measuring the performances was to analyse the composition of the off-gas of the bioleach tanks. The method consists of sampling the off-gas from the tanks by means of a vertical 13-m pipe with open slots at different levels. The pipe is fixed in the tank, near the tank wall and its length goes down to about half a meter from the bottom. Assuming that the tank is perfectly mixed, the gas taken by this way was supposed to be representative of all the off-gas. The sampling well was also used to measure the oxygen concentrations in the slurry (using an oxygen probe) at the level of the open slots. Gas taken from each tank was driven through a network of pipes by a suction pump to a dewatering device and then to two analysers. The oxygen and carbon dioxide concentrations in the inlet and outlet gas of each reactor were measured using a paramagnetic analyser and an infrared analyser respectively (ADC 7000 Gas Analysers). Mass balance of oxygen allowed determination of the oxygen uptake rate for each tank in milligrams of oxygen per litre of slurry per hour by assuming the nitrogen flow in the air fed to the reactor to remain unaffected in the off-gas. It also takes into account the variation of carbon dioxide concentration in the off-gas due to the dissolution of limestone used for pH control. The gas measurements were carried out once a day. The Oxygen Uptake Rate, OUR, in mg.L-1.h-1 from the values of the flow rates of oxygen in mol.L-1 can be expressed as: 32 OUR = n OI 2 − n OO 2 1000xVo (1)
(
)
where: -
n OI 2
is the inflow rate of O2, mol.h-1
-
n OO 2
is the outflow rate of O2, mol.h-1
VO is the operating volume of slurry in the tank considered, m3. This volume is about 1300 m3 depending on the airflow rate and the gas hold-up in the tank. - 32 is the molar mass of O2, g Then, -
n
I O2
= 1000 ×
QOI 2 22.414
(2)
I O2
Q is the flowrate of oxygen in the air fed into the tank in m3.h-1 in normal where conditions (T = 0°C and P = 1 Atm), and 22.414 is the volume of one mol of gas expressed in litres,
n OO 2 = 1000 ×
QOO 2 22.414
(3)
QO where O 2 is the normal flowrate of oxygen in the air going out of the tank in m3.h-1. Moreover, I Q OI 2 = Q air × %O I2
and
Q OO2 = Q Oair × %O O2
(4) and (5)
149
Bioleaching Applications I O where %O 2 and %O 2 are the oxygen concentrations in the air flowing in and out of the I O tank in volume fractions respectively. Q air and Q air are the airflow rates in and out of the
tank respectively in m3.h-1 in normal conditions. In the same way, for nitrogen: I 2
where %N and %N in volume fractions. As
Q ON 2 = Q IN 2
O 2
Q ON 2 = Q Oair × %N O2
and
I Q IN 2 = Q air × %N I2
are the nitrogen concentrations in the air in and out respectively,
, then:
I Q Oair × %N O2 = Q air × %N I2
%N I2 = 1 − %OI2 − %COI2 and %N O2 = 1 − %OO2 − %COO2
(6) (7) and (8)
I O where %CO 2 and %CO 2 are the carbon dioxide concentrations in the air flowing in and out of the tank in volume fractions respectively. Therefore: I Q Oair = Q air ×
1 − %O I2 − %CO I2 1 − %O O2 − %CO O2
(9)
From equations (1), (2) & (3) the Oxygen Uptake Rate (OUR) can be defined as: 32 1 OUR = × × (Q OI 2 − Q OO2 ) Vo 22.414 (10) and combining (10) and (4) and (5), it comes: 32 1 I OUR = × × (Q air × %O I2 − Q Oair × %O O2 ) Vo 22.414
(11)
Then, combining (11) and (9): OUR =
I ⎞ ⎛ Q air 1 − %O I2 − %CO I2 32 × × ⎜⎜ %O I2 − × %O O2 ⎟⎟ O O Vo 22.414 ⎝ 1 − %O 2 − %CO 2 ⎠
(12) From this value of the OUR, the oxygen transfer rate in steady conditions (OTR) will be: O TR = Vo × OUR
preferably expressed in kg O2 /h In the same way, the oxygen transfer efficiency (OTE) is:
OTE =
( nOI 2 − nOO2 ) nOI 2
(13)
x100 (14)
That can also be written: OTE =
( QOI 2 − QOO2 ) QOI 2
And then:
150
x100 (15)
Bioleaching Applications
OTE = 1 −
%OO2 × (1 − %OI2 − %COI2 ) %OI2 × (1 − %OO2 − %COO2 )
(16) A comparison was established by determining oxygen uptake rate by the two methods, i.e. the pyrite oxidation and the gas mass balances. In the chemical balance method based on sulphide oxidation, it was assumed that oxygen reacts according to the stoichiometry of the following reaction: (1) 4FeS2 + 15O2 + 2H2O → 2Fe2(SO4)3 + 2H2SO4 Due to the fact that some flows were partly recycled inside the unit, it was not possible to make a reliable comparison for each bioleach tank and consequently, the comparison is established for the entire bioleach circuit. Dissolved oxygen was measured by means of an Au/Ag electrochemical Orbisphere probe. 3.
RESULTS AND DISCUSSION
3.1 Oxygen uptake rate measurement by O2 and CO2 measurements in the off-gas As shown in figure 1, the sulphide biooxidation activity was particularly variable during the first six-months of 2001. The series of stages of the unit worked well only after the first three months, as it is only after this period that the OUR of the primary stage could be clearly distinguished from the values of the other stages. The unit was, on a number of occasions, affected by mechanical and electrical problems resulting in an interruption of the normal operations of the bioleach tanks. Even after the beginning of April, operating problems like power failures for several hours affected the units’ performance. The sharp drop in activity recorded at the beginning of June is a typical consequence of that kind of problem. However, it is also noticeable that the bioleach primary tanks reacted in a relatively similar way to incidents affecting all the tanks at the same time like at the end of the study period.
Figure 1. OUR as measured by off-gas analysis vs. time in the bioleach tanks of the KCCL plant from January to June 2001 151
Bioleaching Applications
The results showing the comparison between the two methods of determining OUR for the period are shown in figure 2. Generally, it appears that the OUR provided by the gas method applied in the conditions described in this paper is overestimated when compared to the value obtained by material balance. When the operation was more stable, namely after mid-March, the two trends were rather similar, demonstrating that gas analysis as measured in those conditions provided a good method even for a relative picture of the biological activity in the tanks. It must be confessed that the system as installed at the plant suffered several limitations of which the main ones are the following: •
Gas measurements were carried out batch wise and only once a day, which means that the results could be affected by uncontrolled variations of the operating parameters like air flowrate or limestone addition. The gas measurement method gave a snapshot of the performances, whereas material balance was assumed to be more representative of the performances for a period in the range of the residence time in the tank. Moreover, it is assumed that steady state takes place and the calculation by the material balance in the slurry does not take into account the retention time in the bioleach circuit. Actually, as shown by the scattering of the data the steady state was far from being ensured on such a long period.
•
The OUR data by gas analysis as given in figure 2 are the average values for all the bioleach tanks of the unit. These values do not take into account the errors involved in the process of sampling and measurement. The sampling procedures were perfectible for the gas and for the slurry. In particular, sampling of the gas near the wall of the tank was not the appropriate place if gas was not evenly dispersed through the tank. Sampling the entire off-gas would have been the right way, but impracticable. The collection of the off-gas along a radius from the centre to the periphery of the tank would be feasible and probably satisfactory in terms of sampling accuracy.
Figure 2. Values of OUR as obtained by material balance of the sulphide oxidation and by the gas balance for the entire bioleach unit. Eventually, it is reasonable to think that the ratio between the amount of oxygen consumed and the pyrite oxidised was not strictly equal to the stoichiometrical factor as in reaction 1. In particular, the consumption of oxygen by other minor sulphides and the 152
Bioleaching Applications
biological consumption of oxygen was not taken into account. However, the nearness of the two trends for the last bioleach tank in the cascade (fig. 3), hints that the gas measurement can lead to a good indication of the performances.
Figure 3. Values of the OUR as obtained by the material balance method compared to that obtained by the gas analysis method in the last tank of the bioleach unit 3.2 OUR vs. air flow rate The air supply is the largest operating cost item of a bioleach unit and optimising the airflow rate is vital for the plant economy. The gas analysis method can be used to assess the influence of the airflow rate on biological activity. In order to determine the minimum air flow rate in nominal conditions of operation, the OUR was measured in one of the bioleach tanks by the gas analysis method for a range of air flow rates from 7,000 to 15,000 Nm3.h-1. Figure 4 shows the results of the test work.
Figure 4. OUR measured by the gas analysis method vs. airflow rate in a tank of the bioleach unit It appears that an airflow rate between 11,000 and 12,000 Nm3.h-1 is the minimum value to obtain a maximum OUR level. The maximum OUR values obtained are in the 153
Bioleaching Applications
range from 1,350 to 1,400 mg.L-1.h-1, meaning an Oxygen Transfer Rate (OTR) of about 1,750 kg O2.h-1 and an Oxygen Transfer Efficiency (OTE) of ca. 50%. Such a high OTE value was above the original predictions. Measurements of dissolved oxygen at the top of the bioleach tank showed that oxygen concentration was in the range of 1.5 to 2.0 ppm for flow rates from 10,000 to 12,000 Nm3.h-1. In this range of flow rates, the oxygen concentration is above the value of 1.5 ppm. This value of 1.5 ppm is generally agreed as being the minimum oxygen concentration acceptable for the maintenance of stable aerobic conditions required by the bacterial population in bioleach processes [3], though the system would still be efficient at as low as 0.1 ppm [4]. Furthermore, the bacterial growth was, apparently, not affected by a tip speed of the turbine of about 4.5 m.s-1. It must be mentioned that a critical operating aspect for the aeration was the difficulty to equilibrate the air pressure in all the bioleach reactors. Differences in pressure load from one reactor to another frequently resulted in blower trips. 3.3 pH Acidity level of bioleaching medium results from the balance of protons between netconsuming reactions (oxides/carbonates dissolution, arsenopyrite/pyrrhotite/iron oxidation, etc.) and reactions of sulphuric acid production and iron hydrolysis. The optimal pH range is variable from one system to another and one microorganism to another [5]. On one hand, the higher the pH, the easier the acid-producing reactions. However, a relatively high pH, between 1.8 and 2, may lead to the precipitation of iron hydroxide in excess. This would further lead to an increase in the proportion of sterile surface that would interfere with the interaction of the bacteria with the sulphides, increasing the slurry viscosity and making mixing and oxygen transfer less efficient. On the other hand, a low pH value, close to 1.0, of course is harmful to the microorganisms metabolism and can be very selective for acid-tolerant species making the biological system fragile. Other aspects can also be taken into consideration when pH is low. At low pH, ferric iron in solution from iron sulphides increases in concentration, which can reduce bacterial growth of species sensitive to this ion [6]. The lower the amount of a neutralising agent like limestone being added, the lower the amount of carbon dioxide available in situ for stimulating bacterial growth and the less the amount of gypsum being precipitated, which may have consequences on the retention time in neutralisation operation. In the case of refractory gold concentrates, it is observed that low pH reduces the risk of gold being encapsulated and being less soluble during cyanidation treatment [7]. Another aspect is related to nutrients as some authors think that too much calcium could result in less phosphate being available in solution for the bacterial metabolism needs [8]. This could be relevant for ammonium as well, because the formation of jarosite is enhanced by the addition of neutralising agent. It is true that one must be especially aware of nutrients availability in the primary stage of the bioleaching treatment in stirred tanks. The major part of the bacterial growth occurs in the primary stage of bioleaching, and the growth is reduced in the subsequent stages. This is the reason why pH can be increased from the primary to the following stages. A pH range of 1.4-1.6 is probably a good compromise between the risks mentioned above and the technical feasibility of controlling pH in huge stirred tanks. The fact is that considering the evident and unavoidable inefficiency of mixing systems at microscopic
154
Bioleaching Applications
scale, the range of local pH in a tank is actually very large in any situation, from less to 1 to more than 7 probably. 3.4 Temperature The bioleach unit was originally designed to work at 42°C. At laboratory scale, it was shown that the maximum temperature was 46°C. Above this value, biological growth and bioleaching activity were significantly reduced. In practice, it happened that the cooling system was damaged in one of the bioleach tanks and the temperature was left to rise up to almost 50°C in this tank. Actually, the slurry temperature in the bioleach tank could only be kept in the range from 46 to 50°C. However, no significant reduction of the bioleach activity could be observed and the performances were quite similar to the performances in nominal design conditions. It therefore seems that the bacterial system is flexible to the point that new properties can appear when the system has to adapt to new operating conditions. It would confirm the great bacterial diversity in bioleach tanks, especially resulting from the microorganisms naturally occurring with the substrate itself. Unfortunately, no investigation was carried out to identify the change in bacterial population, while temperature had increased. 3.5 Modelling and prediction Modelling the biochemical system in bioleach units is an extremely difficult task and the gain of a possibly efficient mathematical model is unpredictable. A modelling approach with a detailed representation of the medium content is not required. On the other hand, a pragmatical model based on experimental data just for simulation as used for the engineering design of a plant and for the cost estimate is vital and sufficient. Moreover, the ability to understand the phenomena occurring at the reaction interfaces could be quite essential. Indeed, it is, for instance, known that nutrients like ammonium do not only play a role for bacterial growth. They also form compounds with iron covering the sulphides, which compounds do not inhibit the oxidising reactions contrary to what is too often said when no other argument can be used, but these compounds are probably essential components of the catalytic process of biochemical transformations. During the operation at Kasese, a correlation between bioleaching efficiency and ammonium concentration was observed, even when the available amount in solution was presumably largely enough for the bacterial growth (results not shown in this paper). The compromise between the necessary concentration of ammonium for a good bacterial growth, the amount to be precipitated to play the optimum catalytic role in the biochemical process and a reasonable rate of consumption remains to be established at the real scale. 4.
CONCLUSION Operating a bioleach unit at industrial scale reveals the limits of the conclusions made from test work done on small scale. It must be admitted that generally a bioleach plant has better capacity to tolerate changes in operating conditions than expected. In particular, when the biomass is really established it is observed that changes in the normal operating conditions like absence of feed, and other problems related to cooling, mixing, aeration, etc. do not necessarily lead to major production problems. Actually, bioleach ecosystems have a great flexibility and own unknown resources, particularly in the composition of their bacterial population. Already, bioreactors are appropriate confined vessels where new micro-organisms may be isolated like Ferroplasma acidiphilum, an acidophilic, 155
Bioleaching Applications
autotrophic, cell-wall-lacking, mesophilic new micro-organism discovered by serial dilution of the aqueous phase of a bioreactor of a pilot plant treating a gold-bearing arsenopyrite/pyrite concentrate [9]. The diversity of the biological content and the positive selection that bioreactors operate, make that a continuous bioleach unit is a very robust industrial system. However, practise at this scale also shows the limitation of the available tools to monitor the activity. Trouble-shooting is difficult before the consequences of the problem become really serious and result in significant loss of production. The gas analysis system as tested in the frame of the KCCL project was not designed for working in an industrial context. However, it was helpful enough to the operators for the monitoring of the daily bacterial activity. A fully automatic system connected to the general process control system would be easy to design and could inform the operators on critical parameters in real-time. The automatic system would provide the sulphide oxidising performances in steady state operating conditions. ACKNOWLEDGEMENTS The authors are grateful to BRGM’s Direction of Research for the authorisation to publish this paper under the reference No 02493. Mick Rogers of Newmont Mining is especially thanked for his objective help to the installation and the use of the gas analysis system on site. REFERENCES 1. Brierley, J. A. and Brierley, C.L., Biohydrometallurgy and the environment toward the mining of the 21st century, Part A, Elsevier, Amsterdam, The Netherlands, (1999), 81. 2. Briggs, A. and Millard, M., IBS’97 and Biomine’97, Conference proceedings, Australian Mineral Foundation, (1997), M.2.4.1. 3. Dew, D. and Miller, G., IBS’97 and Biomine’97, Conference proceedings, Australian Mineral Foundation, (1997), M7.1.1. 4. Ritchie, I. and Barter, J., IBS’97 and Biomine’97, Conference proceedings, Australian Mineral Foundation, (1997), M14.4.4. 5. Rawlings, D.E., Biomining: Theory, Microbes and Industrial Processes. Springer, Landes Biosciences, Georgetown, TX, USA, (1997), 229. 6. Collinet, M.,N. and Morin, D., Antonie van Leeuwenhoek, 57, (1990), 237. 7. Spencer, P.A., Int. J. Miner. Process, 62, (2001), 217. 8. Barrett, J., Hughes, M.N., Karavaiko, G.I., and Spencer, P.A. Metal extraction by bacterial oxidation of minerals, Ellis Horwood series in inorganic chemistry, New York, (1993). 9. Golyshina, O. V., Pivovarova, T.A., Karavaiko, G.I., Kondrateva, T.F., Moore, E.R.B., Abraham, W.R., Lünsdorf, H., Timmis, K.N., Yakimov, M.M. and Golyshin, P.N., International Journal of Systematic and Evolutionary Microbiology, 50, (2000) 997.
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Bioleaching of natural zeolite – the processes of iron removal and chamfer of clinoptilolite grains I. Štyriakováa, D. Koloušekb, I. Štyriakc, C. Lengauerd, E. Tillmannsd a
Department of Biotechnology, Institute of Geotechnics of the Slovak Academy of Sciences, Watsonova 45, 05343 Košice, Slovakia, E-mail:
[email protected] b Department of Solid State Chemistry, Institute of Chemical Technology, Technicka 5, 166 28 Prague 6, Czech Republic c Department of Microbiology, Institute of Animal Physiology of the Slovak Academy of Sciences, Šoltésovej 4-6, 040 01 Košice, Slovakia d Institut fuer Mineralogie und Kristallographie, Geozentrum, Althanstr. 14, A-1090 Wien, Austria Abstract Natural zeolite, including clinoptilolite, often contains iron and manganese which decrease the whiteness of this sharp angular material. The biological treatment of zeolite enables its use as the substitute for tripolyphosphates in wash powders which have to comply with strict requirements as far as whiteness is concerned and rounded off grain content. Insoluble Fe3+ and Mn4+ in the zeolite could be reduced to soluble Fe2+ and Mn2+ by silicate bacteria of Bacillus spp. These metals were efficiently removed from zeolite as documented by Fe2O3 decrease (from 1.37% to 1.08%) and MnO decrease (from 0.022% to 0.005%) after bioleaching. The whiteness of zeolite was increased by 8%. The leaching effect, observed by scanning electron microscopy, caused also the chamfer of the edges of sharp angular grains. Despite the enrichment by fine-grained fraction, the decrease of the surface area of clinoptilolite grains from the value 24.94 m2/g to value 22.53 m2/g was observed. This fact confirms the activity of bacteria of Bacillus genus in the edge corrosion of mineral grains. Removal of iron and manganese as well as of sharp edges together with the whiteness increase should give a product which is fit for industrial applications. Keywords: zeolite, bioleaching, Bacillus sp., silicate bacteria 1.
INTRODUCTION Natural zeolite and its synthetic counterparts are used as filters, sorbents and ion exchangers with the characteristics of superior selective adsorption properties. Also, many researchers have found applications in air and wastewater pollution control, gas purification, petroleum refining and oxygen concentration. Because a dependable supply of zeolite is important, the synthesis of zeolite was developed by scientists.sBefore a large amount of natural zeolite was found in rock formations from the Cenozoic age, volcanic and sedimentary rocks of volcanic origin, only synthetic zeolite was used in the areas of pollution control, petroleum refining and gas purification processes due to the convenient 157
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supply. If the physical and chemical properties of natural zeolite are defined and improved by special modification processes, adequate supplies for commercial users and a reduction in costs could be achieved. Moreover, it would be viewed as a very competitive material in many fields [1]. In the West Carpathians, deposits of zeolites including clinoptilolite, mordenite, and analcime occur exclusively in Miocene silicate volcanoclastic rocks. Two areas of zeolite occurrences, each having a different genesis and economic significance, can be distinquished in the territory of Slovakia: (a) the East Slovakian basin with regionally widespread zeolitization associated with the Lower Badian sequence of rhyodacite, the so – called Hrabovec tuffites, (b) the southwestern margin of the Kremnické vrchy Mountains having a distinct zone of about 2km2 of Upper Sarmatian – Pannonian zeolitic tuffs [2]. Natural zeolites could be far more utilized, if the modification and purification techniques for these materials made faster progress. With natural zeolites, in addition to acidity modification, the H3PO4 treatment can assist in elimination of impurities, such as carbonates or Fe. Iron is one of the most significant impurities in natural zeolites [3, 4] and it would be of great practical interest to reduce its content without crystal lattice destruction. Berthelin et al. [5] have described the important role of bacteria in iron reduction when an enzymatic mechanism similar to dissimilative nitrate reduction should be involved in this reaction. Fe3+ is mobile only at very low pH values (pH<3). Reduction enables the formation of Fe2+, which is mobile in the normal range of soil pH. Consequently, if microorganisms and plants are able to reduce Fe3+, they can have advantage in competition for available iron. In the case of silicates, an increase in Fe solubility generally occurs with acid and complex secretions, and feldspars and micas can be destroyed [6]. Bacillus spp. play an important role in silicate biodegradation during the process of rock disintegration [7, 8]. The results of such activity involve both geochemical and structural changes in silicate minerals and rocks. Tešič and Todorovič [9] have proposed that so-called "silicate bacteria" belong to the Bacillus circulans group. The mechanism of microbial destruction of silicates and aluminosilicates by these bacteria is not understood yet. However, it is known, for example, that their activity leads to a decrease in Si content of bauxites of lower quality [10], and to the extraction of Al, Ti, U, Au and other elements from silicates and aluminosilicates [11]. This paper describes the treatment of natural clinoptilolite by bioleaching. Clinoptilolite, the most abundant natural zeolite, is a member of the heulandite goup of the natural zeolites. 2.
MATERIALS AND METHODS
2.1 Zeolite sample Zeolite sample (NI) from Nižný Hrabovec deposit is composed of clinoptilolite (5168%), quartz + cristobalite (9-20%), feldspar (4-13%) and mica (13-22%). The ironbearing minerals decrease the quality of this raw material. Its chemical characteristics are shown in Table 1.
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2.2 Bacteria and media Two bacterial strains (Bacillus cereus and Bacillus pumilus) were isolated from a kaolin quarry in Horná Prievrana by a colony reisolation on Nutrient agar No.2 (Imuna, Šarišské Michaľany) plates to obtain pure strain cultures. They were identified by means of the BBL CRYSTAL ID panel (Becton-Dickinson, USA). This panel contains 29 enzymatic and biochemical substrates and a fluorescence control on tips of plastic prongs. The resulting pattern of the 29 reactions is converted into a ten-digit profile number that is used as the basis for identification of a wide variety of microorganisms. For the species identification, the strains were cultivated on Columbia agar plates according to recommendation of the panel producer. For experiment, the bacterial strains were grown in Nutrient broth No.2 (Imuna) at 37 oC for 18 hours. Bacterial cells were subsequently centrifuged at 4000 rpm for 15 min, subsequently washed twice with saline solution (0.9% NaCl) and added in a concentration of 1010 cells per ml to modified Bromfield liquid medium [12]. Bioleaching of the samples was carried out in 3000 ml Erlenmeyer flasks containing 2000 ml of modified Bromfield medium (NaH2PO4 – 0.5g/l, MgSO4.7H2O 0.5g/l, (NH4)2 SO4 – 1.0g/l, NaCl – 0.2g/l, glucose – 10g/l) inoculated with a mixture of both Bacillus cereus and Bacillus pumilus strains. The flasks were incubated statically for 85 days at 28oC. The abiotic controls were cultivated under the same conditions. After incubation, the culture solutions were separated from the biomass by means of membrane filtration. The presence of vegetative bacterial cells in Erlenmeyer flasks and their morphology were regularly examined by light microscopy after Gram staining. 2.3 Chemical analyses Quantitative changes of samples (solid and liquid phases), investigated from the view of element composition stability, were evaluated by standard analytical method – atomic absorption spectrometry on a VARIAN spectrometer AA - 30 apparatus (Varian, Australia) after dissolution of the samples by standard procedure. 2.4 Grain size analysis The particle size distribution of zeolite sample was measured by the laser radiation scattering on a Laser - Particle - Sizer Analysette 22 (Fritsch, Idar – Oberstein, Germany). 2.5 X-ray diffraction analysis The qualitative characterization of the bacterially leached samples was investigated on a Philips X'PERT powder diffractometer with CuKa radiation (40kV, 40mA), equipped with an automatic divergence slit, sample spinner, and a graphite secondary monochromator. For the data collections the range 2-72 deg2theta, a step width of 0.02 deg, and a counting time of 3sec/step were selected. 2.6 Scanning electron microscopy The morphological changes in the surfaces of individual minerals were investigated by SEM (scanning electron microscopy) and the changes of chemical composition by energy-dispersion microanalysis (EDS). All mineral samples were coated with carbon and subsequently examined in a Tesla BS 340 scanning electron microscope. 3.
RESULTS AND DISCUSSION The bioleaching system studied in this work involves Bacillus bacteria and their metabolic products, zeolite particles and the leaching medium. First, we monitored the 159
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elements extraction by bacterial leaching from NI sample using atomic absorption spectrophotometry (Fig. 1). It is demonstrated that bacterial growth and elements extraction from zeolite particles are coupled through biochemical interactions, which involve the different components in this system. The leaching rate of mineral particles was dependent on the fermentation of the organic compound (glucose) during bioleaching discontinuous processes. When the bacteria of Bacillus spp. were inoculated into zeolite, it was observed that gases were produced after a short adaptation phase due to the fermentation of organic compounds during elements extraction. The element extraction in zeolite began gradually increased with the generation of these fermentative gases and ceased to progress when gas generation or colour and pH change of medium was no longer observed. However, very low elements extraction was observed in the flasks without bacterial inoculum (data not shown). This fact suggests that elements extraction is bacterially mediated.
Extraction of elements (mM)
6
K
5
Fe Si
4
Al
3 2 1 0 0
15
30
45
60
75
90
Time (days)
Figure 1. Kinetics of elements extraction as followed by AAS In the presence of the zeolite during bacterial growth, the pH of the Bromfield medium was decreased from 6.5 to about 4.0 within 5 days, because organic acids were accumulated from the fermentation reaction of glucose. That is why this pH decrease was continually neutralized to pH 6.5 during bacterial leaching. Most of the bacteria were adsorbed on the mineral surfaces during discontinuous bioleaching as shown by light microscopy. Therefore, the bulk solution was exchanged with fresh glucose containing Bromfield medium nine times during 85 days without significant loss of active bacteria. The initial Eh was 160 mV and then was decreased to -490 mV during fermentative processes, indicating an anaerobic environment where the reduction reaction elements such as Fe3+ and Mn4+ performed well. The refinement of zeolite was carried out through these processes. Fe (III) can generally serve as an electron acceptor in microbial metabolism around 182 mV [13]. The microbial reduction was more active in ferric hydroxide than in goethite. The other form – hematite, was difficult to reduce by microorganisms. It appears that the lower the degree of crystallization, the higher is the probability of reduction [14]. Chemical composition of zeolite before and after bioleaching is shown in Table 1. The insoluble Fe3+ and Mn4+ was reduced to soluble Fe2+ and Mn2+ by silicate bacteria of 160
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Bacillus spp. The content of Fe2+ in solid samples after bioleaching was increased because it was probably adsorbed by zeolite and bacterial cells. The removal of total iron content was not to exceed 21%. The whiteness of zeolite was increased in 8% after bioleaching and grey-green colour before bioleaching was changed to white-green after bioleaching. Table 1. Effect of bioleaching of zeolite samples on elements removal Chemical composition w.t.(%)
Sample before bioleaching
Sample after bioleaching
SiO2
67.02
64.44
Al2O3
12.25
11.24
K2O
3.12
2.83
Fe2O3
1.19
0.74
FeO
0.16
0.31
MnO
0.022
0.005
TiO2
0.17
0.15
CaO
3.23
1.64
MgO
0.65
1.30
Na2O
0.65
1.29
Li2O
0.004
0.004
SZ*
7.73
12.05
SS*
3.62
3.67
S* 99.82 SZ* loss by ignition (900°C), SS* loss by drying, S* total amount (%)
99.67
Microbial production of organics by fermentation, or reductive dissolution of Fe – Mn mineral phases can greatly accelerate weathering rates of aluminosilicate minerals [15, 16, 17]. The mineral composition of zeolite sample is shown on X-ray diffraction pattern (Fig. 2). This pattern indicates that clinoptilolite, cristobalite, feldspar, celadonite and quartz are major constituents of the NI sample. The portion of quartz and feldspar was lowered by bioleaching as documented by chemical and mineralogical X-ray analyses. Moreover, the changes in Ca, Na, K, and Mg contents in clinoptilolite sample were visible on X-ray pattern because they caused the intensity decrease of individual peaks. This fact suggests a transformation of structural plains of clinoptilolite (hkl –422, 441). Powder X-ray diffraction pattern reveals the clinoptilolite as a predominant mineral, however, the monoclinic habit of this mineral is not easily recognized in the scanning electron microscope. The raw material contained sharp angular grains after pulverizing and before bioleaching (Fig. 3). The plates or blades of clinoptilolite are visible after bioleaching and are generally less than 20µm in length (Fig. 4). The leaching effect, observed by scanning electron microscopy, caused also the chamfer of the edges of sharp angular grains and coating of fine-grained particles on grain surfaces (Fig. 4). The amount of finest-grained fraction from 0.9 to 5.0 µm and the distribution of fraction with particle size from 51µm to 103µm was in NI sample increased, and on the other hand, the distribution of fraction with particle size from 6.0 to 103.0 µm was decreased (Fig. 5). Despite the enrichment of the sample by fine-grained fraction, there was observed the decrease of the surface area of clinoptilolite grains from the value 24.94 m2/g to value 22.53 m2/g. This fact confirms the activity of bacteria of Bacillus genus in the edge corrosion of mineral grains. We are sure that so-called "silicate bacteria" include not only 161
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Bacillus circulans group as proposed by Tešič and Todorovič [9] but also many other Bacillus spp. described in last years. Zeolitic tuffite from Nižný Hrabovec, earlier described as rhyodacite tuffite, was found to contain approximately 40-56% clinoptilolite. The zeolite was formed from volcanic ash as a product of its diagenetic alteration. The crystals of clinoptilolite are impregnated by amorphic phases and cristobalite [13]. Probably, these amorphic phases of zeolite matrix were particularly destructed by bioleaching because this devitrification was observed also on X-ray pattern as a partial increase of intensities of some diffraction lines of clinoptilolite (Fig 2). counts 3000 zeolite Nitrosorb
2500
F C Q
2000
1500
S
1000
nat.
500
BL
0 5
10
15
20
25
30
35
40 °2Theta
Figure 2. X-ray diffraction pattern of the zeolite sample (S-celadonite, C-cristobalite, Q-quartz, F-feldspars, unsigned peaks - clinoptilolite)
Figure 3. Zeolite particle before bioleaching
162
Figure 4. Zeolite particle after bioleaching
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Distribution density (logarith.)
45 Zeolite sample before bioleaching Zeolite sample after bioleaching
40 35 30 25 20 15 10 5 0 0
10
20
30
40
50
60
70
80
90 100 110
Particle size (µm)
Figure 5. The granulometrical analyse of zeolite samples Depending on physical and chemical properties, the zeolite rock from Nižný Hrabovec (Slovakia) should be useful also in many other industrial and agricultural fields (waste water treatment, soil conditioner, absorbent for water removal from refrigerants, for SO2, CO2, and NH3 removal processes from waste gases, etc.) as described by Kozáč et al. [18]. Removal of iron and manganese as well as of sharp edges together with the whiteness increase should give a product which is also fit for industrial applications as the substituent for tripolyphosphates in wash powders which nowadays contaminate the environment. Complex phosphates were identified as major contributors to the de-oxygenation of inland waters. Phosphonates cause the same problems, albeit requiring a greater quantity to do the same damage. Since few companies in the UK are equipped to remove phosphates, they persist in waterways. A phosphate ban was introduced in some Scandinavian countries, and many UK companies now use builder systems based on zeolite, which is similar to pumice stone, with or without polycarboxylic acid, sodium citrate, or carbonates. Zeolite is inert, but it has a positive environmental impact [19]. ACKNOWLEDGMENTS The authors are grateful to the Slovak Grant Agency for Science (Grant No. 2/2107/22 and Austrian Science and Research Liason Office. REFERENCES 1. H.K. Lee, M.J. Shim, J.S. Lee and S.W. Kim, Materials Chemistry and Physics, 44 (1996) 79. 2. E. Šamajová in: D. Kallo and H.S. Sherry (eds.), Occurrence, properties and utilization of natural zeolites. Budapest, 1988.
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3. R.R. Malherbe, C.D. Aguila, E.R. Ruiz, J.F. Lliteras, L.L.Colorado and M. V. Velez, Zeolites, 10, (1990), 685. 4. C. Pozas, C.D. Aguila, E.R. Ruiz and R.R. Malherbe. J.Solid State Chem., 93 (1991) 215. 5. J. Berthelin, M. Bonne, G. Belgy and F.X. Wedrago, Geomicrobiol. J., 4 (1985) 175. 6. M. Robert and C. Chenu in: Stotzky G. and Bollag J.M. (eds): Soil Biochemistry, Vol. 7, New York, 1992. 7. G.I. Karavaiko, Z.A. Frutsko, E.O. Melnikova, Z.A. Avakyan, Y.I. Ostroushko, Role of microorganisms in spodumene degradation, Microbiology-USSR, 49 (1980) 547549. 8. G.I. Karavaiko, N.P. Belkanova, V.A. Eroshchev-Shak, Y.A. Avakyan, Role of microorganisms and some physico-chemical factors of the medium in quartz destruction, Microbiology-USSR, 53 (1984) 795-800. 9. Z.P. Tešič, M.S. Todorovič, Prilog ispitivani ju "Silicatnih Bakterija", Zemljiste i biljka (Belgrad) 1 (1952) 3-18 (in Serbian). 10. V.I. Groudeva, S.N. Groudev, Bauxite dressing by means of Bacillus circulans, In: Travaux ICSOBA Congress, Zagreb (1983), pp. 257-263. 11. S.N. Groudev, Workshop on the aluminosilicate mineral biodegradation: a synopsys, 5th European congress on biotechnology, July 8-13, 1990, Copenhagen, Denmark. 12. S.M. Bromfield, J. Gen. Microbiol., 11 (1954) 1. 13. W.C. Anderson, Water Environment Federation, Alexandria, 1995. 14. E.J.P. Phillips, D.R. Lovley and E.E. Roden, Appl. Environ. Microbiol., 59 (1993) 2727. 15. S.A. Welch and W.J. Ullman, Geochimica et Cosmochimica Acta, 60 (1996) 2939. 16. H.L. Ehrlich, Chemical Geology, 132 (1996) 1. 17. P.C. Bennett and F.K. Hiebert, in: Kharaka, Y.K. and Maest, A.S. (Eds) Water – Rock interaction-7, Vol. 1, Rotterdam, 1996. 18. J. Kozáč and D. Očenáš, Mineralia Slovaca, 14 (1982), 549. 19. information from internet
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Bioleaching of pyrite by defined mixed populations of moderately thermophilic acidophiles in pH-controlled bioreactors Naoko Okibe and D. Barrie Johnson School of Biological Sciences, University of Wales, Bangor, LL57 2UW, U.K. Abstract Pure and defined mixed cultures of moderately thermophilic acidophiles were compared for their abilities to accelerate the oxidative dissolution of pyrite oxidation in pH-controlled bioreactors run at 45°C. Three species of acidophiles used were; a thermophilic Leptospirillum isolate (MT6), Acidimicrobium ferrooxidans (strain ICP) and Acidithiobacillus caldus (strain KU). Microbial populations were analyzed using a combination of cultivation techniques (plate counts on selective solid media) and a molecular approach (fluorescent in situ hybridisation: FISH). Pyrite oxidation in mixed cultures of the two iron-oxidizers was greater than that by Am. ferrooxidans alone, but slightly less than by pure cultures of Leptospirillum MT6, suggesting no synergistic interaction between these bacteria. In contrast, mixed cultures of Am. ferrooxidans and At. caldus were the most effective system tested. Results from FISH and plate counts showed that, in contrast to earlier reports, numbers of iron-oxidizing bacteria were frequently greater than those of At. caldus in mixed cultures, and that the most efficient cultures contained more bacteria and dissolved organic carbon than relatively inefficient pure cultures and consortia. These results indicate that autotrophic At. caldus stimulate mineral dissolution by the "heterotrophically-inclined" Am. ferrooxidans by providing the latter with organic carbon as well as by oxidizing sulfur and polythionates. Keywords: Leptospirillum; mixed cultures; pyrite; synergism; thermophiles 1.
INTRODUCTION Biological oxidation of sulfidic minerals may be mediated by a variety of pure and mixed cultures of acidophilic microorganisms. Frequently, such systems have been categorized on the basis of their optimum temperatures. "Moderately thermophilic" acidophiles are those that grow optimally between ~45-55°C, and include iron-oxidizing bacteria (e.g. Sulfobacillus spp., Acidimicrobium ferrooxidans) sulfur–oxidizing bacteria (e.g. Acidithiobacillus caldus and Sulfobacillus spp.), and heterotrophic bacteria (Alicyclobacillus-like) and archaea (Thermoplasma and Picrophilus spp.) [1]. Some Ferroplasma-like isolates (iron-oxidizing archaea) can also grow at 45°C. In contrast to mesophilic species, the majority of characterized thermophilic iron-oxidizers listed above are either mixotrophic or heterotrophic. A notable exception is Leptospirillum thermoferrooxidans, which was reported to grow at up to 55-60°C [2], though the original (and sole) isolate has subsequently been lost. A thermotolerant Leptospirillum isolate 165
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(MT6) was isolated from a 45°C pilot-plant pyrite bioleaching operation by Okibe et al. [3]; this iron-oxidizer had a temperature optimum of 43°C and grew at up to 50°C. Phylogenetic analysis showed that it was most closely related to the newly designated species, L. ferriphilum [4]. At. caldus, Ferroplasma- and Sulfobacillus-like prokaryotes were also isolated from the stirred tanks, and relative numbers of microorganisms fluctuated as mineral leaching progressed. Given the increasing importance of biological mineral processing, it is somewhat surprising that there are still relatively few detailed accounts of the microflora present in commercial mineral leaching operations [5]. However, it is recognized that consortia rather than pure cultures of acidophiles are involved in mineral biooxidation, and that interactions between microorganisms are of fundamental importance in determining the efficiencies of the process [6, 7]. In this paper, we describe pyrite oxidation by defined pure and mixed cultures of three species of moderately thermophilic acidophiles. Changes in microbial populations as mineral oxidation progressed were monitored using a combination of cultivation (plating) and molecular (FISH) techniques. 2.
MATERIALS AND METHODS
2.1 Bacteria and bioleaching protocols Three species of moderately thermophilic acidophiles were used in the present study: (i) Leptospirillum MT6 [3]; Acidimicrobium ferrooxidans (strain ICP) which was kindly provided by Dr. Paul Norris (Warwick University, U.K.) and Acidithiobacillus caldus (strain KU) which is maintained in the Acidophile Culture Collection at Bangor University. Bacteria were routinely subcultured in 2% (w/v) pyrite (Leptospirillum MT6), 2% pyrite + 0.02% (w/v) yeast extract (Am. ferrooxidans) and 5 mM potassium tetrathionate (At. caldus) liquid media (pH 2.0). 2.2 Shake flask experiments Bioleaching of pyrite (obtained from the Cae Coch mine, north Wales [8]) by various combinations of the three moderately thermophilic acidophiles was compared in shake flasks, containing 100 ml of 2% (w/v) of finely ground ore. Cultures were inoculated (2%, v/v) and incubated, shaken (130 rpm) at 45ºC. Samples were withdrawn at regular intervals to determine concentrations of soluble iron (total and ferrous), sulfate, pH, dissolved organic carbon (DOC) and to enumerate bacteria. 2.3 Pyrite bioleaching in pH-controlled bioreactors Bacteria, grown in batch culture as described above, were used to inoculate 2 L bioreactors (P350; Electrolab, U.K.) fitted with temperature, pH and aeration control. The working volume of the bioreactor was 1.5 L, and 2 x 109 cells of each acidophile (enumerated in a Thoma counting chamber) were added at the start of each bioleaching experiment. Again, various combinations of the three acidophiles were used, and on each occasion two bioreactors were run in parallel to compare either a pure and a defined mixed culture, or else consortia containing different combinations of moderate thermophiles. Cultures were maintained at 45ºC throughout, and aerated at 0.2 L/minute. The initial bioreactor pH was fixed at 1.5, though in the later phases of each experiment this was lowered to 1.2, and finally 1.0 (by addition of sulfuric acid). Samples were withdrawn at regular intervals to determine ferrous iron, sulfate, DOC and microbial populations. The amounts of acid or alkali required to maintain cultures at the pre-determined pH were also recorded. 166
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2.4 Analysis of microbial populations using solid media Mineral leaching samples were vortexed thoroughly to dislodge loosely adhering bacteria from mineral surfaces, a dilution series prepared and inoculated onto solid media. The latter were ferrous sulfate (for Leptospirillum MT6) and ferrous sulfate/tetrathionate (for Am. ferrooxidans and At. caldus) overlay media; these had previously been shown to be highly efficient in promoting colony growth of these bacteria (data not shown). Full details of these solid media are given elsewhere [9]. Plates were incubated at 45ºC for up to 10 days. Colonies were readily distinguished from each other (Leptospirillum MT6 as small Fe3+-encrusted colonies; Am. ferrooxidans as ferric iron-stained “fried egg”-like colonies, and At. caldus as cream/white-colored colonies. 2.5 Analysis of microbial populations by fluorescent in situ hybridization (FISH) The protocols used for FISH analysis are described elsewhere [10]. To analyze relative numbers of microbes using FISH, fixed cells were hybridized with a Cy3-labelled probe that targeted a specific acidophile, and simultaneously with a fluorescein-labelled eubacterial probe (EUB338) that targeted all eubacterial cells. Numbers of a specific acidophile were compared to total numbers of eubacterial cells targeted by the EUB338 probe, and also to total numbers of microorganisms stained with 4',6΄-diamidino-2phenylindole (DAPI) in the same field of view, to work out the relative abundance of a particular acidophile. Probes used to target Leptospirillum MT6, Am. ferrooxidans ICP, At. caldus KU were LF655 [11], ACM995 (P.R. Norris, unpublished) and THC642 [12], respectively. 2.6 Miscellaneous analyses Ferrous iron was determined colorimetrically using the ferrozine assay [13] and total iron by atomic absorption spectrometry. Sulfate was determined turbidometrically as barium sulphate (Hydrocheck; Cambridge, U.K.). DOC was determined using a Protoc DOC analyzer (Pollution & Process Monitoring Ltd., U.K.). 3.
RESULTS
3.1 Shake flask experiments Both Leptospirillum MT6 and Am. ferrooxidans oxidized pyrite, though the latter was not very effective in pure cultures (data not shown); iron solubilization in pure cultures of At.caldus was similar to that in abiotic controls (Fig. 1a). In contrast, mixed cultures of Am. ferrooxidans and At. caldus oxidized pyrite at similar rates as mixed cultures that also contained Leptospirillum MT6 (Fig. 1b). Pyrite leaching by mixed cultures of Leptospirillum MT6 and At. caldus was very similar to those by the iron-oxidizer alone for the first 35 days, though concentrations of soluble iron in the mixed culture were greater after this time (Fig. 1a). This contrasts with culture pH, which was much lower in the mixed cultures than in pure cultures of either acidophile throughout. DOC concentrations were greater in mixed (104 +/-13 mg/L) than in pure cultures (17+/-1 for At. caldus; 88+/6 mg/L for Leptospirillum MT6).
167
Bioleaching Applications (a)
8000
1.6
6000
1.4
4000
1.2
2000
1
0
0.8 10
20
30
40 Time (days)
50
60
70
80
10000
8000 Soluble Fe (mg/l)
1.8
Soluble Fe (mg/l)
10000
0
(b)
2
pH
12000
6000
4000
2000
0 0
5
10
15
20
25
30
35
Time (day)
Figure 1. Oxidation of pyrite in shake flask cultures. (a) pure and mixed cultures of Leptospirillum MT6 and At. caldus (key: n, O Leptospirillum MT6; ▲, ∆ At. caldus; ,
Leptospirillum MT6/At. caldus). (b) pure and mixed cultures of Am. ferrooxidans (key: n, Am. ferrooxidans; ▲, Am. ferrooxidans/Leptospirillum MT6; , Am. ferrooxidans/At.caldus; X, Am. ferrooxidans/At. caldus/Leptospirillum MT6. Data from leaching experiments carried out in controlled pH bioreactors are shown in Fig. 2. Am. ferrooxidans was highly ineffective in leaching pyrite in pure culture, while the mixed culture of this iron oxidizer and At. caldus displayed the most rapid rate of pyrite dissolution of all those tested during the early phase (days 6-20) of the experiment, though this declined dramatically after day 20 (Fig. 2a). The mixed culture containing both iron- oxidizers (Leptospirillum MT6 and Am. ferrooxidans) also oxidized pyrite at a relatively slow rate, and the addition of At. caldus to this consortium resulted in a marked increase in mineral dissolution (Fig. 2b). The consortium of all three moderate thermophiles was also superior in leaching pyrite to the mixed culture containing only Leptospirillum MT6 and At. caldus (Fig. 2c). Acid production in mixed cultures of pyrite-oxidizing moderate thermophiles was determined from the amounts of alkali that were required to maintain culture pH at the pre-set levels. No alkali was consumed in the pure culture of Am. ferrooxidans though, in contrast, 400 mmoles of NaOH was added to the mixed Am. ferrooxidans/At. caldus culture during the first (pH 1.5) phase of the experiment, though no more was consumed when the culture pH was subsequently lowered (data not shown). With other mixed cultures, greater amounts of alkali were consumed in the consortia containing all three moderate thermophiles than in mixed cultures containing only two. Interestingly, no alkali was consumed in the mixed culture of Leptospirillum MT6 and At. caldus. Microbial populations in two of the bioreactor cultures, as determined by plate counts and FISH analysis, together with DOC data from these cultures, is shown in Fig. 3. Plate counts indicated that numbers of both A. ferrooxidans and At. caldus (in the culture containing only these two bacteria) increased during days 1-17, and subsequently declined. There was generally a good correlation between plate counts and relative numbers determined by FISH with this culture; e.g. at day 17 both techniques showed that Am. ferrooxidans was more numerous, whereas at day 28 the two acidophiles were present in similar numbers. Concentrations of DOC increased steadily throughout incubation, and there was a marked increase when culture pH was lowered to 1.0. In contrast, colonies of At. caldus were only observed on two occasions from the Leptospirillum MT6/Am. ferrooxidans/At. caldus consortium, even though this moderate thermophile was detected (at 23-64% of the total population) by FISH analysis. The iron-oxidizers were numerically dominant in this culture, except when the culture pH was lowered to 1.0, when At. caldus 168
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comprised 64% of the total microbial population (Fig. 3b). Again, there was a gradual, though more spasmodic, increase in DOC concentrations throughout incubation, and a large increase when culture pH was adjusted to 1.0. 11000
(a)
10000
Soluble Fe/Sulfate-S (mg/l)
9000 8000
(i)
7000
(ii)
(iii)
(iv)
6000 5000 4000 3000 2000 1000 0
11000 10000
(b)
Soluble Fe/Sulfate-S (mg/l)
9000
(i)
(ii)
(iii)
(iv)
8000 7000 6000 5000 4000 3000 2000 1000 0 11000
(c)
10000
Soluble Fe/Sulfate-S (mg/l)
9000 8000 7000 6000 5000 4000 3000
(i)
2000
(ii)
(iii)
(iv)
1000 0 0
5
10
15
20
25
30
35
40
45
50
Time (days)
Figure 2. Oxidation of pyrite in bioreactor cultures (n,O iron; S, U sulfate-S) (a) Am. ferrooxidans (O, U); Am. ferrooxidans/At. caldus (n, S). (b) Am. ferrooxidans/Leptospirillum MT6 (O, U); Am. ferrooxidans/Leptospirillum MT6/At. caldus (n, S). (c) Leptospirillum MT6/ At. caldus (O, U); Leptospirillum MT6/At. caldus/Am. ferrooxidans (n, S). (i) pH 1.5; (ii) pH controlled removed; (iii) pH 1.2; (iv) pH 1.0. 169
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4.
DISCUSSION AND CONCLUSIONS Whilst it is acknowledged that mixed cultures of acidophiles are involved in mineral biooxidation processes, there are relatively few reports detailing microbial populations in commercial operations. The "primary" microorganisms involved are, in many cases, ironoxidizers such as Leptospirillum spp., since some sulfidic minerals (such as pyrite) are acid-stable, but are oxidized by the ferric iron that these prokaryotes produce. Sulfuroxidizing bacteria, such as At. caldus, whilst having no direct role in the process, exploit the elemental sulfur and polythionates formed as by-products [14]; since the free energy available from oxidizing sulfur and/or polythionates far exceeds that available from ferrous iron oxidation, sulfur-oxidizers may be present in far larger numbers than ironoxidizers in mineral leachate liquors [15]. Since elemental sulfur may form on the surfaces of oxidizing sulfides and inhibit mineral oxidation, At. caldus-like bacteria have been considered to have a beneficial role in biomining operations (e.g. [7]). Another way in which acidophilic microorganisms interact is in the production (by autotrophic species) and consumption (by heterotrophic and mixotrophic species) of soluble organic compounds [6]. This is particularly pertinent at moderately thermophilic temperatures (4050°C) where the many of the known iron-oxidizers appear to have limited capacities for fixing CO2, which restricts their capacities for mineral oxidation when grown in pure cultures without extraneous organic carbon. For example, Sb. thermosulfidooxidans was found to oxidize pyrite effectively in mixed culture with At. caldus or in pure cultures amended with yeast extract, but not in organic C-free medium [7]. Of the two iron-oxidizing moderate thermophiles used in the present experiments, one (Leptospirillum MT6) was an obligate autotroph, and the other (Am. ferrooxidans), whilst having the capacity to fix CO2, is relatively ineffective at this and was originally classified as a heterotroph [16]. This was illustrated by the lack of pyrite oxidation by pure cultures of Am. ferrooxidans, in contrast to the mixed culture with At. caldus, where the organic C requirement of the iron-oxidizer was presumably provided as soluble compounds originating from the sulfur-oxidizer. Concentrations of DOC in the latter were greatest of those measured in any bioreactor culture; it might be expected that that not all of the cell exudates and lysates originating from At. caldus would be utilized by Am. ferrooxidans. In contrast, the mixed culture of the two iron-oxidizers was less effective at oxidizing pyrite, whilst the culture that also contained At. caldus was superior in this respect. Interestingly, the mixed Leptospirillum MT6/At. caldus culture was also a relatively poor pyriteoxidizing consortium (and was inferior to a pure bioreactor culture of Leptospirillum MT6 that was run on a separate occasion; data not shown). One possible reason for this is that these autotrophs competed for CO2, the solubility of which decreases with decreasing pH and increasing temperature, and is potentially a limiting factor to their growth. Interestingly, numbers (from plate counts) of both Leptospirillum MT6 and At. caldus were less than in the parallel culture that also contained Am. ferrooxidans (data not shown).
170
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2
At. caldus
19
Am.ferrooxidans
39
48
52
48
50
81
61
1.0E+10
70
**
***
60
1.0E+09
50 1.0E+08
40 30
1.0E+07
[DOC] (mg/l)
Microbial CFUs (/ml)
*
20 1.0E+06
10
(A) Am. ferrooxidans +At. caldus
1.0E+05 0
5
10
15
20
25
30
35
40
0 50
45
Time (days)
Leptospirillum MT6 Am.ferrooxidans
56
17
23
65
23
19 64 30
*
(B) Leptospirillum MT6 +At. caldus +Am. ferrooxidans
1.0E+08
12
21
38
34
At.caldus
1.0E+09
**
*** 25
20 1.0E+07
15 10
[DOC] (mg/l)
Microbial CFUs (/ml)
28
1.0E+06 5 1.0E+05
0
5
10
15
20
25
30
35
40
45
50
0
Time (days)
Figure 3. Microbial populations and DOC concentrations in consortia of moderate thermophiles. The pie graphs show data from FISH analysis, and the line graphs data from plate counts ( , Am. ferrooxidans; S, At. caldus; , Leptospirillum MT6), and DOC (♦). A variety of chemical reactions may occur in mineral leachate liquors that can either consume or generate protons. At the pH values at which the bioreactors were run, there was no hydrolysis of ferric iron, and no observable formation of jarosites or other secondary ferric iron minerals. The major acid-generating reaction was therefore the oxidation of sulfur and/or polythionates to sulfuric acid, primarily by At. caldus (though this may also be achieved abiotically with ferric iron). Measurements of the amounts of 171
Bioleaching Applications
alkali required to neutralize proton production in the pH-controlled bioreactors was therefore a useful indication of the extent of (microbial) sulfur oxidation. The fact that the greatest amounts of alkali were consumed in bioreactors that contained both At. caldus and Am. ferrooxidans (and which were the most effective leaching consortia) suggests that interactions involving transfer of carbon and oxidation of sulfur/polythionates were important in promoting pyrite oxidation. The combination of plate counts and FISH analysis to assess microbial populations gave useful insights into how these evolved during mineral oxidation, and also when subjected to low pH stress. In many cases, the relative abundances of the different bacteria were similar when analyzed by either method, further validating the use of “overlay” media to assess populations of acidophiles. On some occasions, however, no colonies of At. caldus were recovered, though FISH analysis confirmed that the sulfur-oxidizer was present. The reason for this is unclear, though it may be caused by the physiological state (stress etc.) of At. caldus in the leach liquors. What was apparent from these analyses was that, contrary to earlier reports, numbers of iron-oxidizers in mineral leachates often exceeded those of the “secondary” sulfur-oxidizers. It was also apparent that the most efficient pyrite-oxidizing consortia contained far greater numbers (from plate counts) of iron-oxidizers than the relatively inefficient bioreactor cultures that were run in parallel. One other notable fact from the present study was the finding that Am. ferrooxidans (strain ICP) is an important mineral-leaching organism, though only when grown in mixed cultures containing At. caldus. Mineral oxidation in cultures containing Am. ferrooxidans began far earlier and was more rapid than those where Leptospirillum MT6 was the primary iron-oxidizer. In cultures containing both iron-oxidizers, Am. ferrooxidans was frequently more numerous and presumably had a major role in promoting mineral dissolution. One potentially negative characteristic of Am. ferrooxidans, from the point of view of mineral leaching, is its lower tolerance of ferric iron than some other ironoxidizers [16]. This may be the reason for the cessation of pyrite oxidation beyond day 20 in the mixed culture of Am. ferrooxidans and At. caldus which, until that point, was the most efficient pyrite-oxidizing system of all those tested. This study has shown that some, though not all, mixed cultures of acidophiles are highly effective at oxidizing pyrite. One limitation of the study is that consortia permutations have involved only three moderate thermophiles. Additional experimental work, incorporating other iron-oxidizers such as Sulfobacillus spp. and Ferroplasma spp. would shed further insights into microbe-microbe and microbe-mineral interactions in such environments. ACKNOWLEDGEMENTS Naoko Okibe is grateful for financial assistance provided by The Institution of Mining and Metallurgy (U.K.), the Gen Foundation and Glaxo Ltd.. REFERENCES 1. K.B. Hallberg KB and D.B Johnson. Adv. Appl. Microbiol. 49 (2001) 37. 2. R.S. Golovacheva, O.V. Golyshina, G.I. Karavaiko, A.G. Dorofeev, T.A. Pivovarova and N.A. Chernykh. Microbiology (English translation of Mikrobiologiya) 61 (1992)1056. 3. N. Okibe, M. Gericke, K.B. Hallberg and D.B. Johnson, D.B. Appl. Environ. Microbiol. 69 (2003) 1936. 4. N.J. Coram and D.E. Rawlings. Appl. Environ. Microbiol. 68 (2002) 838. 172
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5. P.R. Norris, N.P.Burton and A.M. Foulis. Extremophiles 4 (2000) 71. 6. D.B. Johnson. FEMS Microbiol. Ecol. 27 (1998) 307. 7. M. Dopson and E.B. Lindstrom. Appl. Environ. Microbiol. 65 (1999) 36. 8. Yahya and D.B. Johnson. Hydrometall. 63 (2002) 181. 9. D.B. Johnson. J. Microbiol. Meth. 23 (1995) 205. 10. S. Kimura, K. Coupland, K.B. Hallberg and D.B. IBS 2003 Proceedings (submitted). 11. P.L. Bond and J.F. Banfield. Microb. Ecol. 41 (2001) 149. 12. K.J. Edwards, P.L. Bond and J.F. Banfield. Environ. Microbiol. 2 (2000) 324. 13. D.R. Lovley and E.J.P. Phillips. Appl. Environ. Microbiol. 53 (1987) 1536. 14. Schippers, P.G. Jozsa and W. Sand. Appl. Environ. Microbiol. 62 (1996) 3424. 15. D.E. Rawlings. Annu. Rev. Microbiol. 56 (2002) 65. 16. D.A. Clark and P.R. Norris. Microbiology 141 (1996) 785.
173
15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Biolixiviation of Cu, Ni, Pb and Zn using organic acids produced by Aspergillus niger and Penicillium simplicissinum R. Galvez-Cloutiera, C. Mulliganb and A.Ouattarac a
Associate Professor, Department of Civil Eng. Laval University, Quebec, Canada, G1K 7P4,
[email protected] b Associate Professor, Department of Civil Eng., Concordia University, Montreal, Canada, H3G 1M8 c Graduate student, Department of Civil Eng. Laval University, Quebec, Canada, G1K 7P4 R Abstract In order to assess extraction potential and identify the mechanisms of solubilisation, washing tests were performed using organic acids produced from the culture of Aspergillus niger and Penicillium simplicissimum. Two types of mining residues were studied: one coming from a Zn and Pb mine in New Brunswick, Canada and the second from a Ni mine in New Caledonia. Metals: Cu, Fe, Mn, Ni, Pb and Zn were in high concentration in the residues from the Canadian mine (from 1130 to 590,000 mg/kg of residues). Both residues are from different mineral backgrounds. For instance, Noranda residues are pyritic with a high sulfur concentration (267,569 mg/kg of residues). The results showed that various factors controlled the efficiency of the extraction. These factors arise from the mineral origin of residues and from the physico-chemical characteristics of the washing solution, which is a mixture of organic acids. Metal distribution and mineral origin of the residues were evaluated using SSE and SEM. The individual concentrations of acids constituting the mixture (citric, maltic, gluconic) were monitored using HPLC. These acids were produced under different periods of incubation, pH. Finally, in order to reduce the process cost, various organic residues were used as substrate (fruit skin or juices). Results show very encouraging efficiencies. The maximum solubilization obtained for Cu was close to 30% for the residues coming from New Caledonia. Keywords: heavy metals, biolixiviation, organic acids, geochemistry 1.
INTRODUCTION
1.1 The mining residues problem In Canadian mines, large amounts of soil and rocks that don’t contain commercially attractive amounts of precious metals (less than 0.5% w/w) will be exposed to the atmosphere. At the surface of the rock piles, wind, rainfall and other weathering factors (i.e. temperature) will produce exhaustive leaching of metals. These leachates are acid and contain high concentrations of heavy metals (e.g. cadmium, lead, copper) that may threat 175
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the health of plants, fish and humans. In Canada, it has been estimated that over 12,500 ha of tailings and 750 million tons of waste rock exist, and are potentially acid generating [2, 6]. Over the world, similar conditions are found. Notably, in mines in South America and Africa where the Canadian Mining Industry is largely present. To rehabilitate these acid generating tailings and waste rock dump sites in Canada, more than $ 10 billion dollars will have to be invested over the next 20 years [7]. Currently, the non-existence of reliable, predictable and cost-effective technologies challenges both the industry and the government management strategies to provide long term environmental protection against the potential risks of mining waste. Therefore, the valorization of mining residues could lead to important breakthroughs towards the sustainable use of mineral resources. 1.2 Current mining residues management strategies Until 1998, mining waste management strategies were mostly focused on the confinement of mining residues (e.g. underwater disposal, surface flooding, dry barriers and porous covers). Bioleaching processes have just recently been considered for metal recovery from mining residues. These bioprocesses are new, under development and, at most, are at batch scale with no proper technico-economical study been done. The preferred organisms for bioleaching are bacteria. For instance, experimental theory on the use of chemolithotrophic bacteria such as the Thiobacillus species for metal leaching has, to limited extent, been documented [1, 4]. Only few natural scale operations have been assayed. In most biological process applications, the process is developed under the exclusive study of the organism growth and its relationship with the aqueous medium in which it develops. Although, in the case of solid residues, geochemistry is fundamental to the success of microorganism growth it is almost never studied or taken in consideration. Bioleaching using fungi is originally new and agrees with sustainable development. Fungi bioleaching relies on the solubilization of metal by means of washing with the organic acids produced by the fungal activity. In many technical aspects, fungi growth is easier to achieve than bacteria. Various groups of native fungi microorganisms are known to produce organic acids. For instance, Aspergillus niger and Penicillium simplicissimum produce high complexing organic acids (e.g. citric, oxalic, gluconic) [5]. Fungi growth can be attained by providing optimal nutrient conditions and optimal substrates. Minerals contained in the mining residues can provide with excellent nutrients (e.g. Fe, Ca, Mg). When geochemistry of residues is known and taken in consideration, nutrients can be provided by the residues. Some organic wastes, such as some food or agricultural wastes, can be excellent providers of carbon. With some minimal waste manipulations carbon from these wastes can be easily available for fungi consumption. In the light of the expected shortage of non-renewable resources, increased efforts are absolutely necessary to seek new sources of raw materials with the aid of new or improved technologies. Bioleaching technology will enhance the release of precious metals to a concentration more commercially attractive and thus will give value to mining residues. The major innovation of this project lies in the facts that it will solve 3 important environmental problems: 1) reduce mining waste, 2) revalorization of mining residues and 3) revalorization of organic waste. Also, it will advance knowledge in organic acid production from fungi, it will incorporate fundamental geochemistry into the comprehension of the bio-leaching process and will systematically develop an industrial bio-leaching process from batch to bench, from column to pilot scale.
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2.
OBJECTIVES OF THIS STUDY The objectives of the study are to: (1) identify and evaluate the parameters (biological and geochemical) that affect the production of organic acids by Aspergillus niger and Penicillium simplicissimum, (2) study the potential use of new carbon sources for fungi growth and evaluate its impact on organic acid production, (3) study the influence and variation of metal geochemistry on the solubilization or immobilization of metals during bioleaching under two percolation setups (growth of fungi within the solid matrix, and growth of the fungi outside the solid matrix). 3.
SITES OF STUDY This study comprised two different solid matrices of two actual mining sites. The First one from New Brunswick, Canada, which is a ground mine of Zn and Pb and a Ni sky cut mine from New Caledonia both belonging to the Noranda Company. Table 1. Residues characteristics Results
Parameters
New Brunswick
New Caledonia
pH
7.7
5.6
Water content (%)
20.5
48.4
Organic Matter (%)
17.6
12.4
267,569
3,636
Cu
1,074
33
Fe
265,264
328,292
Mn
1,045
3,019
Ni
20
8,491
Pb
4,965
9
Zn
11,226
165
Total S (mg/kg)
Total Metal (mg/kg)
4.
MINING RESIDUE GEOCHEMISTRY Mine deposits are different in their geological and mineralogical nature. Their mineralogical composition determines their economical potential by determining the ease of precious metal extraction. Thus, any process seeking the extraction of metals from their solid matrix must take in consideration the geochemistry of its minerals and non-mineral components. Metals are retained in the soil under various forms as shown in figure 1. Manipulating pH, redox and acidity, a selective dissolution procedure can release in one step metal nutrients (e.g. Fe, Mg, Ca) and in a second step precious heavier metals (Pb, Zn, Cu, Ni, etc.). The first step can be done through a chemical washing procedure. The second step can be achieved using organic acids in an optimal ratio of concentrations and dosages (citric:oxalic:gluconic). Metal bioavailability greatly determines metal recovery rates. Also, it is known that heavy metals are in higher or lower scale, toxic to fungi (e.g. Ag > Hg > Cu > Cd >Cr >Ni >Pb >Co > Zn > Ca, Fe). In this sense, the presence of metals in either a soluble or exchangeable phase could significantly affect the biosynthesis of organic acids while using an in-situ bioaugmentation process. Thus, this toxicity needs to be controlled and 177
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studied. On the other hand, strongly sorbed metals could increase operation costs since more organic acids would be needed to dissolve solid phases contributing to metal retention. Metal speciation in mining wastes before and after bioleaching were determined by using a selective sequential extraction (SSE) method [3]. Figure 2 shows the initial metal distribution of the 2 mining residues before bioleaching.
Figure 1. Partitioning of metals in soils under aerobic conditions (from (8)).
New Brunswick
Fraction
100%
Résiduels M.Organique
50%
Ox./Hyd.
0%
Carbonates Cu
Fe
Mn
Ni
Pb
Zn
Metals
Echangeable Solubles
New Caledonia
Fraction
100% 50% 0% Cu Fe Mn Ni Pb Zn Metal
Résiduels M.Organique Ox./Hyd. Carbonates Échangeables Solubles
Figure 2. Metal distribution under natural pH for the 2 residues (Résiduel = Residual, M. Organique = Organic Matter, Oxide/Hydroxides, Carbonates, Échangeable = Exchangeables and Solubles)
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Bioleaching using fungi organisms is principally based on the following mechanisms: acidolysis (consisting of the solubilisation of the matrix by pH reduction), complexolysis (consisting of the complexation of metals by the excreted organic acids or amino acids and shown in Fig. 3), redoxolysis (the reduction of ferric iron which is mediated by the oxalic acid) and bioaccumulation of metals by the organism's mycelium. Metal bioleaching rates depend on acid concentration, complexation kinetics, contact time, pH, and metal geochemistry.
Figure 3. Complexation of heavy metals by organic acids 5.
BIOLEACHING PROCESS Bioleaching studies were funded by Natural Resources Canada and were recently performed at the environmental laboratory in Laval University. Organic acid production can be carried out within the mine residue or produced elsewhere in a reactor. Thus, in this study, organic acids and residues were mixed directly or indirectly using either beaker or column reactors as shown in Figure 4. Each option presents specific advantages and were tested in a preliminary activity. While in-situ bioaugmentation can accelerate the overall treatment process, ex-situ production of organic acids can facilitate the isolation of preferred compounds for more efficient application on the piles and can avoid difficulties related to keeping optimum culture conditions in the field. Indirectly, acids (supernatants) can be separated from fungi growth (biomass) and mixed with various concentrations of residues (1, 5, 7, 10 and 15% w/v). During preliminary experiments, fungus were grown in the presence of the same concentrations of residues up to 15 days. Various bioleaching results are presented in table 2 and figure 4. As it can be seen, indirect bioleaching give in some cases low extraction ratios than direct bioleaching, this is due to the purity and sterile condition in indirect bioleaching were organic acid production is optimized. Targeted metals for extraction were Cu and Pb from the New Brunswick mine and Cu, Ni and Zn from the New Caledonia mine.
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a) Acid production phase for indirect leaching
P. simplicissimum
Aspergillus niger
b) Direct and indirect bioleaching in column Microorganisms culture
Mining Residues
Direct Bioleaching
Indirect Bioleaching
c) Direct bioleaching in beaker
New Caledonia residues
Figure 4. Bioleaching layouts 180
New Brunswick residues
Bioleaching Applications
Table 2. Results in % mass extraction from column bioleaching New Brunswick
Cu Fe Mn Ni Pb Zn
Aspergillus Niger Direct Indirect 8.9 12.3 1.8 1.5 78.7 83.7 0.0 0.1 19.4 12.5 6.0 3.5
Penicillium simplicissinum Direct Indirect 9.4 31.6 1.9 1.4 73.6 80.6 0.0 0.0 18.4 30.5 7.7 9.1
Metal Cu Fe Mn Ni Pb Zn
Aspergillus Niger 12.5 31.3 0.2 0.4 35.4 25.6 63.5 51.1 2.8 100 27.0 75.7
Penicillium simplicissinum 3.1 25 0.09 0.0 20.7 5.6 39.8 16.1 35.6 2.8 43.2 10.8
Metal
New Caledonia
Concentration (mg/kg de sol)
Biolixiviation PEN + CAL
Cu Fe Mn Ni
012 10 216 14 618 4 8
Pb
Jours
Zn
Concentration (mg/kg de sol)
Biolixiviation PEN + NOR Cu Fe Mn Ni Pb Zn
7000 5000 3000 1000 -1000
0
5
10 Jours
15
20
Figure 4. Bioleaching results for Penicillium S.
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Results showed that organic acid production was not inhibited by the presence of mining residues, at least in the range of soil/liquid ratios under study. Preliminary tests were run to obtain optimal soil/liquid ratios for fungi growth. An analytical procedure combining HPLC and GC-MS was developped in order to dose the various organic acids produced. As shown in figure 5 Gluconic acid was produced at the highest concentration followed by citric acid. In a second part of the study, batch experiments evaluated continuous washing with the same initial organic acid load. After five washings, the metal efficiency extraction was: lead, 42%; cadmium, 96%; chromium, 85%; copper, 60% and zinc, 77%. Since pH influence the solubilization/ precipitation of some metal forms (exchangeable, oxides, carbonates), the study included a pH monitoring of both, acid and percolated metal containing effluent solutions. A slightly increase in metal extraction was observed with a decrease in the pH of the acid produced solution. It was also concluded that metal geochemistry played a significant role in metal bioleaching. Organic Acids AN
Concentration (%)
12 Acide gluconique
10
Acide Phytique Acide oxalique
8
Acide Tartarique
6
Acide Isocitrique Acide malique 1
4
Acide citrique Acide malique 2 Sucrose
2 0 37088
37090
37092
37094
37096
37098
37100
37102
Days Figure 5. Organic acids production as function of time for indirect bioleaching Indeed, the SSE results not only allowed to determine metal distribution within the soil matrix but also to conclude that carbonate and amorphous forms of metals were among the more difficult metal particulate species to destroy before extraction was possible. In revenge, exchangeable and oxides were more easily attacked by the acids. Also, a slight redistribution of metal particulate species after bioleaching was observed. We want to further explore the interrelations between metal species distribution, metal species re-adsorption and acid components (e.g. gluconic, citric) selectivity. By developing the application of the SSE method, the information gained will allow a better conception of reliable, and predictable methodologies for on-site metal bioleaching. 6.
USAGE OF ORGANIC WASTES AS SOURCE OF CARBON Parallel experiments were run under direct and indirect leaching conditions in order to evaluate different carbon sources. These experiments were run using a concentration of 10% residues and use only aspergillus niger’s produced acids. Various Carbon sources 182
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included molasses, corncobs, and brewery wastes. These experiments lasted for about three weeks. The results are shown in Figure 6. It is clear that sucrose solution gave the best leaching results under both direct and indirect conditions. The pH was low (pH 3.6) as in the previous set of experiments. Molasses yielded lower leaching results than the sucrose and better than the corncobs. Corncobs yielded reasonably good results (up to 24% copper removal) under indirect leaching conditions. They contain simple sugars that can be used by A. niger. Brewery waste gave poor results. The grains would require pretreatment to release simple sugars for it to be a suitable substrate. (a)
(b) 40
60
Sucrose Molasses
40
Corn cob Brewery waste
20
% Cu leached
% Cu leached
80
Sucrose Molasses Corn cob
20
Brewery waste
0
0 0
10 Time (days)
20
0
10
20
Time (days)
Figure 6. Effect of C source on Cu extraction using indirect (a) and indirect (b) conditions 7.
DISCUSSION AND CONLUDING REMARKS Indirect leaching was clearly more favorable for metal removal. When the flasks were not sterilized, faster growing bacteria took over, causing problems for the fungus and decreasing acid yields. Another important parameter is the degree of selectivity of the organic acid mixture for precious metals rather than for non-precious metals (e.g. Ca, Mg, Na, K). Therefore, the bioleaching process must be conceived in such way of maximizing precious metal extraction, minimizing non-precious metal extraction, optimal organic acids ratios within the mixture, optimal pH to avoid early metal precipitation and optimal To. The process must also include a pre-screening geochemical characterization that will determine the real potential of metal extraction. In conclusion, metal recovery by fungi bioleaching is feasible, however, further applied research must be performed to provide greater understanding on the nature of complexes and bioleaching selectivity, on the influence of metal geochemistry on bioleaching as well as on the reuse of organic acids. REFERENCES 1. Bosecker, Klaus (1997). Bioleaching: metal solubilization by microorganisms. FEMS Microbiology Reviews, n. 20, pages 591-604. 2. Feasby, G. and Tremblay, G.A. (1995). New Technologies to reduce environmental liability from acid mine generating mine waste. Proc. Sudbury ’95 – Mining and the Environment, 6443 Ottawa. CANMET. 3. Galvez-Cloutier, R. (1995). Study of heavy metal accumulation mechanisms in the Lachine Canal sediments. Ph.D. Thesis. McGill University, Montreal.
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4. Krebs, W. et al. (1997). Microbial Recovery of metals from solids. FEMS Microbiology Reviews, n. 20, pages 605-617. 5. Mulligan, C. et al. (1999). Biological leaching of copper mine residues by Aspergillus niger. Proc. of the Int. Biohydrometallurgy Symposium IBS’99. Madrid, Spain. Part A Bioleaching, Microbiology, pages 453-461. 6. Mind Environment Neutral Drainage Program (MEND)(1997). Annual Report. 202 pages. 7. Wheeland, K.G. and Feasby, G. (1991). Innovative decommission technologies via Canada’s MEND program. Proc. of the 12th Nat. Conference, Hazardous Materials. Control/Superfund ’91, Control Res. Inst., pages 23-38. 8. Galvez-Cloutier, R. Yong, R.N., Chan, J. and Bahout, E. (1995) Critical Analysis on Sediment Quality Criteria and Hazard-Risk Assessment Tools. Dredging, Remediation and Containment Contaminated Sediments ASTM STP No 1293, pp. 306-318.
184
15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Biooxidation of pyrite by Acidithiobacillus ferrooxidans in single- and multi- stage continuous reactors C. Canales, J.C. Gentina and F. Acevedo* School of Biochemical Engineering, Catholic University of Valparaiso, Av. Brasil 2147, Valparaíso, Chile Abstract The objective of this work was to study the extent of the bacterial attack on pyrite in the biooxidation of a refractory gold concentrate in continuous stirred tank reactors (CSTR). Two laboratory-scale biooxidation systems were installed. One of them consisted of a single stage 4-litre CSTR; the other one was a four-stage system with a total volume of 4 L. Acidithiobacillus ferrooxidans R2 was used; the concentrate, containing 66.7% pyrite, was suspended in 9K medium without ferrous sulfate. The single-stage CSTR was operated with residence times between 3.5 and 10 days, with Eh values of 600-650 mV. Up to 51% of iron solubilization was obtained, with negligible ferrous iron levels. This result suggests that the bacterial ferrous iron oxidation proceeded at a higher rate than the pyrite attack. The four-stage system operated with total residence times between 7 and 14 days, with maximum iron solubilization of 66% and Eh of 525 to 675 mV. Again, almost no ferrous iron was detected. Decreasing residence times had a large effect in diminishing the bacterial activity especially in the first stage because its low residence time. Oxygen demands measured in each stage revealed that decreasing total residence times caused a displacement of the main microbial activity towards the last stages. These results show that increasing residence times favor the multistage configuration, resulting in higher degrees of pyrite oxidation than in the single CSTR. This is because in the latter case the positive effect of residence time tends to saturation. Keywords: refractory gold concentrate, reactor configuration, CSTR, pyrite biooxidation 1.
INTRODUCTION Continuous biooxidation of refractory gold concentrates in tank reactors is currently a technologically and economically feasible technology that is being applied in several large-scale operations the world over [1, 2]. Reactors present significant advantages over heaps for bioleaching operations, in particular related to a more homogeneous reacting mass and the possibility of exerting close control on the main process variables. The biooxidation of gold concentrates, as well all other bioleaching processes, is as a whole an autocatalytic complex reaction. This autocatalytic character is given by the fact *
[email protected]
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that both cells and ferric iron act as reactants and products. This being the case, the optimal reactor configuration that minimizes the reaction volume for a given conversion is a continuous stirred tank reactor (CSTR) followed by a tubular reactor (TR) [3]. Conversely, given a defined volume, the multi-stage arrangement will render higher degrees of conversion. Because the need to maintain the solids in suspension and to supply oxygen and carbon dioxide rules out the operation of an actual TR, its kinetic behavior is simulated with a battery of CSTR’s connected in series [1, 4, 5]. This configuration is used in large-scale operations utilizing mesophilic bacteria [6] and has been tested at laboratory and pilot scale with extreme thermophiles [7, 8]. The objective of this work was to quantitatively characterize the behavior of a reaction system composed of four CSTR’s in series and compare it with the performance of a single CSTR of a volume equal to the total volume of the series arrangement. 2.
MATERIALS AND METHODS
2.1 Microorganism, concentrate and culture medium Acidithiobacillus ferrooxidans R-2 was used. This strain was cultivated in the presence of the same gold concentrate for over one year before performing these experiments. The refractory gold concentrate contained 15 g gold/tonne, 67.6% pyrite and 9.0% chalcopyrite. It contained 36.1% Fe, 39.7% S and 3.3% Cu. A fraction of particle size less than 75 µm was used. Before each run the concentrate was washed with a 10% v/v aqueous solution of acetone, rinsed with diluted sulfuric acid pH 1.8 and dried overnight at 100ºC. 9K medium was used, replacing the ferrous sulfate with the concentrate at a pulp density of 6% w/v. 2.2 Analytical methods Ferrous ion iron was measured by the modified o-phenanthroline method [9] and total soluble iron was determined by reducing the ferric ion to ferrous ion with hydroxylamine and assaying with phenanthroline. Ferric iron was calculated as the difference between the two. Sulfate was assayed by turbidimetry [10]. Eh was monitored with an Ag/AgCl probe and dissolved oxygen was measured with a polarographic probe. Oxygen consumption rate was determined in each stationary state by the gassing-out method [11]. 2.3 Bioreactor systems Two continuous biooxidation systems were set up, each one of a total working volume of 4 litres. One was a single 4-L continuous stirred tank reactor (CSTR) and the other consisted of four CSTR’s connected in series. The first stage has a working volume of 1.75 L and the other three, which represent a TR, are 0.75 L each. The reactors were made of acrylic plastic and each one has four baffles, heating jacket, variable speed agitator with one pitched-blade turbine pumping down and Eh, pH and dissolved oxygen probes. The CSTR’s had constant geometrical ratios of HL/T = 1.0 and D/T = 0.30. Aeration was supplied by means of a perforated annular sparger. The tanks were fed and discharged by peristaltic pumps. Because of the very low flow rates involved, the pumps were turned on and off periodically by a programmable switch. Fresh pulp was fed from an agitated feed tank.
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2.4 Operation conditions The reactors were operated at 33ºC, with agitation of 600 rpm and aeration of 0.5 volumes of air per volume of liquid per minute (vvm). pH was adjusted to 1.8 with sulfuric acid at the start of the batch phase operation. Pulp density was 6% w/v. These constant conditions were used in all runs of both biooxidation systems. The 4-L CSTR was operated at pulp residence times between 3.5 and 10 days. The four-stage system was operated at residence times in the range of 7 to 14 days. The reactors were operated continuously for at least three residence times before any measurements were done. The different steady states were always established by increasing the pulp flow rate. 3.
RESULTS AND DISCUSSION
3.1 Single stage reactor Figure 1 depicts the results of the operation of the single stage reactor. Under the experimental conditions used in this work, 50% of the pyrite was solubilized as revealed by the total soluble iron measurements. At all times the concentration of ferrous ion was negligible, a fact compatible with the high Eh values attained. The absence of Fe2+ is an indication that the rate of bacterial oxidation of ferrous ion was higher than the rate of ferrous production in the ferric leaching of the concentrate. The low Eh value at a residence time of 3.5 days could be due to the high solids content of the reactor due to the low extraction at that operation condition. This effect has been repeatedly observed by the authors under similar circumstances.
Figure 1. Continuous solubilization of pyrite concentrate in 4-L CSTR at 6% pulp density, 0.5 vvm and 600 rpm Table 1 presents the production rates, oxygen consumption rates and ferric/sulfate ratios obtained at the different residence times. Maximum rates are obtained in the range of residence times of 5.5-7 days, while the consumption rate of oxygen decreases steadily with increasing residence times. The Fe3+/SO42- ratio varies between 0.60 and 0.69, higher than the stoichiometric value of 0.47 predicted by equation 1. It can also be noted that the oxygen consumption rate decreases with increasing ferric/sulfate ratios. These results 187
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suggest that the pyritic sulfur is not completely oxidized to sulfate. This uncoupled behavior of the solubilization and oxygen consumption rates arises from significant changes in cell physiology mediated by the dilution rate. Table 1. Production rates, oxygen consumption rates and ferric/sulfate ratios in the single-stage reactor* Residence time, Fe production, SO42- production, O2 consumption rate, Fe3+/SO42(d) (g/L·d) (g/L·d) (g/L·d) w/w ratio 3.5 0.86 1.43 3.54 0.60 4.5 1.33 2.00 3.34 0.67 5.5 1.36 2.27 2.70 0.60 7 1.37 2.00 2.06 0.69 10 1.10 1.60 1.79 0.69 * Dissolved oxygen concentration was higher than 40% sturation at all operation conditions
4 FeS2 + 15 O 2 + 2 H 2 O → 2 Fe 2 (SO 4 )3 + 2 H 2SO 4
(1)
3.2 Multi stage system Steady state results are presented as a function of cumulative residence time in each reactor for the four total residence times considered. The dissolution of pyrite is presented in Figure 2. No ferrous ion was detected in the liquid. The first stage operated at residence times between 3 and 6 days. The increase of solubilization obtained in that range was over seven fold, significantly higher than the one obtained in the single CSTR under similar conditions (Figure 1). The solubilization that took place in the first CSTR increased steadily with residence time. Blank runs made with non-inoculated pulp showed that chemical leaching was almost nil, so the very low leaching that occurred at 3 days residence time must be due to the activity of a small bacterial population, as revealed by microscopic examination. By the other hand, the contribution of the three 0.75-L reactors was more important at low residence times than at high ones. As a whole, these data confirm the importance of the multi-stage design, pointing to an adequate operation residence time around 9 days. It is worth noting that in the intermediate range of residence times such as 8 or 9 days, highest solubilizations were obtained with two and three stages rather than with the four reactors. This result could be influenced by the fact that some solids accumulated in each stage because of difficulties in attaining a homogeneous pulp because of the small scale of the reaction system. Biooxidation expressed as percent iron extraction is shown in Figure 3. A maximum extraction of 66% is obtained at a cumulative residence time of 14 days. As a rule, the multi-stage system rendered higher extractions than the single vessel. This observation is consistent with the fact that the Eh values were also higher, as can be seen in Figure 4 as compared to Figure 1.The production of sulfate, presented in Figure 5, shows a similar pattern to the iron extraction, although the advantage of the four-stage system is apparent at lower residence times.
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15
80
Fe extraction, %
Soluble iron, g/L
12 9 6 3
60
40
20
0
0 0
3
6
9
12
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15
Cumulative pulp residence time, d
3
6
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Cumulative pulp residence time, d
Figure 2. Continuous solubilization of pyrite concentrate in four-stage system at 6% pulp density. ♦: reactor 1; ■: reactor 2; ▲: reactor 3; x: reactor 4
Figure 3. Iron extraction in the four stage system as a function of cumulative residence time
30 Sulfate concentration, g/L
700
Eh, mV
600
500
400
25 20 15 10 5 0
0
3
6
9
12
15
Cumulative pulp residence time, d
Figure 4. Eh as a function of cumulative residence time in the four-stage system. ♦: reactor 1; ■: reactor 2; ▲: reactor 3; x: reactor 4
0
3
6
9
12
15
Cumulative pulp residence time, d
Figure 5. Sulfate production in the fourstage system. ♦: reactor 1; ■: reactor 2; ▲: reactor 3; x: reactor
The biological oxidation generated ferric/sulfate ratio values of 0.45 to 0.68, lower than the ones obtained in the single reactor unit, pointing to a more complete oxidation of sulfur in the complex arrangement. 189
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The production rates of each stage are presented in Table 2. Except for stage 1, the rates increase with decreasing total residence times. The residence time in stage 1 at total residence time of 7 days across the system is 3 days; this is probably a too short time and did not allow the establishment of a stable cell population of a sufficient size. Data from Table 2 regarding sulfate production rate suggest that the bacterial activity is displaced successively from the first to the following stages as the residence time decreases. This effect can also be appreciated in Table 3 that shows the oxygen consumption rates. From the results of Tables 2 and 3 it can be concluded that the main variation in the behavior of the system is produced in the transition from 7 to 8.5 days of total residence time. The cumulative production rates of iron and sulfate in each stage are presented in Figures 6 and 7. Iron solubilization is less affected by residence time in stages 3 and 4 than in the first two stages, while sulfate production rate in each reactor increases up to a maximum and then decreases steadily with residence time. Cell activity is proportional to the slope of the curves of each stage. Similar behavior can be seen in the first three reactors, in which most of the pyrite solubilisation takes place. This saturation pattern could be due to the exhaustion of available active sites on the surface of the concentrate particles [12]. Table 2. Iron and sulfate production rates in the four-stage system (g/L·d)* Residence Stage 1 Stage 2 Stage 3 Stage 4 time** Fe SO42Fe SO42Fe SO42Fe SO42(d) 7 0.43 0.67 1.67 3.05 1.90 3.81 1.52 4.57 8.5 0.86 1.61 1.25 1.88 1.13 2.51 1.25 4.39 10 1.19 2.51 1.44 2.13 1.01 2.13 1.81 1.60 14 1.63 2.45 0.76 1.14 0.80 1.53 0.11 1.14 * Calculated as the ratio of the increase of production in each stage to the residence time. Dissolved oxygen concentration was higher than 40% saturation at all operation conditions ** Considering the total reaction volume of 4 litres
Table 3. Oxygen consumption rate in the four-stage system (g/L·d) Residence time* (d)
Stage 1
Stage 2
7 0.32 1.41 8.5 1.91 1.58 10 1.98 1.68 14 1.80 1.75 * Considering the total reaction volume of 4 litres
190
Stage 3
Stage 4
3.12 1.47 1.48 1.50
2.66 1.07 1.41 1.30
Bioleaching Applications
14 Sulfate production rate, g/L/d
Fe solubilization rate, g/L/d
8 7 6 5 4 3 2 1
12 10 8 6 4 2 0
0 0
3
6
9
12
15
Cumulative pulp residence time, d
Figure 6. Cumulative iron solubilization rates in the four-stage system. ♦: reactor 1; ■: reactor 2; ▲: reactor 3; x: reactor 4
0
3
6
9
12
15
Cumulative pulp residence time, d
Figure 7. Cumulative sulfate production rates in the four-stage system. ♦: reactor 1; ■: reactor 2; ▲: reactor 3; x: reactor 4
4.
CONCLUSIONS The biooxidation of a pyritic refractory gold concentrate has been compared using a single stage CSTR and a reactor arrangement consisting of four CSTR’s in series of equivalent volume. It is concluded that increasing residence times favor the multistage configuration, resulting in higher degrees of pyrite oxidation than in the single CSTR.
ACKNOWLEDGMENTS This work was supported by the National Commission of Science and Technology through the FONDECYT project 1000284. REFERENCES 1. F. Acevedo F, Electronic J. Biotechnol., 3 (2000) 184. Available from: www.ejb.org, ISSN 0717-3458. 1. J.A. Brierley and C.L.Brierley, Hydrometallurgy, 59 (2001) 233. 2. O. Levenspiel, Chemical Reaction Engineering, 3rd ed., J. Wiley & Sons, New York, 1998, chap. 6. 3. D.H. Dew. In: T. Vargas, C.A. Jerez, J.V. Wiertz and H. Toledo (eds.) Biohydrometallurgical Processing, Vol.1. Universidad de Chile, Santiago, Chile. p. 239, 1995. 4. R. González, J.C. Gentina and F. Acevedo. In: Proceedings XIII National Congress of Chemical Engineering, Antofagasta, Chile, 18–21 October, 1999. 5. D.E. Rawlings (ed.), Biomining. Theory, Microbes and Industrial Processes, Springer Verlag, New York, 1997. 6. M. Gericke, A. Pinches and J.V. van Rooyen, Int. Miner. Process., 62 (2001) 243. 191
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7. P. d’Hugues, S. Foucher, P. Gallé-Cavalloni and D. Morin, Int. Miner. Process., 66 (2002) 107. 8. L. Herrera, P. Ruiz, J.C. Aguillon and A. Fehrmann, J. Chem. Technol. Biot., 44 (1989) 171. 9. Instituto de Hidrología de España, Análisis de aguas naturales continentales, 32. Centro de Estudios Hidrográficos, Madrid, 1980. 10. R. Jurecic, M. Bwerovic, W. Steiner and T. Koloini, Can. J. Chem. Eng., 62 (1984) 334. 11. G.F. Andrews, Biotechnol. Bioeng., 31 (1988) 378.
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Chemical chalcopyrite leaching and biological ferric solvent production at pH below 1 P.H.-M. Kinnunena, V.L.A. Saloa,b, S.O. Pehkonenb and J.A. Puhakkaa a
Institute of Environmental Engineering and Biotechnology, Tampere University of Technology, P.O.Box 541, FIN-33101 Tampere, Finland b Department of Chemical and Environmental Engineering, National University of Singapore, 10 Kent Ridge Crescent, 119260 Singapore
Abstract Chalcopyrite ferric leaching experiments were conducted at pH 1.0 to 2.5 at 50°C and 65°C under air and nitrogen atmospheres. Copper leaching yields were higher (97%) at pH 1.0 than at pH 1.8-2.5 at both temperatures. Increase in the temperature increased the initial rate of leaching, but resulted in lower yields indicating high jarosite precipitation rates. The composition of the gas phase did neither affect redox potentials nor leaching rates. Jarosite and iron hydroxide precipitates were not formed and copper yields increased by using a ferric solution at pH 1.0. In tank reactor, similar copper leaching yields were obtained with different ferric supply regimes. Biological generation of ferric solution at low pH was studied in batch and continuous-flow fluidized-bed reactors (FBR). In batch assays at 35°C, biological iron oxidation rate was not affected by pH of 0.9-1.5 but started to decline at 0.7. The pH of the FBR was gradually decreased to 0.9 without changes in iron oxidation; the maximum iron oxidation rate at pH 0.9 was 10 g Fe2+ dm-3 h-1. The results indicate that biologically produced ferric solvent at pH 0.9 results in high chalcopyrite leaching yields. Keywords: chalcopyrite, ferric leaching, iron oxidation, passivation 1.
INTRODUCTION One of the major problems in hydrometallurgical applications is the formation of a hindering diffusion layer on the mineral surface or a contact hindrance between the leaching solution, microbes and mineral. A better understanding of the surface speciation under leaching conditions is a key factor in improving dissolution kinetics and yields in bioleaching. Passivation layer in sulphide mineral leaching with ferric iron has been proposed to consist of iron hydroxy precipitates or elemental sulphur formed as end products [1,2,3]. Polysulphides may act as transient intermediates of leaching [1]. Iron in the leaching solution precipitates as iron oxides or basic iron salts like jarosite [2,3,4]. Leaching conditions such as pH, temperature and ionic composition and concentration of the medium affect the formation of iron hydroxy precipitates [5]. Ferric iron precipitation diminishes the available ferric iron in the leach solution, forms kinetic barriers and tends to block pumps and valves. The iron precipitation highly 193
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depends on the pH and temperature. [5] Thiobacillus ferrooxidans and Leptospirillum ferrooxidans are the most important mesoacidophilic microorganisms involved in ferrous iron oxidation. The optimum pH for bacterial iron oxidation is generally 2.0 to 2.5. [6] L. ferrooxidans tolerates lower pH values than T. ferrooxidans, which usually does not survive below a pH of 1.0 [5,6]. Even though microorganisms may adapt to physicochemical changes such as pH in their environment, there are limits to the extent to which this may occur [7]. In this study the influence of temperature, pH and atmospheric oxygen on indirect ferric leaching rates and surface precipitates were studied. Since the jarosite precipitation can be prevented at a very low pH, the potential of biologically producing ferric solvent at low pH was studied in batch and continuous flow reactors. The maximum iron oxidation rate at pH of 0.9 was determined. Further, different ferric leach solutions were used to maximize the copper leaching and minimize the iron precipitation. 2.
MATERIALS AND METHODS
2.1 Influence of pH on the leaching rate The influence of pH on the leaching rate of chalcopyrite (Outokumpu Ltd, Pyhäsalmi, Finland, Cu 25.4%; Fe 27.9%) was studied at 50°C and 65°C in 150 cm3 erlenmeyer flasks using 3% (w/v) solids concentration. Leaching solution consisted of 0.4 g dm-3 (NH4)2SO4, 0.25 g dm-3 KH2PO4.2H2O, 0.25 g dm-3 MgSO4.7H2O, 0.02 g dm-3 yeast extract and a stoichiometric amount of ferric iron as Fe2(SO4)3 (Eq. 1). (Eq. 1) CuFeS2 + 4 Fe2+ Æ Cu2+ + 5 Fe2+ + 2 S° The pH was adjusted with H2SO4 or NaOH. Experiments under nitrogen atmosphere were conducted in order to study the influence of atmospheric oxygen on the precipitation. The leaching solution was sparkled with N2 before the addition of chalcopyrite (15 min) and after sampling (5 min). Samples (5 mL) were centrifuged (5000 rpm, 12 min) and the supernatants were used for analyses. Dissolved copper and iron concentrations were analyzed by inductively coupled plasma connected to mass spectrometry (ICP-MS) and the ferrous iron concentration spectrophotometrically by ferrozine method [8]. Ferric iron concentration was determined as the difference of total and ferrous iron concentration. The pH was measured with WTW Sentix 42 electrode and redox potential an ORION combination electrode 9678BN. Samples for SEM/EDX and XRD were centrifuged (4000 rpm, 3 min) and pellets were dried at 40°C. SEM/EDX-analyses were conducted with JEOL JSM-5600LV scanning electron microscope. In addition some of the residues were examined with x-ray diffraction (Cu anode, Kα radiation, λ = 1.54). 2.2 Effect of the ferric addition on chalcopyrite leaching The experiments were conducted in 2 L stirred tank reactors (Figure 1). The reactor conditions were as follows: liquid volume 1.5 L, solids concentration 10% (w/v), and temperature 50°C. The leach solution and a stoichiometric amount of Fe2(SO4)3 was added in either in 1, 2 or 5 shares. The pH of the leach liquor was adjusted to 1.5 with H2SO4 and NaOH at the beginning and at the sampling time. Redox potential was measured using Hamilton Pt-ORP electrode and pH using Orion model SA720 pH -meter after filtration (0.45 µm). The concentrations of dissolved copper and iron were measured using ICP-MS. The ferrous iron and SEM/EDX-analyses were performed as described above. The evaporation losses were accounted for in the results.
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Figure 1. Schematic diagram of 2L agitated reactor for ferric leaching. 1) Hot water inlet from water bath, 2) hot water outlet, 3) mixing device, 4) mixing blade, 5) baffle, 6) sampling socket. 2.3 Influence of pH on iron oxidation The influence of pH on iron oxidation was studied in duplicate 120 mL serum bottles in 60 mL of total volume at 150 rpm in a rotary shaker at 35°C. The mineral medium was autoclaved (121°C, 0.1 MPa, 20 min) before the addition of FeSO4.7H2O after sterile filtration (0.2µm). The final concentration of the mineral medium was 0.35 g dm-3 (NH4)2HPO4, 0.05 g dm-3 K2CO3 and 0.05 g dm-3 MgSO4. The pH was adjusted with H2SO4 to 0.5, 0.7, 0.9, 1.1, 1.3 and 1.5 with 7 g dm-3 Fe2+ in the solution. The bottles were inoculated with 1 mL of an enrichment culture of iron-oxidizers from a fluidized-bed reactor [9]. In the chemical control bottles, sterile H2O was used instead of inoculum. After batch assay determinations of the pH tolerance limits, the pH in the fluidizedbed reactor with activated carbon as the biomass carrier [9] was gradually decreased to 0.9. The composition of the feed solution was 34.75 g dm-3 FeSO4.7H2O, 0.35 g dm-3 (NH4)2HPO4, 0.05 g dm-3 K2CO3 and 0.05 g dm-3 MgSO4 in tap water at pH of 0.9. The maximum iron oxidation rate was determined in the fluidized-bed reactor at the pH of 0.9 with 21 g dm-3 ferrous iron in the feed solution. The ferrous iron concentration was determined using the Shimadzu UV 1601 spectrophotometer by the colorimetric ortho-phenantroline method [10] modified as follows: 2 mL of 1,10-phenantroline (10 g/L) and 1 mL of ammonium acetate buffer were added to 3 mL of sample. Dissolved oxygen and temperature measurements were made using WTW OXI96 meter at the sampling time. 3.
RESULTS
3.1 The influence of pH, temperature and gas phase composition on leaching and precipitation Leaching experiments were conducted at different pH values under air and nitrogen atmospheres at 50°C and 65°C (Figure 2). Copper yields increased with decreasing pH; at 50°C they were 95%, 35% and less than 10% at pH 1.0, 2.0 and 2.5, respectively. At 195
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65°C, the initial rate of leaching was higher, but resulted in lower yields than at 50°C indicating rapid precipitation. Iron precipitation was significant at pH of 1.8-2.5, which affected the leaching efficiency (Figure 3). At pH 1.0, the ferric iron did not precipitate, but was reduced to ferrous iron during leaching. Redox potentials decreased in the course of leaching, being highest at low pH values. The composition of the gas phase did not affect the redox potentials or leaching rates.
Figure 2. Copper leaching yields from chalcopyrite at 50°C (right) and at 65°C (left) In all experiments, SEM/XRD revealed jarosite precipitation layers on the surfaces of the CuFeS2. The layer consisted of large and small regular-shaped cubes with the average composition as presented in Table 1. The mineral surface was much less covered by the precipitation layer at pH 1.0 than at pH 1.5 and above (Figure 4). The surface of the ore (25% Cu, 16% Fe, 33% S) was partly covered by amorphous sulphur rich layer (60-97% S) typically formed under the cubic jarosite precipitates. Temperature and pH influenced the leaching yields, but not the composition of the residue-precipitates. Table 1. The range of elemental-% composition (EDX) of separate layers on CuFeS2 surfaces after 17 to 19 days of ferric leaching at pH 1.8-2.5
O, % S, % Fe, % Cu, % Na/K %
196
Surface composition after leaching
Mineral surface, unleached
Theoretical composition of jarosites
Larger precipitates
Smaller cubes
Typical mass precipitation
--25 11-22 14-24 ---
44-50 13 33-35 --~0
51-63 26-31 4-18 0-1 --
63-77 7.5-13 10-17 0 3-6
55-75 8-12 14-22 --3-4
Bioleaching Applications
40
A
g/L
30 20 10 0 0
5
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30 20 10 0
Figure 3. The fate of iron in chalcopyrite leaching at 50°C under the air atmosphere at A) pH 1.0, B) pH 1.8 and C) pH 2.0. (▲) calculated dissolved iron concentration (Fe from Fe3+-solvent plus Fe dissolved from CuFeS2), (x) dissolved iron (measured concentration), (♦) Fe2+ concentration, (■) Fe3+ concentration A
B
Figure 4. The mineral surface shown by scanning electron microscopy after 17-19 days of ferric leaching at 50°C A) at pH 1.0 and B) at pH 1.8 197
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3.2 Effect of the ferric addition on chalcopyrite leaching Ferric iron was added to the 50°C stirred tank reactor at pH 1.6 in five, two or one doses and resulted in similar copper leaching yields (approximately 40%) with all supply regimes ferric iron becoming the limiting factor (Figure 5). The initial leaching rate was highest, and redox potential and pH decreased fastest when ferric solution was added in one dose in the beginning. The iron precipitation was highest with one dose and lowest with five doses. SEM/EDX revealed elemental sulphur layer on the surface of the mineral at the end of the experiment, but showed also mineral surfaces without passivating layer (Figure 6).
Figure 5. The yield, pH and redox potential with 1 (♦), 2 (▲) and 5 (■) ferric solution additions (shown by arrow). Dissolved iron (◊), Fe3+ (∆), Fe2+ (□) and (x) calculated dissolved iron (Fe from Fe3+-solvent plus Fe dissolved from CuFeS2) concentrations. Addition of ferric solution in five (A), two (B) or one (C) doses
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CuFeS2
S0
Figure 6. Sulphur rich layer on the leached CuFeS2 surface at the end of the ferric addition experiment (day 15) 3.3 Effect of pH on iron oxidation The iron oxidation rate in the batch experiments remained unaffected at pH of 0.9-1.5 but started to decline at 0.7 (Figure 7). The lag phase of three weeks occurred at pH 0.5 and iron oxidation was strongly affected. The pH of the iron oxidizing fluidized-bed reactor was gradually decreased to 0.9 without changes in performance (Figure 8). The maximum iron oxidation rate at pH 0.9 was 10 g Fe2+ dm-3 h-1 (Table 2). The results show that the high-rate ferric production and regeneration in the FBR was possible at pH 0.9. 8 7
Fe2+ (g/L)
6 5 4 3 2 1 0 0
10
20
30
40
50
60
Time (days)
Figure 7. Influence of pH on ferrous iron oxidation by the fluidized-bed reactor enrichment culture in a batch assay. pH 0.5 (■), pH 0.7 (●), pH 0.9 (x), pH 1.1 (*), pH 1.3 (○), pH 1.5 (▲), pH 0.5 uninoculated control (□) and pH 1.5 uninoculated control (∆).
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2,5
1,9 1,7 1,5
1,5
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pH
Fe2+ (g L-1h -1)
2
1,1
1
0,9 0,5
0,7
0 0
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0,5 120
Time (d)
Figure 8. The influence of pH decrease on iron oxidation rate in a fluidized-bed reactor. pH in FBR (■), pH in feed (□), load g Fe2+ dm-3 h-1 (∆) and oxidation rate g Fe2+ dm-3 h-1 (*). Table 2. Iron oxidation rate in the fluidized-bed reactor at 35°C. pH of the feed solution 0.9
Period 1 Period 2 Period 3
4.
Length of the period (days) 10 20 14
Number of data points
Fe oxidation max (%)
Fe oxid. min-max (g Fe2+ dm-3 h-1)
Fe oxid. mean (g Fe2+ dm-3 h-1)
7 4 3
94.6 99.2 99.5
7.3-9.7 9.7-10.4 9.2-10.4
8.1 10.0 9.8
DISCUSSION In this work, jarosite precipitates were typically found on top of the sulphur rich layer indicating the passivation of the mineral surface first by sulphur followed by jarosite precipitates. Increase in temperature from 50°C to 65°C increased the initial copper leaching rate, but resulted in reduced copper yields. This was likely due to faster precipitation at high temperature. Chalcopyrite leaching with ferric iron was initially fast regardless of pH, but slowed down or completely stopped due to intense precipitation. Decrease of the leaching pH resulted in less precipitates and copper yields close to 100%. Leaching proceeded similarly under air and nitrogen atmospheres. The results of this work demonstrated that stepwise adding of ferric iron did not improve the copper yields. For biological ferric regeneration, the pH of the iron oxidation reactor needs to be maintained low. In general the mesophilic iron oxidizers do not survive below pH 1.0 [5, 7], at which the ferric leaching experiments indicated copper yields close to 100%. The other mesophilic iron oxidation studies have been carried out at pH of 1.1-3.2 [11-23]. In this study, the iron oxidation rate remained unaffected at pH of 0.9-1.5. The maximum iron oxidation rate in a fluidized-bed reactor dominated by Leptospirillum like organisms (Kinnunen et al., paper in preparation) at pH 0.9 was 10 g Fe2+ dm-3 h-1, which was similar to the iron oxidation rate at pH 1.4 in the same reactor (8.2 g Fe2+ dm-3 h-1), when air was used for aeration [9]. The iron oxidation rate of this study compared favourably with the mesophilic iron oxidation rate (0.9 g Fe2+ dm-3 h-1) in the fluidized-bed reactor with
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activated carbon as the carrier material at pH of 1.35-1.5 [23]. Ferric regeneration at 35°C was chosen to this study, because the moderately thermophilic iron oxidation rate at 60°C was considerably less than that of mesophilic [24]. In conclusions, the combination of high-rate biological ferric production at 35°C and pH 0.9 and below followed by chemical chalcopyrite leaching at 50°C to 65°C is promising for chalcopyrite leaching. REFERENCES 1. A. Parker, C. Klauber, H.R. Watling and W. van Bronswijk, Proc. Int. Biohydrometall. Symp. (2001), part B. ed. by V.S.T. Ciminelli and O. Garcia, Elsevier, 547-555 2. C. Klauber, A. Parker, W. Bronswijk and H.R. Watling, Int J Miner Process 62 (2001) 65-94 3. M.B. Stott, H.R. Watling, P.D. Franzmann and D. Sutton, Miner Eng 13 (2000) 11171127 4. S. Prasad and B.D. Pandey, Miner Eng 11 (1998) 763-781 5. M. Nemati, S.T.L. Harrison, G.S. Hansford and C. Webb, Biochem Eng J 1 (1998) 171-190 6. K. Bosecker, FEMS Microbiol Rev, 20 (1997) 591-604 7. D.B. Johnson, Hydrometallurgy 59 (2001) 147-157 8. L.L. Stookey, Anal Chem 42 (1970) 779-781 9. P.H.-M. Kauppi, H.J. Hautakangas and J.A. Puhakka, Proc. Int. Biohydrometall. Symp. (2001), part A. ed. by V.S.T. Ciminelli and O. Garcia, Elsevier, 385-392 10. Anonymous, Standard Methods for the Examination of Water and Wastewater. 18th ed. by A.E. Greenberg, L.S. Clesceri and A.D. Eaton, 1992, American Public Health Association 11. A.W. Breed and G.S. Hansford, Biochem Eng J 3 (1999) 193-201 12. M.J. Garcìa, I. Palencia and F. Carranza, Process Biochem (1989) 84-87 13. H.R. Diz and J.T. Novak, J Env Eng 125 (1999) 109-116 14. D.G. Karamanev and L.N. Nikolov, Biotechnol Bioeng 31 (1988) 295-299 15. M. Nemati and C. Webb, Appl Microbiol Biot 46 (1996) 250-255 16. H. Olem and R.F. Unz, Biotechnol Bioeng 19 (1977) 1475-1491 17. D.G. Karamanev, J Biotechnol 20 (1991) 51-64 18. M. Nemati and C. Webb, Biotechnol Bioeng 53 (1997) 478-486 19. S. Sandhya and R.A. Pandey, J Environ Sci Heal A27 (1992) 445-461 20. Mazuelos, I. Palencia, R. Romero, G. Rodríguez and F. Carranza, Miner Eng 14 (2001) 507-514 21. Mazuelos, R. Romero, I. Palencia, N. Iglesias and F. Carranza, Miner Eng 12 (1999) 559-564 22. Mazuelos, F. Carranza, I. Palencia and R. Romero, Hydrometallurgy 58 (2000) 269275 23. S.I. Grishin and O.H. Tuovinen, Appl Environ Microb 54 (1988) 3092-3100 24. P.H.-M. Kinnunen, W.J. Robertson, J.J. Plumb, J.A.E. Gibson, P.D. Nichols, P.D. Franzmann and J.A. Puhakka, Appl Microbiol Biot 60 (2003) 748-753
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Comparative study of the bioleaching of two concentrates of chalcopyrite using mesophilic microorganisms in the presence of Ag(I) A. López Juárez and R. E. Rivera Santillán* Departamento de Ingeniería Metalúrgica, Facultad de Química, UNAM. Ciudad Universitaria, México D. F. 04510, México Abstract A bioleaching study of two copper concentrates from Sonora and Zacatecas, (Mexico), containing basically chalcopyrite, was performed. X-ray diffraction studies, XRD, showed certain variations in the composition of the concentrates. One of them, C1, showed a greater secondary copper sulfide presence and pyrite (23.2% Cu and 20.6% Fe). The concentrate C2, practically did not have secondary copper sulfides and it had very little pyrite amount (21.85% Cu and 31.14% Fe). Mesophilic microorganisms at 35ºC in presence and absence of Ag(I) as catalytic agent were used. The Ag(I) ion showed an important catalytic effect on concentrate C1 bioleaching, whereas on concentrate C2 the effect was not noticed, and an smaller copper extraction was observed. The copper extraction increased doping C2 concentrate with pyrite and chalcocite. Keywords: bioleaching, chalcopyrite, pyrite, chalcocite, catalytic effect, mesophilic 1.
INTRODUCTION The environmental requirements imposed on pyrometallurgical processes of sulfide mineral concentrates in the copper industry have forced the development of hydrometallurgical routes as alternatives for the conventional treatment of sulfide minerals concentrates in order to avoid the SO2 production. Those processes involve sulfide oxidation either to sulfur or sulphate using oxidating agents such as O2 or Fe(III) ions or by a direct anodic oxidation in an electrolyte. This oxidation can be considered as an electrochemical reaction (1), with the cathodic reduction of the oxidant and the anodic oxidation of sulfide (2). The first idea that the existing chemical interactions on the surface of minerals could be of electrochemical nature was proposed by Salamy and Nixon (3). Bacterial leaching is an economical and widely used method for metal extraction, but until now its application has been limited to low metallic content minerals (4). During the last two decades, this process has been successfully applied to refractory mineral of Au and Ag treatment, and recently has been used for cobalt recovery from pyritic material in Uganda (5). * Corresponding author. Main address: Depto de Ingeniería Metalúrgica. Circuito Institutos s/n., Facultad de Química UNAM. Ciudad Universitaria. México D. F. 04510. México. E-mail:
[email protected] Phone +525556225241, fax: +525556225228.
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Nevertheless, in many cases, the slow kinetic of the biooxidation processes has limited its commercial application. This slowness is attributed to different parameters such as biological, physicochemical, electrochemical and mineralogical factors (6). Among mineralogical factors affecting the bioleaching systems is the initial composition of the material that, will cause that the systems will respond different ways to the same treatment conditions. As an attempt to accelerate the kinetics of chalcopyrite dissolution, the use of catalytic agents has been proposed. The catalytic effect of silver, Ag(I), during the chalcopyrite leaching has been reported (7, 8, 9). The improvement in the dissolution rate is attributed to the formation of a film of Ag2S on the surface of chalcopyrite particles (8). According to the semi-conducting characteristics of sulfide minerals (6), the electrochemical interactions (galvanic pairs formation) originated among different sulfide minerals in a same bioleaching system could improve the selective dissolution of the most active minerals. Although these interactions have been known for some time, it is complicated either to explain or predict their effects on the bioleaching rate in a sulfide mineral mixture (10). In this work the results of bioleaching tests with mesophilic microorganisms, in absence and in the presence of Ag(I) of two concentrates of chalcopyrite, as well as the results of bioleaching of the concentrate C2 doped with pyrite and chalcocite, also in absence and in the presence of Ag(I) are presented. 2.
MATERIALS AND METHODS
2.1 Minerals Two concentrates of chalcopyrite from Sonora and Zacatecas (Mexico), C1 and C2, respectively, were used. The samples of C2 were doped with 10% of pure minerals of pyrite (Py) or chalcocite (Ct), also from Zacatecas, Mexico. 2.2 Cultures The mesophilic microorganisms used were mixed cultures obtained from the own microflore of the concentrates, adapted by successive steps and developed in 100 mL of nutrient medium and 5g of concentrate. Three successive steps were carried out. 2.3 Bioleaching tests The bioleaching studies were made in an orbital incubator with temperature and stirring controlled to 35ºC and 150 rpm, respectively. The bioleaching tests were made with 90 mL of medium, 10 mL of inoculum and 5g of concentrate. All the tests were made in Erlenmeyer flasks of 250 mL. 2.4 Monitoring and control techniques Periodic measurements of pH and redox potentials, ORP, were made. The pH was fixed to the necessary value by addition of a diluted solution of H2SO4. The bacterial growth evolution was obtained determining the cellular concentration in solution samples counting the cells using an optical microscope with a Neubauer chamber. The analysis of metallic values in solution (Cu and Fetot) was conducted by atomic absorption spectrophotometry. The main mineral phases present in the concentrates were identified by means of xrays diffraction, XRD. 204
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3.
RESULTS AND DISCUSSION
3.1 Chemical analyses and characterization of the samples The chemical analyses of the resulting solutions of an acid attack to the concentrates allowed determining the following concentrate composition: Table 1. Chemical composition of the concentrates C1: 23.20% Cu, 20.60% Fe C2: 21.85% Cu, 31.14% Fe
Studies of XRD showed variations in the initial composition of the concentrates, exhibiting one of them, C1, a greater secondary copper sulfide and pyrite presence. The other concentrate, C2, did not have secondary copper sulfides, and contained very little pyrite and an important amount of arsenopyrite, figures 1A and 1B.
Figure 1A. Diffractogram of C1, and Figure 1B. Diffractogram of C2. (Cv: covellite, SiO2: silica, Cp: chalcopyrite, Py: pyrite y As: arsenopyrite) In figures 1A and 1B a clear difference between C1 and C2 is observed. With respect to the secondary copper sulfide, covellite, it was present in the C1 sample but not in the C2 sample. The pyrite presence in C1 was greater than in C2. The presence of arsenopyrite in C2 was not detectable in C1. Diffractograms of synthetic composites (results not shown) presented the signal corresponding to pyrite (sample C2Py) and chalcocite (sample C2Ct). 3.2 Copper and iron extraction Figure 2A shows an important difference between the reactors of C1, with and without Ag(I), results of reference (12), and the reactors of C2, with and without Ag(I). These last ones practically did not differ between each other. It is necessary to mention that the concentrate C1 contained secondary copper sulfides, which are less recalcitrant to the acid dissolution, present a faster dissolution wich was on a greater copper extraction: in 22 days 45% without Ag(I) and 62% with Ag(I). For C2 a copper extraction of only 16% in almost 80 days was reached in both cases. In figure 2B it can be observed that the iron extraction in systems with C1 and C2 was very similar with the exception of the C1 reactor with Ag(I), where a greater extraction in a smaller residence time than with C2 was reached: 25% of extraction in 30 days. The recalcitrance of C2 with respect to C1 was greater. In the reactors with C1, figure 2A, an important amount of copper came from secondary sulfides, because in the dissolution of the chalcopyrite equal amounts of both 205
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copper and iron are practically dissolved. Except for the C1 with Ag(I), all of them had a very similar behavior for iron dissolution.
Figures 2A and 2B. Cu and Fe dissolution respectively. ▲C1 without Ag(I), • C1 with Ag(I), ■ C2 without Ag(I) and ♦ C2 with Ag(I) Under the testing conditions, a slower dissolution of C2 was obtained, with a residence time of 80 days. Because of the poor results obtained in the bioleaching tests of C2, it was decided to doped systems of this concentrate with pure minerals of pyrite and chalcocite in order to corroborate the influence of these sulfides on the copper extraction using Ag(I). Figures 3A and 3B show the copper and iron extraction in the bioleaching systems under different testing conditions: pyrite and chalcocite doped C2 systems in presence of Ag(I). In figure 3A it is observed that, at the beginning of the experiments, the reactors with greater copper dissolution were the chalcocite-C2 doped systems, due to the acid attack given by: CuS + 2H+ → Cu2+ + H2S
(1)
Figures 3A and 3B. Cu and Fe extraction respectively. ♦C2 + Ag(I), ■ Pyrite-C2 doped + Ag(I) and ▲Chalcocite-C2 doped + Ag(I) The massive dissolution of the chalcopyrite initiates after 70 days, obtaining a substantial increase in the copper extraction in the pyrite-C2 system reaching around 50% in 84 days. In all systems a greater copper dissolution is observed around day 70, agreeing this with the dissolution of iron, attributed, therefore, to the attack of the chalcopyrite. The higher copper recovery is obtained in the pyrite-C2 doped system, 50% in 84 days; surpassing by almost 20% that of the system also doped but with chalcocite for the same residence time. 206
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The iron extraction curves, figure 3B, present a slightly different behavior from the curves of copper extraction, being at the end of the tests the reactors pyrite-C2 doped and C2 those reached the greater extraction. In the chalcocite-C2 doped systems and in the systems containing only C2, it can be seen a slight increase in the iron dissolution. A possible reason for this behavior is the acid attack only on chalcocite, and therefore, the attack to the crystalline structure of the chalcopyrite was almost negligible. Figures 4A and 4B show the copper and iron extraction in the bioleaching pyrite-C2 and chalcocite-C2 doped systems in absence of Ag(I).
Figures 4A and 4B. Cu and Fe extraction respectively. ♦C2, ■ Pyrite-C2 doped and ▲Chalcocite-C2 doped It can be seen, in figure 4A, the acid attack to chalcocite crystalline structure occurred based on the high copper extraction in that chalcocite doped system, almost the same that with the pyrite doped system, 31% in 84 days. This result verifies the fact that the Cu in solution in these systems came mainly from the secondary sulfides that were added, and in the case of the system containing only C2 it resulted from the slight attack to the chalcopyrite structure. Results of MEB (not shown) demonstrated a preferential attack on secondary copper sulfides greater than on the chalcopyrite particle surface. In the case of the pyrite doped system without Ag(I) a slight reduction in the time needed to initiate the massive iron dissolution, 5 days with respect to others systems was observed. The same behavior was observed in the copper extraction curve of C2, figure 4A and 4B. This agrees with observations made by other authors (6,10,11) in the sense that, in a bioleaching system where different sulfide minerals are present, their respective rest potentials (Erep) will create galvanic interactions that will influence the selective dissolution of the most active minerals. 3.3 pH and redox potential, ORP In figures 5A, 5B, 6A and 6B the results corresponding to representative curves of redox potential, ORP, and pH evolution for pyrite-C2 and chalcocite-C2 doped systems with and without Ag(I), respectively, are showed. Previous work have showed the effect of the presence of Ag(I) on the redox potential behavior on bioleaching systems (12). In figure 5A it is observed that the ORP evolution was affected by the presence of Ag(I) in the catalyzed systems.
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The way in which Ag(I) affects the chalcopyrite dissolution is depicted by the following electrochemical reaction between Ag(I) and chalcopyrite: CuFeS2 + 4Ag+ → 2Ag2S↓ + Cu2+ + Fe2+ (2) Later, the silver sulfide generated in (2) reacts with the ferric ion, Fe(III), to form: (3) Ag2S(sup) + 2Fe3+ → 2Ag+ + 2Fe2+ + Sº This reaction consumes Fe(III) ion driving a decrease of the ORP value corresponding to Nernst equation on the Fe(III)/Fe(II) ratio. In the representative curves of catalyzed doped systems pH, figure 5B, a very similar behavior in all the cases was observed, except on chalcocite doped systems, where a slight increase of pH was recorded during the first 5 days due to the H+ consumption by the acid attack on chalcocite (reaction 1).
Figures 5A and 5B. Redox potential and pH respectively. ♦C2 + Ag(I), ■ Pyrite-C2 doped + Ag(I), ▲Chalcocite-C2 doped + Ag(I) From curves 6A it can be appreciated that the system that reached the highest value of ORP, and also with greater iron dissolution, figure 4B, was the pyrite-C2 doped system. In the other cases the values of redox potential are rather similar and only in those systems in which chalcocite was added the values of ORP were slightly low. In the final stage of the experiments, figure 6B, the reactors showed a slight increase in the pH value due to the ability of protons to attack the chalcopyrite (13) and due to a decrease in bacterial activity.
Figures 6A and 6B. Redox potential and pH of pyrite-C2 and chalcocite-C2 doped bioleaching systems in absence of Ag(I)
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4.
CONCLUSIONS •
The initial composition of sulfide minerals affected the kinetics of bioleaching, which was influenced by the galvanic interactions generated when putting in contact minerals of different rest potentials.
•
The galvanic interactions follow the thermodynamic prediction of galvanic series of sulfides.
•
The chalcocite presence did not improve the copper extraction from chalcopyrite.
•
The presence of pyrite improved the chalcopyrite bioleaching.
•
The catalytic effect of the Ag(I) was affected by the presence (or absence) of pyrite in the bioleaching systems.
•
The highest copper extraction was reached in the pyrite-C2 doped system in the presence of Ag(I): 50% in 84 days.
ACKNOWLEDGEMENTS One of the authors, A. L. J., wishes to acknowledge the support of CONACyT and DGEP-UNAM for a Ph. D., scholarship. Both authors gratefully acknowledge DGAPAUNAM for the economical support through Project IN210000. REFERENCES 1. R. S. McMillan, D. J MacKinnon, and J. E. Dutrizac. J. of App. Electrochem 12 (1982) 743-757. 2. D. F. A. Koch. J’OM Bockris and B. E Conway, Modern Aspects in Electrochemistry. Plenum Press. 10 (1971) 211-237. 3. S. G. Salamy and J. C. Nixon, 2nd International Congress on Surface Activity 3 (1957) 369. 4. E. Gómez, A. Balleter, M. L. Blázquez, F. González, Hydrometallurgy 51 (1999) 3746. 5. T. A. Fowler, F. K. Crundwell. In Biohydrometallurgy and the Environmental Toward the Mining of the 21st Century, IBS’99 R. Amils and A. Ballester, (eds), Part A, (1999) 273. 6. K. A. Natajaran, Metal. Trans. 23B (1992) 5-11. 7. P. B. Munoz, J. D. Miller, M. E. Wadsworth. Metal. Trans. B. 10 (1979) 149-158. 8. J. D. Miller, H. Q. Portillo in 13th International Mineral Processing Congress J. Laskowwski (ed), Warsaw, (1979) 851-894. 9. M. L. Blázquez, A. Álvarez, A. Ballester, F. González and J. A. Muñoz. In Biohydrometallurgy and the Environmental Toward the Mining of the 21st Century, IBS’99 R. Amils and A. Ballester, (eds), Part A, (1999) 137. 10. Y. A. Attia and El-Zeky. Inter. J. of Min. Proc. 30 (1990) 99-111. 11. D. W. Dew, C. Van Buuren, K. McEwan and C. Bowker. In Biohydrometallurgy and the Environmental Toward the Mining of the 21st Century, IBS’99 R. Amils and A. Ballester, (eds), Part A, (1999) 229. 12. A. López-Juárez, R. E. Rivera-Santillán, A. Ballester Pérez, M. L. Blázquez Izquierdo and J. A. Muñoz Sánchez. In Biohydrometallurgy: Fundamentals, Technology and Sustainable Development. V. S. T. Ciminelli and O. Garcia Jr. (eds), Part B, (2001) 661. 13. A. Schippers and W. Sand. App. Environ. Microbiol. 65 (1999) 319-321. 209
15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Comparison of air-lift and stirred tank batch and semi continuous bioleaching of polymetallic bulk concentrate D. R. Tiprea, S. B. Vorab and S. R. Davea* a
Department of Microbiology, School of Sciences, Gujarat University, Ahmedabad 380 009, Gujarat, India b Gujarat Mineral Development Corporation, Khanij Bhavan, 132 ft. Ring Road, Ahmedabad 380 052, Gujarat, India
Abstract Metal bioextraction from GMDC polymetallic bulk concentrate was studied in stirred tank and air-lift reactors with a working volume of 5 dm3 and 20% (w/v) pulp density. The inoculum used in these studies was a consortium mainly consisting of Acidithiobacillus ferrooxidans, Leptospirillum ferrooxidans and Acidithiobacillus thiooxidans. Under the optimum conditions air-lift reactor gave 41, 75 and 65% of copper and zinc extraction and galena oxidation as compared to the STR giving 76, 70 and 80% respectively in 5 to 9 d of incubation time. During the process, 590 and 547 mV redox potential, 0.05 and 0.07 g/l of soluble ferrous and pH of 1.55 and 1.78 were recorded in air-lift reactor and STR respectively. When the total 20% (w/v) pulp density was added in four equal fractions in STR study, instead of one lot addition, the metal extraction rate increased by 1.22 and 1.1 fold for copper and zinc respectively. Use of semi-continuous fed-batch STR process further enhanced the copper and zinc extraction rate by 1.6 and 1.54 fold as compared to batch STR process. The highest metal extraction of 92 and 87% of copper and zinc respectively was achieved in semi continuous process. Acid consumption was 75% less in fed-batch semi continuous process as compared to batch process. Optimisation of the process and use of developed inoculum reduced the bioleaching time from 30 d to as low as 5 and 7 d for copper and zinc respectively with overall increased percent metal extraction. The reduction in contact time could make the polymetallic bulk concentrate bioleaching process economically viable. Keywords: polymetallic concentrate, semi continuous, air-lift, stirred tank 1.
INTRODUCTION Bioleaching has generated intense research activity in late nineties which, resulted in important findings in the field, such as the development of economics and engineering factors of biometal extraction [1]. Bioextraction processes are now applied on commercial scale for the extraction of copper, cobalt, gold and nickel from refractory ores and concentrates [2, 3]. Microbial leaching has been used at laboratory scale for the base metal sulphides of Co, Ga, Mo, Ni, Pb and Zn. Biohydrometallurgical processes are developed * Corresponding author: S.R. Dave, E-mail:
[email protected]
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for extraction of nickel and cobalt from their sulphidic concentrates using mesophilic iron and sulphur oxidising bacteria in stirred tank reactor [2]. Several factors such as oxygen requirement, nutrient availability, agitation speed and inoculum size plays an important role in metal extraction. However, the operating solid concentration constitutes one of the most critical parameter of a bioleaching process in terms of size of the equipment. At higher operating pulp concentration the above mentioned factors could limit the bioleaching efficiency [4-6]. The previous laboratory pilot scale studies were mainly carried out only up to 10% solids with about 50 to 80% metal extractions in 15 to 20 days [1, 7, 8]. The indirect two-stage bioleaching process from complex sulphidic Cu-Pb-Zn concentrate is developed in which, metal extraction is done at 70°C temperature [9]. But, the direct bioextraction process for such complex polymetallic bulk concentrate is poorly developed. India has well scattered small reserves of polymetallic ores and Ambamata Mine situated in Gujarat is one of them. Biomineral processing holds great potential for such reserves in India. In India, inspite of cheap labour, liberalised market, vast consumer base and strong foundation of hydrometallurgy, there exists a wide gap between the existing potential and the potentials to be exploited for economic metal growth [10, 11]. In biohydrometallurgical processes the stirred tank and air-lift reactors are widely used at laboratory scale but pulp density studied usually is upto 12.5% [12]. In this context, in the presented work air-lift and stirred tank batch and semi continuous bioleaching of polymetallic bulk concentrate was carried out after optimisation at shake flask leaching scale. Comparative bioleaching profiles of both these reactors at laboratory scale with 20% pulp density is discussed. 2.
MATERIALS AND METHODS
2.1 Polymetallic concentrate Polymetallic concentrate was procured from Gujarat Mineral Development Corporation (GMDC), Ambamata Multimetal Project, Gujarat, India. Major constituents of the concentrate were sphalerite, chalcopyrite, galena and pyrite. The concentration of copper, lead, zinc, iron and sulphur in the concentrate were 2.5, 13, 30, 9 and 27% respectively. Detailed composition is given elsewhere [13]. The bulk concentrate received was of mixed particle size ranging between +150-325 # B.S.S. and was used as received. In all the experiments 20% (w/v) pulp density was used. 2.2 Inoculum The inoculum used for extraction of metals was developed from shake flask leaching experiment in the form of leachate. Leachate used in the experiment mainly consisted of a consortium of acidophilic chemolithotrophic auto- and heterotrophic iron and sulphur oxidisers [14]. All the experiments were carried out with 20% (v/v) leachate as inoculum and M-2 modified medium [15] prepared in tap water. 2.3 Reactors Bioextraction studies were compared at laboratory scale with two reactors. First, an Air Lift Glass Reactor (ALGR) which was fabricated in the laboratory consisting 5 dm3 working volume. Detailed design set up and configuration is given elsewhere [12]. The other bioreactor tested was a laboratory scale Stirred Tank Reactor (STR) with axial turbine type impellers which was designed in our laboratory and fabricated by Texfab 212
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Manufacturers, India. The reactor consisted 5 vessels (3 in cascade and 2 in series) each 5 dm3 capacity with working volume of 3 dm3. The rate of agitation, aeration and CO2 supply was 300 rev/min, 0.5 l/min/v and 0.2% (v/v) of compressed air respectively. The detail configuration is quoted elsewhere [10, 11, 15]. Temperature during the investigation was 30 ± 5°C, i.e. ambient room temperature. 2.4 Bioleaching study In both the reactors typical batch leaching trial was operated until the logarithmic growth and steady leaching conditions were well established. In STR the pulp addition mode was tested with the four equal parts i.e. instead of addition of pulp 20% (w/v) at the 0th h, it was added in four fractions of 5% (w/v) each at 24 h of interval. The semi continuous bioleaching study was carried out in STR with 7 d of residence time. The detail design is given in the previous article [5]. 2.5 Analysis Supernatant of the leaching system was analysed periodically. pH and redox potential (Eh) was determined using Systronics Digital µ pH system 361 with platinum SCE couple electrode. Soluble ferrous iron was estimated by potassium dichromate titrimetric method. Copper, zinc and lead were estimated by spectrophotometric (Systronics UV–Vis spectrophotometer 119), polarographic (ELICO Polarograph Model CL-25 D) and titrimetric (tannic acid as external indicator) methods respectively from the leached solutions [16]. The analysis was confirmed by Atomic Absorption Spectrophotometer (Varian AA-175 model). 3.
RESULTS AND DISCUSSION After the amenability and bioleaching optimisation study of GMDC polymetallic bulk concentrate at shaken flask level [14, 17], the batch and semi continuous fed batch processes were performed with 20% (w/v) pulp density in a laboratory scale reactor. The performances of two reactors were compared for batch leaching with airlift glass reactor and stirred tank reactor, which were challenged with the consortium developed at shake flask as 20% (v/v) leachate. As can be seen from the data presented in Table 1, in ALGR highest copper extraction achieved was 40.1% while, lead and zinc were 65 and 75% respectively at 9 d of contact time in batch leaching. Throughout the study, the extraction of lead represent oxidation of galena to lead sulphate, which remain in insoluble form with the pulp. The system was stabilised after 6 d of residence time which can be noted from the pH, redox potential and soluble ferrous being 1.55, 590 mV and 0.05 g/l respectively at the end of 9 d. The total acid consumption recorded was 19.3 g sulphuric acid/Kg concentrate. This acid requirement was to bring pH back to 2.0 could be due to the acid consuming material present in the pulp and the chemical oxidation of metal sulphide [12].
Thereafter, the gradual decrease in the pH indicates the enhancement in biooxidation of sulphide content of the concentrate.
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Table 1. Batch bioleaching profile of Air Lift Glass Reactor Bioleaching time (days) Leaching profile
Before inoc. 2.0 308 0.62
After inoc. 2.35 320 0.97
pH Redox potential (mV) Soluble Fe2+ (g/l) Acid consumption 2.9 (g/Kg concentrate) Copper (%) 2.2 4.8 Zinc (%) 12.91 15.92 Lead (%) 4.5 4.9 n.d.: not determined; inoc.: inoculation
2
4
6
8
9
2.7 330 1.06
2.6 370 0.09
1.6 520 0.06
1.55 580 0.05
1.55 590 0.05
6.4
10.0
-
-
-
11.22 24.38 n.d.
24.36 50.35 n.d.
41.0 64.98 n.d.
31.6 71.48 n.d.
40.1 75.0 65.0
When the leaching profile of STR and ALGR batch studies were compared (Table 2), STR performance was proved better than that of ALGR. The maximum copper, lead and zinc extraction recorded in STR were 76, 70 and 80% in 6, 9 and 9 d which comes out to be 2.84, 1.06 and 1.07 folds higher as compared to ALGR. The acid consumption was 10.7 g acid/Kg concentrate which was 1.8 times lower than the ALGR. This indicates highly stabilised system, which was due to the positive effect of aeration and agitation system adopted during the process [18]. The observed high chalcopyrite leaching indicates dominance of biological leaching over chemical leaching as chalcopyrite was very difficult to leach chemically. Table 2. Bioleaching profile of batch STR process Bioleaching time (days) Leaching profile pH Redox potential (mV) Soluble Fe2+ (g/l) Acid consumption (g/Kg concentrate) Copper (%) Zinc (%) Lead (%)
Before inoc. 2.0 326 0.66
After inoc. 2.52 344 1.22
2
4
6
8
9
2.5 354 1.53
2.29 382 0.99
1.84 429 0.73
1.78 528 0.09
1.78 547 0.07
-
5.73
4.97
-
-
-
-
3.8 8.4 5.1
8.4 20.68 5.5
34.36 39.36 n.d.
57.14 48.68 n.d.
76.0 62.93 n.d.
66.8 72.0 n.d.
65.96 80.0 70.0
Looking to the promising leaching result of STR, further experiment was performed with fractional pulp addition and results are shown in Table 3. As can be seen from the data 92.7% copper extraction was achieved in 140 h of contact time followed by 83.62 and 83% zinc and lead extraction respectively in 215 h of reactor run. The high metal extraction during this fractional addition of pulp could be due to the less shear effect generated and better gas exchange rate owing to low pulp density at any particular time throughout the process. The high biological activity throughout the process was responsible for constant neutralisation of the alkaline gangue present in the fraction pulp added at that time, which resulted in 1.96 times less acid consumption compare to the STR batch process. Even when 30% (w/v) pulp was added in 10+10+10% (w/v) fractions, extraction rates of 0.84 and 6.36 g/l/d for copper and zinc respectively (data not shown) were achieved. This suggests that with the developed inoculum it is possible to get the 214
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considerable metal extraction even at such a high pulp density, which was almost equivalent to that achieved at 15% pulp density with fractional pulp addition [5]. Table 3. Bioleaching profiles during fractional addition of pulp in STR Bioleaching time (days) Leaching profile
Before inoc. 1.95 322 0.42
After inoc. 2.24 332 0.50
2
4
6
8
9
2.16 373 0.75
2.25 412 0.94
2.12 430 1.11
1.93 446 0.80
1.82 525 0.09
-
1.82
1.82
1.82
-
-
-
Copper (%)
6.2
18.4
56.80
64.57
92.7 (140h)
71.34
Zinc (%)
12.8
25.15
58.0
63.54
69.08
74.34
Lead (%)
5.3
6.0
n.d.
n.d.
n.d.
n.d.
pH Redox potential (mV) Soluble Fe2+ (g/l) Acid consumption (g/Kg concentrate)
83.62 (215h) 88.0 (215h) 83.0 (215h)
The better extraction results achieved during fractional addition of the pulp in STR study opened the way for semi continuous fed batch process. The semi continuous metal extraction was started with a stabilised leaching system which, reached to 75% metal extraction, thus even at initiation time the concentration of the extracted metals present in leachate was 33 ± 2.0%. At the time of calculation of the percent metal extraction this amount was substracted, and the results presented in Table 4 are of net average extraction of 15 cycles. The highest metal extractions of 92, 87 and 80% were achieved with copper, zinc and lead respectively. The metal extraction time was reduced by 35 h for copper while for lead and zinc it was reduced by 65 h as compared to the bioleaching with fractional pulp addition STR process. The bioleaching time is the major factor affecting the economy of the process. The other cost factor in the process was the external acid addition or acid consumption, which was 3.99 fold less or 75% reduction compare to STR batch leaching process. Table 4. Polymetallic bioleaching profile in semi continuous STR process based on average of 15 cycles Bioleaching time (days) Leaching profile
Before inoc.
After inoc.
2
4
6
7
pH Redox potential (mV) Soluble Fe2+ (g/l) Acid consumption (g/Kg concentrate)
2.06 321 0.79
2.23 338 1.06
2.15 397 1.23
2.06 448 0.94
1.96 502 0.08
1.80 521 0.07
-
1.73
0.95
-
-
-
Copper (%)
31.59
37.4
71.86
92.0 (105h)
73.54
65.98
Zinc (%)
32.63
45.68
57.07
63.08
73.81
Lead (%)
35.12
36.4
n.d.
n.d.
n.d.
87.0 (150h) 80.0 (150h) 215
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The comparative metal extraction rates of all the four studied processes are depicted in Table 5. As can be seen from the data, a higher metal extraction rate was obtained in the STR semi continuous process as compared to any of the processes studied. This was 43.81 and 348.0 mg/l/h, which comes out to be 1.05 and 8.35 g/l/d copper and zinc solubilisation respectively. The total metal extracted from the polymetallic concentrate was 4.6 g/l copper, 52.5 g/l zinc and galena oxidised equivalent to 21.12 g/l or 105.60 g/Kg lead in leached residue in the form of lead sulphate. Table 5. Comparative soluble metals in leachate, galena oxidised and the overall metal extraction rate Metals Zinc
Copper
Process g/l
mg/l/h
g/l
mg/l/h
ALGR 2.05 14.64 45.0 209.30 STR (batch) 3.80 27.14 48.0 223.36 STR (fraction) 4.64 33.14 52.8 245.58 STR (semi 4.60 43.81 52.5 348.00 continuous) *: equivalent lead due to oxidation of galena to lead sulphate
g/l* 17.16 18.48 21.91 21.12
Lead g/Kg mg/l/h* concentrate 85.80 79.81 92.40 85.95 109.55 101.92 105.60 140.80
The selection and development of the efficient bioleaching consortium and modification of pulp addition resulted in reduction of leaching time from 30 d at shake flask level with wild type consortium [14] to as low as 5 to 6 d. The observed high metal extraction as well as high rate proved suitability of the semi continuous process over the other three processes studied. Moreover, the reduction in retention time as well as external acid consumption leads to economization of the process. On the basis of obtained data the bioextraction process was successfully scaled-up to 600 dm3 STR level in our laboratory and was efficiently operated for 17 cycles with more than 80% average metal extractions which is detailed elsewhere [13]. The noteworthy developments of this investigation are the room temperature operations, high pulp density as well as the higher zinc extraction rate obtained as compared to the reported values in literature [1, 7, 8, 19-21]. The zinc extraction data obtained in this investigation can be compared with those of Steemson et.al [21] who have reported 93.8% zinc extraction in the overall residence time of 3.7 days at 40-45°C temperature and 6-8% solids. The highest overall rate of zinc extraction achieved in this work is 348 mg/l/h. When the amount of zinc extraction is compared per day on one litre pulp volume basis, the extraction works out to be 8.35 g/l/d, which is comparable with the calculated value of 8.75 g/l/d from the Steemson et.al report [21]. What is interesting is, that the presented work was performed at ambient temperature compared to 40-45°C. Moreover, the pulp density is 3 times higher in this investigation compared to the above reference. When adopted on commercial basis the process may be more cost effective in terms of capital as well as operating costs. 4.
CONCLUSION To summarise, the achieved equivalent zinc extraction rate, higher than that reported by others, even at lower temperatures and higher pulp density using indigenously developed consortium, indicates comparatively higher cost effectiveness of the bioleaching process developed for GMDC polymetallic bulk concentrate. Interestingly, the 216
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investigation was carried out with a Cu-Pb-Zn system, which allowed as much as 90% copper extraction and 80% oxidation of lead sulphide simultaneously. ACKNOWLEDGMENTS We are thankful to Gujarat Mineral Development Corporation for the project grant and research fellowship to D. R. Tipre. REFERENCES 1. Y. Rodriguez, A. Ballester, M. L. Blazquez, F. Gonzalez and J. A. Munoz, In: Biohydrometallurgy: Fundamentals, Technology and Sustainable Development, Part A, V. S. T. Ciminelli and O. Garcia Jr. (eds.), Proc. Intl. Biohydrometallurgy Symp., Elsevier, Amsterdam, (2001) 125. 2. J. A. Brierley and C. L. Brierley, In: Biohydrometallurgy and the Environment toward the Mining of the 21st Century, Part A, R. Amils and A. Ballester (eds.), Proc. Intl. Biohydrometallurgy Symp., Elsevier, Amsterdam, (1999) 81. 3. C. L. Brierley, In: Biomining: Theory, Microbes and Industrial Processes, D. E. Rawlings (ed.), Springer-Verlag, New York, (1997) 3. 4. A. D. Bailey and G. S. Hansford, Biotechnol. Bioengg., 42(10) (1993) 1164. 5. S. R. Dave, D. R. Tipre and S. B. Vora, In: Biohydrometallurgy: Fundamentals, Technology and Sustainable Development, Part A, V. S. T. Ciminelli and O. Garcia Jr. (eds.), Proc. Intl. Biohydrometallurgy Symp., Elsevier, Amsterdam, (2001) 561. 6. D. R. Tipre, S. B. Vora and S. R. Dave, In: Mineral Biotechnology, L. B. Sukla and V. N. Misra (eds.), Proc. Nat. Seminar, Allied Publishers Pvt. Ltd., New Delhi, (2002) 81. 7. R. E. Rivera, A. Ballester, M. L. Blazquez and F. Gonzalez, In: Biohydrometallurgy and the Environment toward the Mining of the 21st Century, Part A, R. Amils and A. Ballester (eds.), Proc. Intl. Biohydrometallurgy Symp., Elsevier, Amsterdam, (1999) 149. 8. A. Chin, In: Biotechnology comes of age, BIOMINE'97, Proc. Intl. Biohydrometallurgy Symp., Australian Mineral Foundation, Australia, (1997) M1.2.1. 9. R. Romero, I. Palencia and F. Carranza, Hydrometallurgy, 49 (1998) 75. 10. A. R. Tipre, Ph. D. Thesis, Gujarat University, Ahmedabad, Gujarat, India, 1999. 11. S. R. Dave, D. R. Tipre and S. B. Vora, In: Proc. Intl. Seminar on Mineral Processing Technology, S. Subramanian (ed.), Bangalore, (2002) (in press). 12. A. R. Tipre, S. B. Vora and S. R. Dave, Ind. J. Microbiol., 41 (2001) 173. 13. S. R. Dave, D. R. Tipre and S. B. Vora, In: Application of Chemical Engineering for Utilisation of Natural Resources, G. K. Roy, C. R. Mishra and K. Sarveswara Rao (eds.), New Age International Publishes, New Delhi, (2001) 169. 14. A. R. Tipre, S. B. Vora and S. R. Dave, J. Scientific Ind. Res., 57 (1998) 805. 15. D. R. Tipre, S. B. Vora and S. R. Dave, In: Biohydrometallurgy and the Environment toward the Mining of the 21st Century, Part A, R. Amils and A. Ballester (eds.), Proc. Intl. Biohydrometallurgy Symp., Elsevier, Amsterdam, (1999) 219. 16. A. I.Vogel, A Textbook of Quantitative Inorganic Analysis, 3rd ed., ELBS and Longman, London, 1962. 17. A. G. Menon and S. R. Dave, In: Biohydrometallurgical Technologies, Vol. 1, A. E. Torma, J. E. Wey and V. L. Lakshamanan (eds.), Proc. Intl. Biohydrometallurgy Symp., TMS, Wyoming, (1993) 137. 18. D. R. Tipre, S. B. Vora and S. R. Dave, In: Biotechnology of Microbes and Sustainable Utilisation, R. C. Rajak (ed.), Scientific Publishers (India), Jodhpur, (2002) 263. 217
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19. T. A. Fowler and F. K. Crundwell, In: Biohydrometallurgy and the Environment toward the Mining of the 21st Century, Part A, R. Amils and A. Ballester (eds.), Proc. Intl. Biohydrometallurgy Symp., Elsevier, Amsterdam, (1999) 273. 20. P. Chawakitchareon, B. Buddharuksa and S. Pradit, In: Biotechnology comes of age, BIOMINE'97, Proc. Intl. Biohydrometallurgy Symp., Australian Mineral Foundation, Australia, (1997) M1.3.1. 21. M. Steemson, F. Wong, B. Goebel, In: Biotechnology comes of age, BIOMINE'97, Proc. Intl. Biohydrometallurgy Symp., Australian Mineral Foundation, Australia, (1997) M1.4.1.
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Effect of pH and temperature on the biooxidation of a refractory gold concentrate by Sulfolobus metallicus I. Ñancucheo, J. C. Gentina* and F. Acevedo School of Biochemical Engineering, Catholic University of Valparaiso, Av. Brasil 2147, Valparaíso, Chile Abstract The biooxidation of refractory gold concentrates using thermophilic microorganisms is considered an interesting alternative to current processes. Many operation conditions influence the process, pH and temperature being two important parameters among others. The objective of this work was to evaluate the influence of these two variables on the biooxidation of an enargite-pyrite gold concentrate containing 40 g Au/ton, using the thermophilic archaeon Sulfolobus metallicus. The experiments were run in shake flasks with 1% pulp density and particle size under 200 mesh. The pH was kept constant at 1.0, 1.5 and 2.0 by daily addition of 0.5 M NaOH. One experiment was run at non-controlled condition with initial pH at 2.5. Every pH condition was tested at 60, 65, 70 and 75ºC. The extent of biooxidation was measured through the iron solubilisation. The best run at constant pH condition and at all different temperatures was 1.5, obtaining the highest percentage of iron extraction at 65ºC which amounted 75%. Iron extractions at pH 1.0 were between 10% and 24% of those obtained at pH 1.5, while at pH 2.0 they were in the range of 75% to 95%. All experiments run at 75ºC showed almost no iron solubilisation. However, the non-controlled pH experiment with initial pH at 2.5 was shown to be a more suitable condition for biooxidation. The highest iron extraction was 83% at 65ºC which is 10% higher than the one obtained at pH controlled at 1.5. It is concluded that, under the studied conditions, the best pH and temperature to perform the biooxidation are 65ºC and a non-controlled pH policy with initial pH of 2.5. Keywords: enargite, biooxidation, thermophiles, bioleaching, archaea 1.
INTRODUCTION Biooxidation has become an attractive alternative as a pretreatment of refractory gold concentrates in order to facilitate gold extraction by cyanidation [1-4]. Biooxidation competes against roasting and pressure oxidation presenting several advantages such as low capital demand, low energy input, mild operation conditions and low environmental impact. Nowadays there are operating no less than ten large scale biooxidation plants located in South Africa, Brazil, Australia, Ghana, Peru and USA [5, 6].
*
[email protected]
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Most of these commercial gold processing plants use stirred tank bioreactors for the biooxidation of refractory gold concentrates, although a few of them use heap leaching to pre-treat low-grade ores and tailings [7]. The advantage of using stirred tank bioreactors is the control that can be exerted on important environmental variables like pH and temperature [6, 8, 9]. Microorganisms that are involved in these operations occurs naturally at mineral sites and are mostly of the genus Acidithiobacillus preferentially Acidithiobacillus ferrooxidans, Acidithiobacillus thiooxidans and Acidithiobacillus caldus together with Leptospirillum ferrooxidans [11-13]. All of them are acidophiles and mesophiles, being used at operating pH between 1.2 to 2.2 and temperature between 30 and 40ºC [14]. Several attempts have been made in order to evaluate the performance of thermophilic bacteria and more recently hyperthermophilic archaea which grow at temperatures higher than 60ºC. Some examples of these archaea are Sulfolobus metallicus, Sulfolobus acidocaldarius, Sulfolobus solfataricus, Acidianus brierleyi and Acidianus infernus [15-17]. Although these archaea have natural environments different from mineral sites, they have adapted very well to oxidize different mineral species like pyrite, arsenopyrite, enargite, chalcopyrite, chalcocite and covellite [18-21] The bioleaching mechanisms used by archaea are less understood than those of Acidithiobacillus and much more basic information is needed. This work presents results that show the influence of pH and temperature on the solubilisation of a refractory gold concentrate with high content of enargite, a recalcitrant species present in several gold minerals in Chile using the archaeon Sulfolobus metallicus. 2.
MATERIALS AND METHODS
2.1 Experimental conditions A strain of Sulfolobus metallicus, kindly supplied by Dr. Antonio Ballester (Faculty of Chemical Science, Universidad Complutense, Madrid), was used throughout this study. The cells were cultured in a medium containing 0.4 g/L of (NH4)2SO4, 0.5 g/L of MgSO4.7H2O, 0.2 g/L of KH2PO4, 0.1 g/L of KCl and 1% (w/v) of gold concentrate as energy source. The mineralogical composition of the gold concentrate (El Indio Mining Company, La Serena, Chile) was: 40.7% enargite, 42.8% pyrite, 3.9% chalcopyrite, 0.8% chalcosine and 0.3% covellite. Its elemental composition was: 42 g Au/ton, 21.1% Cu, 22.6% Fe, 37.8% S and 7.7% As. Particle size was lower than 200 mesh. Experiments were carried on in 500 mL erlenmeyer flasks with 100 mL of medium, incubated in a rotary shaker at 200 rpm. The different culture pH were kept constant at 1.0, 1.5 and 2.0 by daily addition of 0.5 M NaOH. One experiment was conducted under non-controlled pH policy with an initial pH of 2.5. Each pH condition was tested at 60, 65, 70 and 75ºC. The evaporated water was quantitatively replaced on a daily basis adding distilled water. The inoculum was obtained from a fully adapted culture grown on a medium of the same composition as described above. 2.2 Analytical procedures Total soluble iron, after reduction of ferric ion with hydroxylamine, was determined by the o-phenanthroline method [22]. Eh was measured with a Ag/AgCl probe.
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3.
RESULTS AND DISCUSSION Figures 1, 2 and 3 depict the total soluble iron concentration profiles at all pH and temperature conditions studied, with the exception of the experiments run at 75ºC, a temperature at which after eight days iron solubilisation was not observed. This temperature was too high as to allow this archaeon to grow, a result that is consistent with the culture temperature reported by different authors that use Sulfolobus species in biooxidation studies [23, 24]. Under the operating conditions used in this work, the biooxidation of the mineral stopped after eight to ten days. Also, it was observed that under the different tested conditions biooxidation started after three to four days of lag, a situation that is related to the difficulties that the cells face when they are transferred as inoculum to a fresh medium. Experiments run at pH 1.0 showed almost no iron extraction independently of the cultivation temperature. It is thought that the initial iron solubilisation should be due mostly to a chemical action more than an initial biological activity. As pyrite is the most abundant mineral component of the refractory gold concentrate, the iron obtained in solution as a consequence of biooxidation must come from pyrite and therefore constitutes a measurement of the extent of pyrite oxidation. In this respect, from the results it is clear that the highest iron extractions were obtained at a temperature of 65ºC. In the case of the biooxidation experiments conducted at controlled pH, the highest extraction was obtained at a pH of 1.5, amounting to 75% of the iron present initially in the refractory gold concentrate. Table 1 summarizes the percentages of Fe extraction obtained under the different operating pH and temperature considered in this work. The extraction values obtained at pH 1.0 are between one fifth to one tenth of those obtained at pH 1.5, which reveals the importance to keep biooxidation pH above 1.0 in order to avoid a marked reduction on the process yield.
Figure 1. Iron solubilisation kinetics during the biooxidation of a refractory gold concentrate at 60ºC and different pH (nc: non-controlled)
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Figure 2. Iron solubilisation kinetics during the biooxidation of a refractory gold concentrate at 65ºC and different pH (nc: non-controlled)
Figure 3. Iron solubilisation kinetics during the biooxidation of a refractory gold concentrate at 70ºC and different pH (nc: non-controlled) Table 1. Iron extraction by biooxidation of a 1% pulp density refractory gold concentrate at different pH conditions and temperature (%) Temperature (ºC) 60 65 70
Culture pH Non-controlled 64 83 69
1.0 14 7 14
1.5 58 75 67
2.0 44 55 64
In the case of experiments conducted under pH controlled condition and as a way to predict the optimal pH and temperature to carry out the biooxidation of the refractory gold concentrate using Sulfolobus metallicus, the experimental data was fitted to a second order polynomial function to generate a surface response. This equation represents the 222
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solubilised iron concentration [Fe] as a function of biooxidation pH and temperature The following equation was drawn using a statistical software fed with the experimental results tabulated according to a central composite rotatable experimental design:
[ Fe] = −19.2326 + 7.4686 ⋅ pH + 0.4244 ⋅ T − 3.1629 ⋅ pH 2 + 0.046 ⋅ pH ⋅ T − 0.0036 ⋅ T 2
(1)
2
The values of the coefficient of determination (R ) and the standard deviation are 0.975891 and 0.138646 respectively. Calculating the first partial derivatives of equation (1) the optimum culture pH and temperature were quantified as 1.68 and 69ºC, respectively. These operating conditions predict a soluble iron concentration of 1.72 g/L, which represents a 76% of iron extraction The optimal pH and temperature fit in the ranges reported in the literature for Sulfolobus metallicus [25, 26]. However, as can be seen in Table 1, this figure still is lower than the iron extraction obtained in the experiment run at 65ºC and under non-controlled pH condition. In this case the iron extraction was 83%, showing that the biooxidation of a 1% pulp density refractory gold concentrate by Sulfolobus metallicus was more efficient when conducted under non-controlled pH condition. Although there is not a precise explanation for this phenomenon, it has been noted that sometimes microbial activity is higher under a policy of initially adjusting the pH and letting it change freely during the time course of the oxidation. The pH varied moderately during the biooxidation as can be seen in Figure 4, leveling off at a lowest value of 1.7, very close to the one predicted by equation (1). This policy is thought to be worthwhile for the biooxidation of solid suspension of low pulp density as in this work, since the pH downfall is restricted by the total amount of iron that can be extracted. As pulp density goes up, the total iron solubilisation increases and consequently the pH drop should become more important, reaching values that would interfere with microbial activity and extent of the biooxidation process. Also in Figure 4 is observed a typical Eh behavior during biooxidation, starting from low values and increasing as solubilisation of different elements goes up, especially iron.
Figure 4. pH and Eh evolution during the biooxidation of a refractory gold concentrate at 65ºC and under non-controlled policy with initial pH of 2.5 From the data contained in Figures 1, 2 and 3, it is possible to calculate the global volumetric productivity of iron solubilisation at each experimental condition. According to Table 2, again this parameter is maximum for biooxidation conducted at 65ºC and pH 223
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controlled at 1.5. This time the global volumetric productivity is slightly higher at the pH controlled condition compared to non-controlled operation. This situation is clearly confirmed in Figure 2 where it is possible to infer that the rate of iron solubilisation at pH controlled at 1.5 is faster than at the non-controlled pH condition. Table 2. Global volumetric productivity of iron solubilisation of a batch biooxidation of a 1% pulp density refractory gold concentrate at different pH conditions and temperature (g/L·d) Temperature (ºC) 60 65 70
Culture pH Non-controlled 0.132 0.208 0.171
1.0 0.036 0.023 0.041
1.5 0.129 0.211 0.160
2.0 0.099 0.206 0.179
4.
CONCLUSIONS The results allowed the definition of the pH and temperature values that maximized the iron extraction from a refractory gold concentrate through biooxidation using Sulfolobus metallicus, a thermophilic archaeon. Under the experimental conditions studied in this work, the best temperature and pH to perform the biooxidation are 65ºC and a noncontrolled pH policy starting with initial pH of 2.5. On the other side, operating at a pH controlled condition, using surface response methodology it is shown that pH controlled at 1.68 and a temperature of 69ºC maximize the extent of iron solubilisation. On the other hand, the highest value of the global volumetric productivity of iron solubilisation is found in the pH controlled operation mode and correspond to a biooxidation conducted at pH 1.5 and 65ºC. ACKNOWLEDGEMENTS This work was supported by the National Commission of Science and Technology through the FONDECYT projects 1000284 and 1020768. REFERENCES 1. S. Hutchins, J. Brierley and C.L. Brierley, Mining Eng., April (1988). 2. C.L. Brierley, Australian Gold Conference, Kalgoorlie, Australia, 1992. 3. J.A. Brierley, R.Y. Wan, D.L. Hill and T.C. Logan, Biohydrometallurgical Processing (T. Vargas, C.A. Jerez, J.V. Wiertz and H. Toledo, eds.), Vol. I, p. 253, Universidad de Chile, Santiago, 1995. 4. A.W. Breed, C.J.N. Dempers and G.S. Hansford, J. South African Inst. Min. Metall. (2000) 389. 5. S. Gilbert, C. Bounds and R. Ice, CIM Bull., 81 (1988) 89. 6. F. Acevedo, Electronic J. Biotechnol., 3 (2000). Available from Internet: http://www.ejb.org/content/vol3/issue3/full/4/index.html. 7. C.L. Brierley, Biohydrometallurgy and the Environment toward the Mining of the 21st Century (R. Amils and A. Ballester, eds.), Part A, p. 91, Elsevier, Amsterdam, 1999. 8. A. Pinches, J.T. Chapman, W.A.M. te Riele and M. van Staden, Biohydrometallurgy (P.R. Norris and D.P. Kelly, eds.), p. 329, Science and Technology Letters, Surrey, 1988. 9. F. Acevedo and J.C. Gentina, Bioprocess Eng., 4 (1989) 223. 224
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10. D.W. Dew, Biohydrometallurgical Processing (T. Vargas, C.A. Jerez, J.V. Wiertz and H. Toledo, eds.), Vol. I, p. 239, Universidad de Chile, Santiago, 1995. 11. D.W. Dew, E.N. Lawson and J.L. Broadhurst, Biomining: Theory, Microbes and Industrial Processes (D.E. Rawlings, ed.), p. 45, Springer Verlag, Berlin, 1997. 12. D.E. Rawlings, H. Tributsch and G.S. Hansford, Microbiol-UK, 145 (1999) 5. 13. F.K. Crundwell, Biotechnol. Bioeng., 71 (2001) 255. 14. R.W. Lawrence and P.B. Marchant, Biohydrometallurgy (P.R. Norris and D.P. Kelly, eds.), p. 359, Science and Technology Letters, Surrey, 1988. 15. P. Norris and J.P. Owen, FEMS Microbiol. Rev., 11 (1993) 51. 16. D.W. Dew, C. van Buuren, K. McEwan and C. Bowker, Biohydrometallurgy and the Environment toward the Mining of the 21st Century (R. Amils and A. Ballester, eds.), Part A, p. 229, Elsevier, Amsterdam, 1999. 17. Y. Konishi, S. Asai and M. Tobushige, Biotechnol. Progr., 15 (1999) 681. 18. F. Kargi and M. Robinson, Biotechnol. Bioeng., 24 (1982) 2115. F. Torres, M.L. Blázquez, F. González, A. Ballester and J.L. Mier, Metall. Mater. Trans. B, 26B (1995) 455. 20. Y. Konishi, S. Yoshida and S. Asai, Biotechnol. Bioeng., 48 (1995) 592. 21. P. d’Hugues, S. Foucher, P. Gallé-Cavalloni and D. Morin, Int. J. Miner. Process., 66 (2002) 107. 22. L. Herrera, P. Ruiz, J.C. Aguillon and A. Fehrmann, Chem. Tech. and Biotechnol., 44 (1989) 171. 23. Y. Konishi, M. Tokushige and S. Asai, Biohydrometallurgy and the Environment toward the Mining of the 21st Century (R. Amils and A. Ballester, eds.), Part A, p. 367, Elsevier, Amsterdam, 1999. 24. M. Nemati, J. Lowenadler and S.T.L. Harrison, Appl. Microbiol. Biotechnol., 53 (2000) 173. 25. J.L. Mier, A. Ballester, F. González, M.L. Blázquez and E. Gómez, J. Chem. Tech. Biotechnol., 65 (1996) 272. 26. M. Gericke, A. Pinches and J.V. van Rooyen, Int. J. Miner. Process., 62 (2001) 243.
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Effect of the pulp density and particle size on the biooxidation rate of a pyritic gold concentrate by Sulfolobus metallicus P. Valencia, J.C. Gentina and F. Acevedo* School of Biochemical Engineering, Catholic University of Valparaiso, Av. Brasil 2147, Valparaíso, Chile Abstract It is recognized that increasing pulp densities and decreasing particle sizes have positive effects in the volumetric rate of biooxidation of refractory gold concentrates. It has also been noted that a variety of phenomena can limit this positive effect. The objective of this work was to determine the values of pulp density and particle size that maximize the volumetric rate of solubilisation of iron from a pyritic gold concentrate. The leaching was carried on in agitated flasks with the thermophilic archaeon Sulfolobus metallicus. The concentrate contained 66.7% pyrite, and the constant operation conditions were 220 rpm, 68ºC initial pH of 2.0. Pulp densities were 2.5, 5, 10 and 15% w/v and the size fractions were 150-106, 106-75, 75-38 and –38 µm. Total solubilised iron concentrations were in the range of 8 to 25 g/L. In the 2.5 and 5% pulp density runs, iron extractions were in the range of 80 to 100%. After 15 to 25 days of leaching the rate declined to almost zero in the runs with 2.5 and 5% pulp densities, while the same occurred in the other two runs after 20 to 45 days. A complete experimental design of 16 runs allowed the definition of response surfaces from which the optimal conditions that maximize the rate of iron solubilisation were determined. These optimal conditions are 7.8% pulp density and particle size of 35 µm. Keywords: optimal conditions, response surface, iron solubilisation, thermophilic archaeon 1.
INTRODUCTION The rate of biooxidation of refractory gold concentrates is influenced by several operational factors. It is recognized that increasing pulp densities and decreasing particle sizes have a positive effect in the volumetric rate of biooxidation, as both situations result in an increase in surface area. Nevertheless, it has also been noted that the interaction among these factors together with a variety of associated phenomena such as mechanical effects, metabolic stress and inhibitory concentrations of ferric ion, can limit this positive effect and even result in declining leaching rates [1].
*
[email protected]
227
Bioleaching Applications
The negative effects of high pulp densities and small particle sizes were early reported in bioleaching with mesophilic bacteria [2-3]. The detrimental effect of high pulp densities is likely to be larger in operations with archaea because of their weaker cell wall [4-5] that make them susceptible to mechanical damage and metabolic stress caused by the intense agitation needed for maintaining a homogeneous suspension [6-9]. On the other hand, decreasing particle size can reduce the leaching rate probably because of difficulties in cell attachment when the diameters of the particles and cells become of similar magnitude. It is also likely that the rate of collision between particles increases as particle size diminishes [9-10]. The objective of this work was to determine the optimal values of pulp density and particle size that maximize the volumetric rate of solubilisation of iron from a pyritic gold concentrate when using the thermophilic archaeon Sulfolobus metallicus in shake flasks. 2.
MATERIALS AND METHODS
2.1 Microorganism and culture conditions A strain of the thermophilic archaeon Sulfolobus metallicus, kindly supplied by Dr. Antonio Ballester from the Universidad Complutense, Madrid, was used throughout this work. The microorganism was maintained in Norris medium [11] (0.4 g/L (NH4)2SO4, 0.5 g/L MgSO4.7H2O, 0.2 g/L K2HPO4, 0.1 g/L KCl, H2SO4 to pH 2.0) in the presence of the gold concentrate. All experiments were run in 1-L shake flasks with 90 mL of the same culture medium with the concentrate at pulp densities of 2.5, 5, 10 and 15% w/v. The flasks were inoculated with 10 mL of a culture of adapted cells with total soluble iron concentration of 10 g/L, so in all runs initial concentration was 1 g Fe/L. Uninoculated flasks were run for each pulp density. Other culture conditions were 68ºC, initial pH of 2.0 and agitation in orbital incubator of 220 rpm. 2.2 Gold concentrate The concentrate contained 15 g gold/tonne and was supplied by Las Ventanas Copper Refinery, Las Ventanas, Chile. Its mineralogical and elemental composition is presented in Table 1. Table 1. Gold concentrate composition Component Mineralogical Pyrite (FeS2) Chalcopyrite (CuFeS2) Sphalerite (ZnS) Others Gangue Elemental S Cu Fe Zn Others 228
% w/w
67.55 8.98 0.56 0.71 22.20 39.47 3.32 34.43 0.38 22.40
Bioleaching Applications
Four fractions of the original concentrate were used in this work: 150-106, 106-75, 75-38 and –38 µm. 2.3 Analytical methods Ferrous ion was measured colorimetrically by the modified o-phenanthroline method [12]. Total soluble iron was determined by the same method after reduction of the ferric iron with hydroxylamine. Ferric ion was calculated as the difference between total and ferrous iron. Sulfate was determined volumetrically using the Cole-Parmer (Vernon Hills, IL) sulfate kit Nº 05542-23. 2.4 Experimental design A 24 factorial design was used. The factors were particle size and pulp density. The complete set of experiments is shown in Table 2. The coded variables were generated by defining the highest value of each variable as 1 and the lowest as –1. Table 2. Experimental design with the independent physical and coded variables Run
PS (µm)
PD (% w/v)
X1
X2
1
-38
2.5
-1
-1
2
75-38
2.5
-0.312
-1
3
106-75
2.5
0.132
-1
4
150
2.5
1
-1
5
-38
5
-1
-0.6
6
75-38
5
-0.312
-0.6
7
106-75
5
0.132
-0.6
8
150
5
1
-0.6
9
-38
10
-1
0.2
10
75-38
10
-0.312
0.2
11
106-75
10
0.132
0.2
12
150
10
1
0.2
13
-38
15
-1
1
14
75-38
15
-0.312
1
15
106-75
15
0.132
1
16
150
15
1
1
PS: particle size; PD: pulp density, X1: coded particle size; X2: coded pulp density
Results were modelled by multiple regression analysis [13] based in the empiric model: Q P = b 0 + b1 x1 + b 2 x 2 + b11 x12 + b 22 x 2 + b12 x1 x 2 (1) Equation (1) was used to generate response curves of the effect of particle size and pulp density on the maximum volumetric rate of production of soluble iron, QP. Maximum iron solubilisation rates were calculated as the slopes of the straight lines drawn from the iron concentration at time zero passing tangent to the solubilisation curve.
229
Bioleaching Applications
3.
RESULTS AND DISCUSSION
3.1 Bioleaching kinetics The results of the 16 runs are presented in Figure 1. At all particle size fractions, pulp density of 15% proved to be deleterious, a similar result than that obtained by Nemati and Harrison [14] with a similar Sulfolobus strain. This effect can be due to a number of factors, namely mechanical, metabolic stress and gas transfer limitations [7, 10, 14-16]. Attrition of the cells by the concentrate particles can result in mechanical damage due to the delicate nature of the archeal cell envelopes, while this same situation can cause metabolic stress. It is believed that stress can be overcome by an extended adaptation period. High pulp densities can cause oxygen and carbon dioxide demands higher than the actual supplies, as has been suggested by Gerike el al. [15], d’Hugues et al. [16] and Boogerd et al. [17]. High percent iron solubilisations were obtained for pulp densities of 2.5 and 5%, as can be seen in Table 3. Iron solubilisation was affected by increasing pulp densities and by coarser particles. In fact, most of the soluble iron obtained with 15% pulp density and the coarser fraction (run 16) came from chemical leaching as evidenced by the results of the uninoculated flasks (not shown). The contribution of chemical leaching became less important with decreasing pulp densities and particle sizes, and was not significant in the 2.5 and 10% pulp density runs. Table 3. Maximum solubilisation rates and percent solubilisation at different particle sizes and pulp densities Run
PS* (µm)
PD* (% w/v)
Iron solubilisation (%)
Solubilisation rate (g/L·h)
1
-38
2.5
86.2
0.819
2
75-38
2.5
64.0
0.740
3
106-75
2.5
82.9
0.712
4
150
2.5
69.9
0.776
5
-38
5
81.4
0.948
6
75-38
5
73.7
1.042
7
106-75
5
68.1
0.985
8
150
5
59.0
0.646
9
-38
10
45.6
1.009
10
75-38
10
40.9
1.089
11
106-75
10
29.3
0.794
12
150
10
31.5
0.420
13
-38
15
32.9
0.618
14
75-38
15
21.5
0.582
15
106-75
15
29.3
0.318
16
150
15
6.4
0.084
* PS: particle size; PD: pulp density
230
Bioleaching Applications
Total iron (g/L)
25
25
150-106 µm
20
20
15
15
10
10
5
5
0
0
10
20
30
40
50
0
106-75 µm
0
10
Time (d)
Total iron (g/L)
25
25
75-38 µm
20
15
15
10
10
5
5
0
10
20
30
Time (d)
30
40
50
40
50
Time (d)
20
0
20
40
50
0
-38 µm
0
10
20
30
Time (d)
Figure 1. Iron solubilisation kinetics from a pyritic gold concentrate by Sulfolobus metallicus in shake flasks at 68ºC initial pH 2.0 and agitation of 220 rpm. Pulp densities: ■ 2.5%; • 5.0%; ▲ 10%; ▼15%
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Bioleaching Applications
3.2 Iron solubilisation rates Maximum iron solubilisation rates are given in Table 3. Multiple regression analysis of these results allowed the evaluation of the coefficients of equation (1): Q P = 0.9591 − 0.2032x1 − 0.1862x 2 − 0.1304x12 − 0.3039x 22 − 0.1305x1x 2
(2) All the coefficients of equation (2) have a significant effect on QP as their F values are much higher than the tabulated F at the 95% confidence limit (F5,10 = 3.3, p < 0.05) [13]. F values also allow the conclusion that the interaction between particle size and pulp density had the weakest effect (F = 9.24) and that pulp density (F = 81.24) showed a stronger effect on QP than particle size (F = 14.15). The response surface generated by equation (2) is showed in Figure 2 and the corresponding contour plot is depicted in Figure 3. It can be seen that a maximum value of QP occurs at a pulp density of 7.8% and particle size of 35 µm, a value very near the fraction 75-38 µm. The -38 µm exhibits a negative effect on the biooxidation rate, but this effect is not as strong as that observed by Nemati et al. [10], who found a complete inhibition of microbial action at particle sizes under 25 µm.
1,0 QP (g Fe/L·d)
0,8 0,6 0,4 0,2
lp Pu
2.5
( ity ns de
5.0 10
% v) w/
15 -38
75 -3 8
10 675
15 010 6
(µm size e l c i t Par
)
Figure 2. Response surface of the effect of particle size and pulp density on the rate of iron solubilisation from pyrite by Sulfolobus metallicus in shake flasks at 68ºC initial pH 2.0 and agitation of 220 rpm
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Bioleaching Applications
Particle size (µm)
150-106
Productividad (g Fe/L·d) 0.99 -- 1.10 0.88 -- 0.99 0.77 -- 0.88 0.66 -- 0.77 0.55 -- 0.66 0.44 -- 0.55 0.33 -- 0.44 0.22 -- 0.33 0.11 -- 0.22 0.00 -- 0.11
106-75
75-38
-38
2.5
5.0
10
15
Pulp density (%w/v)
Figure 3. Contour plot of the effect of particle size and pulp density on the rate of iron solubilisation from pyrite by Sulfolobus metallicus in shake flasks at 68ºC initial pH 2.0 and agitation of 220 rpm 4.
CONCLUSIONS It is concluded that the operation variables particle size and pulp density have an effect on the rate of iron solubilisation from pyrite by Sulfolobus metallicus in shake flasks. Under the experimental conditions used in this work, the set of these variables that produce the highest rate is 35 µm particle size and 7.8% pulp density. It is also concluded that, in the range studied, the interaction between the variables is weak and that pulp density has a much stronger effect than particle size. ACKNOWLEDGMENTS This work was supported by the National Commission of Science and Technology through the FONDECYT projects 1000284 and 1020768. REFERENCES 1. A.D. Bailey and G.S. Hansford, Biotechnol. Bioeng., 42 (1993) 1164. 2. A.E. Torma, C.C. Walden and R.M.R. Brannion, Biotechnol. Bioeng., 12 (1970) 501. 3. A.E. Torma, C.C. Walden, D.W. Duncan and R.M.R. Brannion, Biotechnol. Bioeng., 14 (1972) 777. 4. M.T. Madigan, J.M.Martinko and J. Parker, Brock: Biología de los Microorganismos, 8th ed., p. 741, Prentice Hall, Madrid, 1998. 5. Α. Balows, K.H. Schleifer, H.G Truper, M. Dworkin and W. Harder (eds.), Prokaryotes, Vol. 1, p. 684, Springer Verlag, New York. 6. M.K. Toma, M.P. Ruklisha, J.J. Vanags, M.O. Zeltina, M.P. Leite, N.I. Galinina, U.E. Viesturs and R.P. Tengerdy, Biotechnol. Bioeng., 38 (1991) 552.
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7. P. d’Hugues, D. Morin and S. Foucher, Biohydrometallurgy: Fundamentals, Technology and Sustainable Development (V.S.T. Ciminelli and O Garcia Jr., eds.), Part A, p. 75, Elsevier, Amsterdam, 2001. 8. R.P. Hackl, F.R. Wrigh and L.S. Gormly, IBS ’89 International Symposium Proceedings, p. 533, Jackson Hole, WY, 1989. 9. D. Howard and F.K. Crundwell, Biohydrometallurgy and the Environment Toward the 21th Century (R. Amils and A. Ballester, eds.), Part A, p. 209, Elsevier, Amsterdam, 1999. 10. M. Nemati, J. Lowenadler and S.T.L. Harrison, Appl. Microbiol. Biotechnol., 53 (2000) 173. 11. P.R. Norris, IBS ’89 International Symposium Proceedings, p. 3, Jackson Hole, WY, 1989. 12. L. Herrera, P. Ruiz, J.C. Aguillon and A. Fehrmann, J. Chem. Technol. Biot., 44 (1989) 171. 13. J. Lawson, J. Madrigal and J. Erjavec, Estrategias Experimentales para el Mejoramiento de la Calidad en la Industria, p. 181, Grupo Editora Iberoamericana, México, 1992. 14. N. Nemati and S.T.L. Harrison, Biohydrometallurgy and the Environment Toward the 21th Century (R. Amils and A. Ballester, eds.), Part A, p. 473, Elsevier, Amsterdam, 1999. 15. M. Gericke, A. Pinches and J.V. van Rooyen, Int. Miner. Process., 62 (2001) 243. 16. P. d’Hugues, S. Foucher, P. Gallé-Cavalloni and D. Morin, Int. Miner. Process. 66 (2002) 107. 17. F.C. Boogerd, P. Bos, J.G. Kuenen, J.J. Heijnen and R.G.J.M. van der Lans., Biotechnol. Bioeng. 35 (1990) 1111.
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Enhancement of chalcopyrite bioleaching capacity of an extremely thermophilic culture by addition of ferrous sulphate A. Rubioa and F.J. García Frutosb a
Instituto Geológico y Minero de España (IGME), La Calera nº1. 28760 Tres Cantos, Madrid, Spain b Centro de Investigaciones Energéticas, Medioambientales y Tecnológicas (CIEMAT), Av. Complutense 22, Edificio 20, 28040 Madrid, Spain Abstract One of the mainly problems of the bioleaching processes applied to copper concentrates and refractory ores are the low kinetics of the reactions, with high residence time that does not permit it to be an economic process. For this reason, current researches into this field are focused on how to increase this bioleaching rate. Apart from improving engineering design of bioreactors, the possibilities to increase bioleaching rates depend on the use of catalyst and isolation and adaptation of new microorganisms with high capacity to leach these ores. Many researches have investigated the possibility of using thermophilic microorganisms to improve metal-leaching rates instead to mesophilic microorganisms. In this sense, it was obtained a mixed natural thermophilic culture from a typical chalcopyritic copper concentrate of the Spanish Pyritic Belt. This culture has selectivity with respect to leaching chalcopyrite, when this is present together with other sulphides in ores and concentrates, and presents high copper leaching rates at high pulp density [1]. In this paper the study of enhancement of the bioleaching capacity of this culture on a chalcopyritic copper concentrate of the Spanish Pyritic Belt when ferrous sulphate is added is presented. The results obtained show that an initial addition of 1.8 g/L of iron as ferrous sulphate increases the copper bioleaching rate in all pulp densities studied. These results are according with other investigations that suggest the enhancement in copper leaching of chalcopyrite obtained by addition of ferrous sulphate [2]. From these results authors suggest an indirect in situ mechanism of bioleaching of chalcopyrite for this culture in which ferrous ion have a significant role. Keywords: extremely thermophilic culture, ferrous iron catalysis, chalcopyrite, bioleaching 1.
INTRODUCTION The resources of high ores in the world are becoming more and more scarce being necessary the processing of more complex ores. Conventional mineral processing on complex sulphide ores, carried out by differential flotation, often produces concentrates 235
Bioleaching Applications
not enough clean and difficult to commercialise. Therefore, plenty effort has been put into of hydrometallurgical process development for treatment of these ores, but the greatest number of proposed methods are complex and expensive [3]. Biohydrometallurgical processes appear as an alternative. These processes were first applied industrially to copper and uranium extractions using bio-assisted heap, dump and in-situ technologies, are today successfully used in extraction of gold from refractory sulphide-bearing ores and concentrates [4]. However, for other metal concentrates this technology isn’t still viable alternative to conventional pyrometallurgical extractions. In case of chalcopyritic concentrates there is a more complicated situation due to natural refractivity of chalcopyrite. The leaching of chalcopyrite is slow and incomplete in relation to other sulphides, and this is thought to be as a result of formation of a passivating layer. The use of new microorganisms isolated and adapted and of catalyst could improve the bioleaching rates. Thermophilic microorganisms had been used in bioleaching process because can be to increase leaching rates due to high temperatures, tolerant capacity and metabolic characteristic [5, 6]. Moreover, utilisation of these microorganisms that naturally thrive on ore samples and in its aqueous environments could be good options, since that, those probably have chalcopyrite specificity and much higher capacity for adaptation. In this sense, the authors have obtained a mixed natural thermophilic culture from a typical chalcopyritic copper concentrate of the Spanish Pyritic Belt. This culture has selectivity with respect to leaching chalcopyrite, when this is present together with other sulphides in ores and concentrates [7]. The use of catalytic ions like silver can increase bioleaching copper rate. But other ions can be accelerating this process too. Although ferrous iron is a reagent involved in the process, its role in bioleaching process is not still overall understanding. In acidic solutions, chalcopyrite is oxidised by ferric ions and dissolved oxygen to release copper ions. Ferrous ions are rapidly oxidised to ferric ions in presence of iron oxidising bacteria such as Acidithiobacillus ferrooxidans. It appears that role of ferrous iron in leaching is only as a source of ferric ions. The role of iron in bioleaching of sulphur ores is complex and depends of the interactions between these in its different oxidation states, bacteria and minerals particulate. However, there are reports, which suggest that ferrous iron contributes to copper extraction from chalcopyrite. There is a critical potential at which the leaching rate is very great and the rate suddenly decreases above this potential. As the suspension potential increases with increasing ferric to ferrous ratio, these results indicate that the leaching rate is faster with an optimum concentration of ferrous ions that without ferrous ions [2]. Further, if addition of catalytic ions to bioleaching can effectively increase the oxidation rate, the combination of both thermophilic microorganisms and catalyses using ferrous iron could aid to increase the slow kinetic problem that exhibit this process. This paper is presented in context of copper bioleaching, where it is assumed that the ferric ions and protons produced by microbial action, act as the leach agents in copper dissolution. The objective of this study is to establish the rate of chalcopyrite dissolution as a function of to added ferrous iron, determining the role of this in copper extraction with this thermophilic culture.
236
Bioleaching Applications
2.
METHODS
2.1 Ore sample A chalcopyritic copper concentrate obtained from conventional differential flotation was used to obtain the mixed thermophilic culture. This concentrate came from the Spanish Pyritic Belt and its composition is given in Table 1. The mineralogical composition of this sample shows chalcopyrite and pyrite as main mineralogical species and galena and sphalerite as secondary mineralogical phases. The particle size distribution presents a passing d80 of 20µm. Table 1. Chemical composition of copper concentrate Chemical analysis (wt %) Cu
Zn
Pb
Fe
S
23.37
2.58
2.52
32.63
38.00
2.2 Microbial inoculum A mixed thermophilic culture (MTC) of native microorganisms, isolated on this chalcopyritic concentrate was used as inoculum. The optimal condition of this culture is a 65°C temperature, with a pH of 1.30. For all tests done, the inoculum was obtained by the following way. The final pulp of the bioleaching test was filtered, obtaining the leach liquor and the solid residue. The solid was intensively stirred during two hours with a solution at pH 4. This pulp was filtered again and the final liquid obtained, which contains most the bacteria, which were attached to the solid residue, was filtered through a 0.22 µm Millipore filter where bacteria were retained. Finally, the bacteria were re-suspended in 50 mL of the leach liquor obtained in the first filtration in order to get the inoculum volume (5% v/v) for the next bioleaching test. The cultures were successfully adapted to higher pulp density of copper concentrate until obtaining high copper extractions in short residence times. 2.3 Bioleaching experiments Bioleaching experiments were carried out in 1 L glass cylindrical reactors provided with a cap with four holes to allow mechanical stirring (at 130 rpm), aeration (10-15 L/h) and sampling. These reactors were placed in a thermostatic bath to keep the temperature constant at 65°C. During the experiments the pH was kept at 1.3 by the addition of 10N H2SO4 when were necessary. This was made to avoid the precipitation of iron in form of jarosites, which damage the bioleaching process. Redox potential and pH were measured daily, while the levels of copper, zinc, and iron in solution were analysed daily or every two days, depending on the test. Water was added to the reactors in order to compensate for evaporation losses. Once bioleaching tests were finished, solids were removed by filtration, and chemically characterised as well as the leachate.
237
Bioleaching Applications
Sterile bioleaching tests were carried out with ore sterilised by autoclaving at 121°C, 30 minutes and 1atm of pressure, and adding a solution of 10% ethanol to the leaching media. To study the effect of the addition of ferrous iron on the bioleaching process, the test were carried out using ferrous sulphate solutions of varying concentration that were added initially to the leaching media. The ferrous iron concentration varied from 400 mg/L to 9000 mg/L depending of the tests. 2.4 Analysis Soluble species of copper, zinc, lead, total iron and minor elements were analysed by ICP [8] using a spectrophotometer ICAP-61 Thermo Jarrel Ash. The ferrous iron was analysed by a volumetric method by titration with potassium dichromate [9]. Copper, zinc, lead and iron content samples and leaching solid residues were analysed by XRF [10] using a spectrophotometer Philips PW-1404, and minor elements by ICP. Total sulphur was gravimetrically determined and elemental sulphur was analysed by toluene extraction in a Soxhlet apparatus. The pH was measured with a 704 pH-meter Metrohm. The redox potential (Eh) was measured with a platinum electrode with an Ag/AgCl reference electrode. Mineralogical composition was determined by XRD using a diffractometer PW-1700 Philips. 3.
RESULTS AND DISCUSSION Initially the were carried out at 1% pulp density and using concentrations of ferrous iron of 450, 900, 1800, 3600 and 4500 mg/L. Figure 1 shows copper extraction during the tests carried out with more representative ferrous iron concentration. Initial Fe2+addition increase copper overall leaching, obtaining with 1800 mg/L more than 96% of extraction in only 69 hours. With 3600 mg/L of ferrous iron addition the catalyst effect is maintained, but with higher ferrous iron concentrations, 4500 mg/L, copper extractions was diminished. At early hours, this catalysed effect is not appreciable, having lower copper extractions when ferrous iron was added respecting to the control (no addition of ferrous sulphate). 100
Cu extraction (%)
90 80 70 60 50
Fe2+(mg/L) control 900 1800 3600 4500
40 30 20 10 0 0
15
30
45
60
75
90
105
120
135
150
165
Time (Hours)
Figure 1. Copper extraction evolution with the different ferrous iron concentration added. (1% pulp density (w/v), MTC culture at pH 1.30 and 65°C) 238
Bioleaching Applications
Relationship between both total iron and ferrous iron in solution are shows in Figure 2. In all tests during the first hours iron in solution was in ferrous form, without apparent oxidation of Fe2+ added. This form of iron was increased in solution with all ferrous iron concentration added, but after 45 hours of assay the concentration was practically constant in solution. Only is observed a reduction in the solution ferrous iron concentration, with 900 and 1800 mg/L of ferrous iron added, when copper had been leached. 7
FeT concentration (g/L) Control
0.9
1.8
7 3.6
4.5 6
5
5
4
4
3
3
2
2
1
1
0
2+
concentration (g/L)
Control
6
0
Fe
0 15 30 45 60 75 90 105 120 135 150 165 0 Time (Hours)
0.9
1.8
3.6
4.5
15 30 45 60 75 90 105 120 135 150 165 Time (Hours)
Figure 2. Relationship between iron forms present in solution. (Left) Total iron. (Right) Ferrous iron. (1% pulp density (w/v), MTC culture at pH 1.30 and 65°C) In this sense, it seems that both ferrous iron, ferrous iron added and ferrous iron produced by chalcopyrite leaching reaction, only are being oxidised when the chalcopyrite was not available. Ferric iron in solution was not higher than 30% of total iron in all tests. Figure 3 shows the zinc bioleaching of the copper concentrate during the tests. This present a typical kinetic observed in all tests, in which the zinc did not leached whilst copper was bioleaching. Thus, in 1800 mg/L of ferrous iron addition test, can be clearly observed zinc bioleaching after total copper bioleached (72 hours). 100
Zn extraction (%)
90
control
Fe2+ (mg/L) 900 1800
3600
4500
80 70 60 50 40 30 20 10 0 0
15
30
45
60
75
90
105
120
135
150
165
Time (Hours)
Figure 3. Zinc bioleaching evolution with the ferrous iron concentration added. (1% pulp density (w/v) (MTC culture at pH 1.30 and 65°C) 239
Bioleaching Applications
pH evolution was similar in all tests, increasing in the first hours but was enclose in 1.30 adding sulphuric acid (10N) to avoid iron precipitation. XRD analyses of final solids of all test is showed in Table 2. The copper leaching produced elemental sulphur as principal species, increasing the presence of jarosites when ferrous addition was increased. Table 2. XRD diffraction of final solids obtained in bioleaching tests Addition of Fe2+ (mg/L) Control 900 1800 3600 4500
Main species
Secondary species
Minor species
Anglesite, Pyrite Pyrite, Elemental sulphur Pyrite, Elemental sulphur Pyrite, jarosite, Elemental sulphur Pyrite, Elemental sulphur, jarosite
Anglesite -
Elemental sulphur Jarosite Jarosite, anglesite -
At the same way as done with 1% of pulp density, bioleaching tests adding ferrous iron were carried out at 5% and 10% of pulp density using ferrous iron addition of 900, 1800, 2400, 3600 and 9000 mg/L. In Figure 4 can be observed that with 1800 mg/L of ferrous iron addition, also better copper extractions was obtained, with a copper extraction of 80% in only 165 hours. 100
Cu extraction (%)
Fe2+(mg/L)
90
Control
80
900
70
1800
60
2700
50
3600
40
9000
30 20 10 0 0
15
30
45
60
75
90
105
120
135
150
165
Time (Hours)
Figure 4. Copper extraction evolution with ferrous iron concentration added. (5% of pulp density (w/v), MTC culture at pH 1.30 and 65°C) At 10% of pulp density, copper extraction of 90% in 270 hours was obtained when 1800, 2700 and 3600 mg/L of ferrous iron were added (Fig. 5). The evolution for the rest of parameters studied, such as ferrous and ferric iron in solution and zinc bleaching, were similar as obtained at 1% of pulp density. In all bioleaching tests carried out, addition of ferrous iron increase copper extraction with this thermophilic culture. Addition of 1800 mg/L of ferrous iron seems to have the optimum catalysed effect in all pulp densities studied (1, 5 and 10% (w/v)). This optimum extraction could be related wit lesser presence of jarosites in solid residues. Most of the soluble iron, was present in ferrous form and the ferric concentration was negligible. The ferrous iron was produced by chalcopyrite leaching, and only when chalcopyrite is almost oxidised, ferrous iron oxidation is observed. Ferrous oxidation by this thermophilic culture is very low in liquid medium (data confirmed in early tests in 9K medium). 240
Bioleaching Applications 100
Cu Extraction (%) Fe2+ addition (mg/L) control 1800 2700 3600
90 80 70 60 50 40 30 20 10 0 0
50
100
150
200
250
300
Time (Hours)
Figure 5. Copper extraction evolution with ferrous iron concentration added (10% of pulp density (w/v), MTC culture at pH 1.30 and 65°C)
Trying to explain leaching mechanisms with this thermophilic culture, authors think that, in acidic solution, chalcopyrite can be principally oxidised by dissolved oxygen according to the following reaction: CuFeS2 + O2 + 4H+ = Cu2+ + Fe2+ + 2So+ 2H2O (1) In addition to this, the ferric ion that can be produced by oxidation of ferrous iron in acidic medium through Eq. (2) was rapidly utilised to leach the chalcopyrite by Eq. (3). 4Fe2++ 4H+ + O2 = 4Fe3+ + 2H2O (2) 3+ 2+ 2+ o CuFeS2 + 4Fe = Cu + 5Fe + 2S (3) taking place the chalcopyrite oxidation. Considering the final products obtained in our case, an indirect leaching mechanism could be postulated. However, ferric iron in solution was minimum and MCT possess very low capacity to oxidise ferrous iron in solution. Besides, according with rest potentials, sphalerite should be the most easily mineralogical species to leach, but this is not true, in our case. Because of that, the authors think according to postulated in (2), ferrous oxidation to ferric ions is produce onto chalcopyrite surface, where ferrous ions added are adsorbed on chalcopyrite surface. This ferrous iron is oxidised on surface by culture not in leaching media. This process that occur also with no addition of ferrous iron, have a catalyse effect when initial ferrous ions were added. These, provide of ferric ions that rapidly oxidises chalcopyrite producing more ferrous iron that can be adsorbed in other points of the chalcopyrite surface acting in the same way. When the chalcopyrite leaching is almost exhausted, ferric ion concentration is high passing to the solution and could leach sphalerite by ferric leaching, as can be observed in Figs 2 and 3. 4.
CONCLUSIONS A natural mixed thermophilic culture was obtained from a chalcopyritic copper concentrate with an ability to preferentially leach chalcopyrite in concentrates and with high leaching rates at high pulp densities. In our case, the enhancement of chalcopyrite bioleaching capacity of the mixed thermophilic culture by addition of ferrous ions confirm the indirect in situ leaching mechanism, postulated in our previous work [1], for this culture respecting the chalcopyrite. 241
Bioleaching Applications
The results obtained in this study confirm the main role of iron forms in the bioleaching processes of chalcopyrite. From these laboratory results this culture can be considered as a promising advance in the biohydrometallurgical treatment of chalcopyritic concentrates and its potential use on an industrial scale. REFERENCES
1. A. Rubio and F.J. García Frutos (2002) Mineral Engineering, 15, 689-694. 2. Hiroyosi, N et al. Hydrometallurgy. 47 (1997), 37-45. 3. J.L. Alvarez, Simposio sulfuros polimetálicos de la Faja Pirítica Ibérica, Huelva, (1996). 4. M. A. Jordan, S. McGuiness and C.V Philips, Minerals Engineering, 9, No.2 (1996) 169. 5. C. L Brierley, International conference and workshop application of biotechnology to the mineral Industry, Australian Mineral Foundation (1993) 2.1. 6. D. A. Clark and P.R. Norris, Minerals Engineering, 9, No.11 (1996) 1119. 7. A. Rubio. PhD Thesis (1998). Dpto. Biología Molecular. Universidad Autónoma de Madrid. 8. S. Del Barrio, Boletín Geologico y Minero (Special issue) (1992). 9. I.M Kolthoff, Análisis químico cuantitativo, Ed. Niger SRL (1979). 10. J.A Martín Rubí, IV Simposio Internacional de sulfuretos polimetálicos de la Faixa Piritosa Ibérica, C10 (1998) 1.
242
15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Evaluation of microbial leaching of uranium from Sierra Pintada ore. Preliminary studies in laboratory scale Patricia Silva Paulo*1, Diego Pivato1, Ana Vigliocco1, Julian López2, Alberto Castillo3 1
Unidad Aplicaciones Tecnológicas y Agropecuarias, Centro Atómico Ezeiza Unidad de Operación de Instalaciones Nucleares, Centro Atómico Ezeiza 3 Unidad de Proyectos Especiales y Suministros Nucleares Comisión Nacional de Energía Atómica, Av. del Libertador 8250 (1429) Buenos Aires, Argentina 2
Abstract The feasibility of bacterial leaching was studied in our laboratory with mineral from Sierra Pintada mine (Province of Mendoza, Argentina). It is a low U grade ore with 6% of carbonates and low quantity of pyrite. The experiments were performed in shake flasks and scaled up to static flood tanks. Shake cultured leaching studies were carried out with 0.5 inch crushed ore and a pulp density of 10%. Acidithiobacillus ferrooxidans (ATCC 33020) was grown in 9K medium and used as inoculum. Ferrous sulphate (FeSO4) was added as energy source. Percent of U leached and acid demand to maintain acidity values of pH=2 were measured during the experiment. Three conditions were tested: ore, ferrous sulphate and inoculum; ore and FeSO4 and a control with ore but without inoculum or FeSO4. After 90 days, the acid added in the flask with bacteria and FeSO4 was 20 times less than the control, and the U leached yielded 90.5%, that is 83% more than the control. Scale up experiments were performed in static flood tank system. Sixteen kg of 0.5 inch crushed ore were placed in a static tank reactor and flooded with 32 liters of 9K medium without FeSO4. Ferrous sulphide (SFe) was added as energy source. Gentle fine bubble aeration was accomplished by passing air through diffusers, and liquid recirculation was forced by a pump. At. ferrooxidans was grown in 9K medium and then used as inoculum. Three different conditions were tested: ore, inoculum and SFe; ore and inoculum, and a control without inoculum (ore only). Daily pH was measured and sulphuric acid added to maintain pH=2.0 was recorded. Weekly U leached was measured. The assay was carried on for five months. An initial 7 weeks lag phase was observed for all three conditions. This was associated to pH adjustments required for neutralization of acid demand due to the high carbonate presence in the ore. After five months, 72.8% of U was extracted in the At. ferrooxidans and SFe tank, with a total of 43.0 g of sulphuric acid added per g of U leached. An uranium extraction of 32.3% and 73.2 g acid/g U leached was observed in the ore and At. ferrooxidans (no energy source added) tank, while the control showed 3.6% of U leached and 73.4g acid added /g U leached. * Corresponding author. E-mail:
[email protected]
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The slowness of the process may be due to gypsum and jarosite precipitation on the mineral, impairing bacterial action and liquid recirculation as well. Both, shake flasks and static flood systems results, showed that the addition of an energy source is needed to obtain a reasonable rate of uranium leaching. A second static flood system experiment was performed, with an acid cure treatment before bacteria addition and counter current media flow. Different energy sources were also tested. The four conditions were: ore, inoculum and carbon steel scrap as energy source; a blank without inoculum; ore, inoculum and exogenous pyrite as energy source, and its corresponding blank without bacteria. After a month, 90.7% of U was extracted from the inoculated and carbon steel scrap tank with an acid consumption of 60.9 g/g U leached. In the tank with pyrite and inoculum an 89.6% of U leached was observed with 54.9g acid /g U. The second static flood experiment was notoriously effective in diminishing the leaching time, and avoiding precipitation that cause blockages in tubing. Both scrap and pyrite have shown to be suitable and economic energy sources. The experiments describe here are a promising first step in the evaluation of a possible pilot scale application. 1.
INTRODUCTION In Argentina the uranium industry is controlled by the federal government through the Atomic Energy National Commission (CNEA). In the fifties, the geological group found several uranium deposits around the country. In the sixties, the mining exploitation began to supply the fuel to the nuclear power plants Atucha and Embalse de Río Tercero. Atucha (RWU type) works with low enriched uranium and Embalse de Río Tercero (Candu type) only with natural uranium, consuming nearly 120 tnU/year. From 1979 to 1995 CNEA produced, by heap leaching, 1,000 tU in yellow cake form in Sierra Pintada plant (Mendoza province), to supply the fuel to the two nuclear power plants. During the nineties, due to the lower price in the spot market and overvaluation of the currency, the Sierra Pintada mine and the yellow cake production plant were shut down and Argentina begun to buy yellow cake in the spot market. Nowadays both activities have been started again and, the engineering group started to work on some other methods to reduce the cost and improve leaching and waste treatment. The mineral of this mine is sandstone with high concentration of calcium carbonate. This situation demands huge amounts of sulfuric acid to extract the uranium from the rock. The sulfuric acid, the mining operation and the manpower, are the most significant costs in yellow cake production. The main tasks implemented were: new design in the open pit, changing heap leaching by flooded leaching, and studies on bioleaching. In 1999, combined work began between two groups from CNEA, the Engineering and the Microbiology groups, to investigate bioleaching, as reported in this paper. Microbial leaching is a simple and effective technology used for metal extraction from low-grade ore. Metal recovery is based on the activity of chemolithotrophic bacteria, mainly Acidithiobacillus ferrooxidans, At. thiooxidans and Leptospirillum ferrooxidans (1-3). 244
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Most of the uranium ores occur as a mixture of mineral containing uranium in either the tetravalent or the hexavalent state. Uranium is soluble only in its most oxidized hexavalent state (4). Tetravalent uranium can be oxidized to the hexavalent state by ferric iron, but oxidation occurs much more rapidly in the presence of the iron-oxidizing At. ferrooxidans(5). Bacterial leaching of uranium occurs via an indirect mechanism in which the bacteria oxidize the pyrite in the ore, generating an acidic ferric sulfate solution which carries out the chemical oxidation of the tetravalent uranium to the soluble hexavalent state. The studies reported here were initiated in order to determine the feasibility of microbial leaching of Sierra Pintada ore and whether bioleaching could contribute in the costs reduction by increasing the uranium recovery rate, reducing the acid consumption and / or diminishing the leaching time. Bioleaching of uranium from Sierra Pintada ore was carried out using Acidithiobacillus ferrooxidans at bench scale; exogenous ferrous compound were added for laboratory tests due to the low content of pyrite in the ore. 2.
MATERIALS AND METHODS
2.1 Bacterial culture preparation A pure strain of Acidithiobacillus ferrooxidans (ATCC 33020) was used in this study. The bacteria were cultured in 2L Erlenmeyer flasks with 1L of basal medium containing 0.1g KCl, 3.0g/l (NH4)2SO4, 0.5g/l MgSO4.7H2O, 0.5g/l K2HPO4.3H2O, 44.22 g/l FeSO4.7H2O, 0.01 g CaNO3 ( 9K medium) (6). The pH of the medium was adjusted to 1.80 with sulfuric acid. The culture was kept at 30°C in a rotary shaker at 100 r.p.m. The culture in exponential growth was used directly as inoculum in each experiment. 2.2 Analytical methods The majority of the uranium analyses were performed by LivestockGroup laboratory. The uranium content of the ore and pulp residues was determined by the Laboratory Section of Complejo Minero Fabril San Rafael, Province of Mendoza. Uranium in leach liquor and pulp residues was determined spectrophotometrically by using dibenzoylmethane method (8). Daily pH was measured and the volume of sulfuric acid added to maintain pH 2.0 was recorded. The bacterial viability in both shake flasks and static tanks was checked by subculturing 1 ml of supernatants in 9ml of 9K medium twice a month. 2.3 Uranium ore The Sierra Pintada ore consisted of moderately well-sorted grains of quartz, feldspar and rock fragments cemented by calcite with minor clay replacement. This mineral is a sandstone with high quantity of carbonates. It is a low-grade ore, with an average of U3O8 content of 0.15%, 3.8% of carbonates and a low quantity of pyrite. The radioactive mineralization occurs mainly as uraninite, brannerite and coffinite. Analysis of the ore gave average values of 0.26% magnesium, 0.14% phosphorous, 2.6% calcium and 1.72% total iron (9).
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2.4 Pyrite Pyrite from an outcrop near Sierra Pintada mine was characterized and used in leaching experiments. Sulfur and iron contents were 30% and 38% respectively. 2.5 Shake culture leaching Four Erlenmeyer flasks with two liters of medium containing 0.5g/l K2HPO4, 0.5g/l (NH4)2SO4, 0.5g/l MgSO4.7H2O at pH: 2.0 (TyK iron-free medium) were inoculated with 100 ml of a 96 hour culture of Acidithiobacillus ferrooxidans (7). Two hundred grams of Sierra Pintada ore (>1.0, <1.3cm) was added to the medium (10% pulp density) in each flask. Ferrous sulfate (FeSO4.7H2O) 2% w/w was added as energy source when it was necessary. The initial pH was 2.0 in all flasks. 2.6 Static flood tank system leaching The system consisted of two plastic tanks (38cm x 30 cm) connected in series. Gentle fine bubble aeration was supplied by an aquarium air pump (fig. 1). One tank contained TyK medium and the other was packed with 16 kg of uranium ore crushed and screened to > 1.0 <1.3 cm. The iron free TyK medium was continuously cycled through the ore by means of a submerged centrifugal water pump. Different energy sources were tried: 2% w/w of ferrous sulphide (1.5 cm); 2% w/w exogenous pyrite (100 Mesh) and 0.5% w/w of carbon steel scrap. A
B
E D
C
F
Figure 1. Scheme of static flood tank system (A: Aquarium air pump; B: T y K medium; C: Centrifugal water pump; D: Valve; E: Ore and medium flooding; F: Filter bed; --- Fluid circulation direction) 3.
RESULTS
3.1 Shake-culture leaching Four flasks containing 10% w/w uranium ore were incubated, three of them with bacterial inoculum and one abiotic control. Ore with neither inoculum nor FeSO4 (Flask A) worked as control. The other had ore and inoculum (Flask B); ore, FeSO4 and inoculum (Flask C) and ore with FeSO4 (Flask D). The shake cultures were in TyK iron-free medium at room temperature and shaken at 100 rpm. The uranium grade of the ore used in the assay was 1600 µg U/g ore. Bioleaching was monitored by following the uranium content in the supernatants. Acid demand to maintain acidity values of pH 2.0 were measured and recorded. The assay was 246
Bioleaching Applications
carried out for 90 days. At this stage the uranium in the pulp residues was analyzed. The data are presented in Table 1 and Fig. 2. Table 1. Shake culture leaching of Sierra Pintada ore with At. ferrooxidans (At. fe) for 90 days Flask A- Control B- Ore+At. fe C- Ore +Fe2+ + At. fe D- Ore + Fe2+
% Uranium in leached liquor
% Uranium in pulp residue
6.9 74.0 90.5 80.7
85.0 22.5 6.2 10.7
Acid Consumed g/ Kg ore g/ g U leached 312 2227 193 132 15 20 124 71
The yield of leaching was about 90% after 90 days in the flask C with bacteria and FeS04 with the lowest acid consumption. In contrast, only about 7% yield of uranium was achieved in Flask A, the control (Table 1). Flask B containing only At. ferrooxidans extracted 74% of the uranium from the ore. The reason for this high percentage of uranium recovery respect to the control could be due to the iron present in the ore which was used by the bacteria. The time course of leaching is seen in Fig. 2. 100.0
% U Leached
80.0 60.0 40.0 20.0 0.0 0
20
40
60
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100
Time (Days) Control Flask
Flask B
Flask C
Flask D
Figure 2. Shake flasks microbial leaching of uranium by At. ferrooxidans using ferrous sulfate as energy source
At the end of the assay, flask D showed 80% of U leached. This result could be explained by confirmation of At. ferrooxidans presence by sampling flask D supernatant and subculturing in 9K medium. It could have been contaminated during manipulation of samples. This first experiment showed that bioleaching process is possible in Sierra Pintada ore and that an energy source is necessary for a faster uranium extraction. 3.2 Static flood tank leaching system According to the results obtained in shake flasks leaching, a static flood leaching system was tested and scaled up 80 times respect to the flasks (Fig. 1). Sixteen kg of >1.0 <1.3 cm crushed ore were placed in a static tank reactor and flooded with 32 liters of TyK iron-free medium by using industrial grade reagents and tap water. Three different 247
Bioleaching Applications
conditions were tested: tank 1: ore, inoculum and ferrous sulphide (FeS); tank 2: ore and inoculum without external energy source and control tank without inoculum nor FeS. The assay was carried out at room temperature. The pH was measured daily and the sulfuric acid added to maintain pH 2.0 was recorded. The assay was monitored by measuring the uranium content in leach liquor. Periodically aquarium pump valve and tubing were checked and the blockages removed. After five months of incubation, the experiment was finished because of the increase of uranium leached in the liquor stopped. The uranium mineral in Sierra Pintada is hosted in sandstone deposits mostly containing 3.8% carbonates and minor amounts of pyrite (9). The leaching from this carbonate bearing sandstone uranium ore requires sulfuric acid addition to neutralize the carbonates. As a consequence during the first seven weeks of the assay, important pH adjustments were made (Fig. 3). It is obvious that carbonates buffer the acid bioleaching solution and may adversely affect the metabolic activity of acidophilic At. ferrooxidans (10,11).
Acid consumed (g)
800
600
400
200
0 0
5
10
15
20
25
Time (weeks) Tank 1
Tank 2
Control Tank
Figure 3. Acid consumption in the static flood tank system during bioleaching assays
Table 2 shows the performance of the static flood bioleaching system. Table 2. Bacterial static flood leaching of Sierra Pintada ore at room temperature with At. ferrooxidans for five months. Ferrous sulphide (Fe2+) added as energy source Tank 1- Ore +Fe2++At. fe 2 - Ore + At. fe Control - Ore
% Uranium in leached liquor
% Uranium in pulp residue
72.8 32.3 3.6
33.5 50.2 86.0
Acid Consumed g/ Kg ore g/ g U leached 45.3 43.0 34.3 73.2 38.1 734.6
After five months, 73% of uranium was extracted in tank 1 with bacteria and FeS, with a total of 45 g of sulfuric acid added per kg of ore during the experiment. Thirty three percent of uranium recovery and 34 g of acid/kg ore was recorded for tank 2 without energy source addition, while the blank without inoculum or FeS (Control tank) showed 3.6% of uranium leached and 38 g acid/kg ore (Fig. 4). It can be seen in Fig. 4 that the rate of uranium leaching becomes reasonable after 7 weeks. In that initial period, acid additions were necessary to maintain a pH value in At. ferrooxidans at optimal range. 248
Bioleaching Applications 100.0
% U leached
80.0 60.0 40.0 20.0 0.0 0
5
10
15
20
25
Time (weeks) Tank 1
Tank 2
Control Tank
Figure 4. Microbial leaching of uranium by At. ferrooxidans using ferrous sulphide as energy source: static flood tank system
The percentage of uranium recovery reached 72.8% in tank 1 after fourteen weeks and it becomes almost constant. In later stages of the experiment, bioleaching rates decreased probably because during the process gypsum (CaSO4) was formed and precipitated on the ore particles in the same way as jarosite did, diminishing the available surface for bacterial attack. Moreover, recirculation and aeration systems were frequently blocked by gypsum, jarosites and other inorganic precipitates. This blockage altered the optimal conditions for At ferrooxidans activity and periodic maintenance of tubing became necessary. At the end of the experiment the efficiency of At. ferrooxidans activity in uranium extraction was evident. Both, uranium leaching rate and acid consumed to leach one gram of uranium were notoriously better in tank 1 with inoculum and FeS, than in the other two conditions (Table 2). The addition of an energy source was necessary to obtain a reasonable rate of uranium leaching as it can be seen when comparing tank 1 and tank 2 performances (Fig. 4). This response to addition of Fe2+ was much more evident in static tank system than in shake flasks (Fig. 2). Based on the results and conclusions of the static flood system, a new experiment was carried out in order to improve the performance of the system. Fourteen kg of >1.0 <1.3 cm crushed ore were placed in a tank reactor and flooded with 28 liters of TyK iron free medium. The media recirculation was in a counter current direction. An acid cure treatment was performed before leaching to diminish the leaching time and to avoid the gypsum precipitation which causes blockages in pump valves and tubing. A better media recirculation through the ore and aeration were expected. At the same time a better bacteria performance should be observed. The cure acid treatment was carried out for a week. A total 55.0g of concentrated sulfuric acid was added per liter of TyK media and recirculated through the ore to permit carbonates dissolution and to maintain a pH value between 2.0 and 3.0. Four different conditions were tested: tank 1: ore, inoculum and carbon steel scrap; control 1: ore and carbon steel scrap without bacteria; tank 2: ore, inoculum and exogenous pyrite, and control 2: with ore and pyrite but without inoculum. After the acid cure treatment, 0.5% w/w of carbon steel scrap (tank 1 and control 1) and 2.0% of pyrite (tank 2 and control 2) were added. The next day, tanks 1 and 2 were inoculated with At. ferrooxidans and was considered the day one of the experiment. As it 249
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was described before, the assay was monitored by measuring U content in the liquor weekly. The assay was carried out at room temperature, and acid demand for pH 2.0 maintenance was measured and recorded for each tank. After 33 days, the assay was finished because most of the uranium contained in the ore was leached. The remain uranium in the pulp residues was determined. The final data are summarized in Table 3 and the time course of process is showed in Figure 5. Table 3. Static flood bioleaching system of Sierra Pintada ore using carbon steel scrap and pyrite as exogenous sources of energy. Results obtained after 33 days of incubation Tank 1- Ore + scrap + At. fe Control 1- Ore + scrap 2- Ore + pyrite+At. fe Control 2- Ore + pyrite
% Uranium in leached liquor
% Uranium in pulp residue
90.7 91.1 89.6 48.5
9.3 8.9 10.4 51.5
Acid Consumed g/ Kg ore g/ g U leached 74.6 60.9 74.3 60.4 66.5 54.9 61.3 93.6
100.0
% U Leached
80.0 60.0 40.0 20.0 0.0 0
1
2
3
4
5
Time (weeks) Tank 1
Tank 2
Control 1
Control 2
Figure 5. Static flood tank system: microbial leaching of uranium using At. Ferrooxidans. Carbon steel scrap and pyrite were used as energy source
After 33 days of incubation the tank with bacteria and carbon steel scrap extracted 90.7% of uranium with 60.9 g of acid added per gram of uranium leached. Tank 2 with pyrite as energy source showed 89.6% of U extracted and 54.9 g acid/g U leached (table 3). Carbon steel scrap is a waste product of carbon steel pieces turned in a turning lathe. In this experiment, both carbon steel scrap and pyrite were used to tested weather it could be an alternative of commercial FeS and the results obtained showed that it is possible (table 3, Fig. 5). Comparison of acid consumed in tank 1 and 2 showed that it was slightly minor (9.0%) in tank 2 (containing pyrite (S2Fe)) than in tank 1 (scrap). This difference could be attributed to At. ferrooxidans acid production by sulphur oxidation from pyrite. However that difference was lower than we had expected. The absence of adaptation of the bacteria to pyrite as energy source could be a reason. In future experiments a major advantage of pyrite as substrate will be evaluated.
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As can be seen in Fig. 5 initially control 1 performance was according to a blank. Afterwards the U leached was increasing and at the end of the experiment it was as in tank 1 with bacteria. The At ferrooxidans presence in control 1 was confirmed by microscope observation and supernant subculturing in 9K medium. The gypsum formation was not evident and tubing clogging did not occur. The performance of the assay was notoriously improved by diminishing the leaching time approximately five times respect to the first static flood assay. The uranium recovery was also increased in almost 20%. 4.
COMMENTS AND CONCLUSIONS Laboratory bench scale approaches to bioleaching evaluation usually involves shaken flask techniques. It is a suitable methodology for screening and preliminary testing because complete metal recovery is reached in a few days. However it has some limitations. The steady state cannot be reached due to continuous change of conditions and the ore sample homogeneity is difficult to achieve because the amount of ore quantity employed is usually small. The ore property and characteristics should be considered when a biological leaching process is being evaluated on a bench scale. No two ores are identical and within each ore deposit the mineralogical composition and the concentration of metals show heterogeneity. These variations demand experimental evaluation of multiple samples even from a single ore body, because they have a major effect on the microbial leaching of the material (12). After a few approaches to bioleaching using shake flasks, a static tank reactor technique was attempted. This technique represented a scale up of 80 times over the shake flask method. The ore material homogeneity had less influence in the results than in the reduced scale. Moreover the ore material could be more representative of the material that will be used in large-scale commercial application. The tank leaching studies have longer duration and develop different zones within the ore that differ in redox potential, iron precipitation, chemical and physical gradients. These characteristics are similar to large-scale applications, which is an advantage when the objective is a pilot plant scale. The presented results showed different performance of bioleaching between shake flask and the static flood techniques tested. We concluded that static flood tank system gave better results through being a larger scale system. The first static flood system experiment ran with some difficulties, mainly design variables, that were solved by introducing counter current flow leaching and acid cured treatment. This impaired the undesirable precipitates that cause tubing blockages. Counter current flow leaching improved recirculation and reduced ore compaction. As a consequence, aeration and pH value improved, and a more optimal At. ferrooxidans activity condition was reached. An improved kinetics of the leaching process and considerably shorter leaching period were attained. Biological leaching profiles of mineral samples from every ore are unique. In this paper the ore tested was poor in pyrite content so different ferrous sources were tested that permitted a reasonable leaching rate. Commercial ferrous sulphide, carbon steel scrap and pyrite were attempted. Both, carbon steel scrap and pyrite were effectively used as energy 251
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substrates by At. ferrooxidans and can also be considered as suitable iron sources for bioleaching. The Sierra Pintada ore is alkaline in nature; as a consequence the bioleaching process using acidophilic bacteria becomes a slow process and has a high acid demand to neutralize the carbonates. In future experiments, an At. ferrooxidans culture adapted to pyrite will be used as inoculum in order to become an acid producer reaction diminishing the acid consumption. The presented preliminary results indicate that it is possible to use bioleaching for uranium recovery from Sierra Pintada ore; that is necessary to add an exogenous iron source and the results also are a promising first step in the evaluation of a possible pilot scale application in Sierra Pintada ore. REFERENCES
1. Norris, P., (1990). Acidophilic bacteria and their activity in mineral sulphide oxidation. p. 3-29. In Microbial Mineral Recovery: Ehrlich, H.; Brierley C. Mc Graw-Hill Publishing Company ISBN 0-07-00781-9. 2. Rawlings, D.E., (1997). Mesophilic, autotrophic bioleaching bacteria: Description physiology and role. P. 229-241. In: Biomining: Theory, Microbes and Industrial Processes. Berlin: Springer-Verlag. 3. Sand, W., Rhode, K., Sobotke, B. and Zenneck, C., (1992). Evaluation of Leptospirillum ferrooxidans for leaching. Applied and Environmental Microbiology. 58:85-92. 4. Brierley, C.L., (1978). "Bacterial leaching. CRC Critical Reviews". Microbiology 6: 207-262. 5. Lundgren, D.G.; Silver, M. (1980). "Ore leaching by bacteria". Ann. Rev. Microbiol. 34: 263-283. 6. Silverman, M.P., Lundgren, D.G. (1959) "Studies on the chemoautotrophic iron bacterium Thiobacillus ferrooxidans. An improved medium and a harvesting procedure for securing high cellular yields". J. Bacteriol. 77: 642-647. 7. Touvinen, O. H., and Kelly, D. P. (1973) "Studies on the growth of Thiobacillus ferroxidans. I. Use of membrane filters and ferrous ion agar to determine viable numbers and comparison with 14CO2 -fixation and iron oxidation as measures of growth". Arch. Mikrobiol. 88: 285-298. 8. Blanchet P. (1957) "Dosage colorométrique de l'uranium par le dibenzoylméthane". Anal. Chim. Acta. 16:44-45. 9. Eva Arcidiacono, M. E. Saulnier (1979) "Estudio sobre la asociación mineral de los yacimientos y manifestaciones de Uranio del area de Sierra Pintada. San Rafael. Mendoza. Informe D. E. E. 12-79. Comisión Nacional de Energía Atómica. 10. Bosecker K., Neuschutz D., Scheffler U., (1978) "Microbial leaching of Carbonaterich German Copper-shale" Metallurgical Applications of Bacterial Leaching and related Microbiological Phenomena, ed. L. E. Murr, A. E. Torma, and J.A. Brierley (New York, Academic Press) 389-401. 11. Khalid A.M., Anwar M.A., Shemsi G., Niazi G., Akhtar K., (1993) "Biohydrometallurgy of low-grade, carbonate bearing sandstone Uranium ore". Biohydrometallurgical Technologies. Edited by A. E. Torma, J. E. Wey and, V. L. La Kshmanan. The Minerals, Metals and Materials Society. 12. Touvinen, O. (1990) "Biological fundamental of mineral leaching processes". P55-77. In Microbial Mineral Recovery: Ehrlich, H.; Brierley C. Mc Graw-Hill Publishing Company ISBN 0-07-00781-9. 252
15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Extraction of copper from mining residues and sediments by addition of rhamnolipids C.N. Mulligan and B. Dahrazma Dept. Building, Civil and Environmental Engineering, Concordia University 1455 de Maisonneuve Blvd. W., Montreal, Quebec, H3G 1M8 Abstract Rhamnolipids are anionic biosurfactants produced by the bacteria Pseudomonas aeruginosa. They are less toxic and more biodegradable than many synthetic surfactants. They were evaluated as agents to enhance removal of heavy metals from mining tailings and sediments. Mining tailings were obtained from a copper oxide mine. Another media for extraction of metals was also examined, contaminated sediments from a canal. A series of washings was performed on the mining tailings and sediments using 0.5%, 1% and 2% rhamnolipid solutions with and without 1% NaOH. The mining residue contained 8,950 mg/kg of copper and the sediments contained 140 mg/kg copper and 4,850 mg/kg zinc. Biosurfactants were added to slurry of either the mining residues or sediments with 10% mining solids. For the mining residues, with a 2% concentration of rhamnolipid, 28% of the copper could be removed. Adding 1% NaOH with the 2% rhamnolipid increased extraction up to 42% over a 6 day period. Lower concentrations of biosurfactant (0.5 and 1.0%) removed lower amounts of copper. Addition of 1% NaOH always showed higher removal rates than without the NaOH. Copper was also analyzed in the supernatants after washing the sediments. Maximal removal was 15.6% copper with 1% NaOH. For example, for a 2% rhamnolipid concentration, approximately 250 mg/kg of zinc were removed compared to 20 mg/kg of copper (0.5% rhamnolipid with NaOH). Therefore, copper can be removed from two different metal containing media by rhamnolipids. Keywords: rhamnolipids, mining tailings, sediments, copper 1.
INTRODUCTION Cadmium, copper, lead, mercury, nickel and zinc are considered the most hazardous heavy metals and are included on the EPA's list of priority pollutants [1]. Recently, the EPA has announced that the decontamination of sediments will receive the highest priority. Sources of metals include domestic and industrial effluents, the atmosphere, runoff and lithosphere. Once heavy metals are allowed to pass through the municipal waste treatment facility, they return to the environment where they are persistent, cannot be biodegraded and can thus follow a number of different pathways. The metals can adsorb onto the soil, runoff into rivers or lakes or leach in the groundwater, an important source of drinking water. Exposure to the heavy metals through ingestion or uptake of drinking water (particularly where water is reused) and foods can lead to accumulation in animals, plants and humans. 253
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Between 1850 and 1990, production of these three metals increased nearly 10-fold, with emissions rising in tandem [2]. The world mining industry has made some remarkable improvements in worker safety and cleaner production but yet it remains one of the most hazardous and environmentally damaging industries. Mining of heavy metals and of coal and other minerals is a major route of exposure since during the extraction process, a part of the heavy metals remains in the ore after discarding. Treatment methods for contaminated sediments are similar to those used for soil and include pretreatment, physical separation, thermal processes, biological decontamination, stabilization /solidification and washing [3]. However, compared to soil treatment, few remediation techniques have been commercially used for sediments. Solidification /stabilization techniques are successful but significant monitoring is required since the solidification process can be reversible. In addition, the presence of organics can reduce treatment efficiency. Vitrification is applicable for sediments but expensive. Only if a useful glass product can be sold will this process be economically viable. Thermal processes are only applicable for removal of volatile metals such as mercury and costs are high. Biological processes are under development and have the potential to be low cost. Since few low cost metal treatment processes for sediments are available, there exists significant demand for further development. Pretreatment may be one of the methods that can reduce costs by reducing the volumes of sediments that need to be treated. Methods to determine the requirements for remediation are necessary for optimal design. Biological removal of heavy metals has been the subject of several studies. Mulligan and Galvez-Cloutier [4] presented the result that heavy metals could be removed from mining residues by growth of the fungus Aspergillus niger through organic acid production They report that up to 65% of the copper can be removed from the sediments by rhamnolipids. Chartier et al. [5] could wash 51% of copper from the same sediment by biological treatment using Thiobacillus ferroxidans. Mulligan et al. [6] also used biosurfactants to remove heavy metals from oil-contaminated soil. In their study, they found that a combination of 2% rhamnolipids with 1% NaOH provided maximum removal of 25% of the copper from the soil. 1.1 Surfactants A surfactant is a "surface active agents". They are also called surface-active substances and surface-active compounds [7]. These materials are able to lower the surface tension of a solvent. Meanwhile they form aggregates, micelles, in aqueous media [8]. This property is very important, as their effectiveness depends on their ability to reduce surface tension [9]. An effective surfactant is able to reduce the air-water interface to 35 mN/m and the oil-water interfacial tension to 1 mN/m [10]. An important characteristic of surfactants is that the hydrophobic portion has little affinity for the bulk medium while the hydrophilic portion is attracted to the bulk medium. Surfactants have three important characteristics: surface tension lowering, hydrophobic-lipophilic balance or HLB, and critical micelle concentration or CMC. Low CMC values represent a more efficient surfactant since less surfactant is needed to decrease the surface tension [11]. 1.2 Biosurfactants Biosurfactants are a group of surfactants naturally produced by certain types of microorganisms. Although the low level of toxicity is a big advantage of biosurfactants over synthetic surfactants, it is not the only one. Other remarkable advantages of biosurfactants are: small molecular size, increased biodegradability, effectiveness, and ease of synthesis. Biosurfactants are made by microorganisms and this increases the 254
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possibility of in-situ production [12, 13] since they are synthesized as metabolic byproducts. The composition and yields depend on the fermentation conditions including, pH, nutrient composition, substrate, and temperature used [10]. The biosurfactant that is used in this study, a rhamnolipid, is from the glycolipids group and is made by Pseudomonas aeruginosa [7]. There are four types of rhamnolipids. Rhamnolipid type I and type II are suitable for soil washing and heavy metal removal while type III is for metal processing, leather processing, lubricants, pulp and paper processing. Type IV is usually used in textiles, cleaners, foods, inks, paints, adhesives, personal care products, agricultural adjuvants, and water treatment [14]. In the present study, the feasibility of using a biosurfactant (rhamnolipids) for the extraction of copper from a low-grade oxide mining residue and sediment were investigated. The effect of various parameters was investigated to enhance copper removal. 2.
MATERIALS AND METHODS
2.1 Ore The ore is obtained from a copper mine in the Gaspé region, Quebec, Canada. The oxide ore was already crushed into smaller particles that remained larger than 2.54 cm in nominal diameter. The ore was crushed in the lab into smaller size and sieved to find the best size for the ore particles in the extraction process. The maximum sieve size chosen was number 5 and the minimum was number 200. To remove colloidal materials from the particles, the ore was washed with tap water until no more colloids were suspended in the water. Using method ASTM D422 [15], the washed materials were placed in an oven and dried at 100°C for 24 hours. After 15 minutes in shaking sieving, the collected ores from the sieves were kept separately in a dry keeper. The method recommended by Environment Canada [16], for digestion of ores containing copper, was used to digest the prepared ore. The copper in the ore was measured by an Atomic Absorption Spectrophotometer (Perkin Elmer Aanalyst 100). The results are shown in Table 1. 2.2 Chacterization of the sediment sample The metal-contaminated sediment sample was obtained from a canal area which was surrounded by metal and steel industries. The sample was air-dried. The grain size distribution of the sediment indicated 10% sand, 70% silt and 20% clay. X-ray analysis indicated the presence of quartz (30%), feldspar (36%), illite (2%), kaolinite (27%), chlorite (3%) and carbonate (0.5%) as performed by Mulligan et al. [4]. Total organic matter was 20% (w/w). The sediments were digested by the method recommended by Environment Canada [16] and then analysed by a Perkin Elmer Atomic Absorption Aanalyst 100 Spectrophotometer for heavy metal content. These results are summarized in Table 1. Table 1. Characterization of sediment sample for heavy metal content Parameter Chromium Copper Nickel Lead Zinc
Concentration (mg/kg) in mining residues ND 8,950 27 ND 201
Concentration (mg/kg) in sediments 145 140 76 572 4,854 255
Bioleaching Applications
2.3 Biosurfactants The rhamnolipids, used in this study, were biosurfactants type I and type II from the glycolipid group made by Pseudomonas aeruginosa with the trademark JBR215 from Jeneil Biosurfactant Co. JBR215 is an aqueous solution of rhamnolipid at 15% concentration. It is produced from a sterilized and centrifuged fermentation broth. Two major types of rhamnolipids, RLL (R1) and RRLL (R2), are present in the solution. Several tests done by the manufacturer and independent laboratories show the degree of biodegradability and toxicity of JBR215 match the EPA requirements. The CMC was found to be 0.035 g/L through conductivity measurement at various dilutions. This value is equivalent to 0.003% rhamnolipid. Therefore, for all experiments, a concentration above the CMC was used to ensure the formation of micelles. 2.4 Washing of mining residue To evaluate the effect of various parameters on the extraction of copper from the lowgrade mining residue as well as to optimize extraction, several tests were performed. All tests were duplicated and the difference between tests never exceeded more than 5%. Parameters were optimized step by step. Since the results of each step were used for the next parameter, tests were repeated between 2 to 8 times. All of them showed reasonable agreement. The ore was placed in batch reactors for copper extraction. In some cases samples were placed on a rotary shaker for a desired period of time. For analysis, the ore particles were allowed to settle. The supernatant solution was then decanted from the ore and digested according to the procedure of APHA [17], Method 3030E. To release the biosurfactant from the copper, 30% solution of hydrogen peroxide was added until no reaction was observed. The following steps were then followed: • • • •
50 mL of the sample was transferred to a 125 mL beaker 25 mL of concentrated nitric acid and a few boiling chips were added to the beaker The beaker was then heated on a hot plate and brought to a slow boil. Once the solution became a clear light-colored liquid, the digestion was considered complete • The solution was cooled to room temperature, filtered through filter #40 ashless paper, and the volume was readjusted to 50 mL with distilled water. The quantity of copper was measured by an Atomic Absorption Spectrophotometer. 2.5 Washing of the sediment with the biosurfactants The biosurfactant solution was diluted with distilled water or 1% NaOH as required. A quantity of 1.5g from the sediment of each particle size was placed in individual 50mL vials. 15.0mL of surfactant solution was added to each vial. The samples were kept in the temperature incubator at constant temperature (25°C) for 1 day. Blank samples including 15.0mL distilled water and 1.5g sediment were provided. Vials were placed on the side in order to achieve the maximum contact surface between the ore particles and the biosurfactant. The pH of the rhamnolipid with NaOH was 13 and without NaOH was 6.0. Blanks included the same additives as for the biosurfactant studies without the presence of the biosurfactant. Washing of the supernatants was performed for 6 days, and then removing the supernatants (3,000 x g, 30 min). All experiments were performed in triplicate. Results were reproducible to ±10%. The solutions from each sample were collected. The samples were digested and the concentrations of copper and other metals in each sample were measured by the Atomic Absorption Spectrophotometer In order to 256
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release the heavy metals trapped in the biosurfactant micelles, the organic copper colloids were oxidized by adding 30% H2O2 slowly until no reaction was observed. The method chosen for biosurfactant digestion was 3030E, which is recommended by APHA [17] and approved by the EPA. 3.
RESULTS AND DISCUSSION
3.1 Mining residue studies 3.1.1 Effect of particle size Seven ore samples with different sizes were collected from sieves and 1g of each was placed in individual vials and 10mL of the rhamnolipid solution of 1% (pH 6.5) was added. The control consisted of the same ore with distilled water. The test was repeated while the samples were placed on a shaker. The shaker was set at 100 rpm and after 5 days, the solutions were digested and analysed. A comparative diagram in Fig. 1 is presented. Controls were not shown since they were not significant. Since shaking did not remarkably improve the extraction, the experiments were continued without shaking. The results show that a maximum of 17.7% copper extraction occurs in the unshaken samples containing ore particles between 0.15mm and 0.3mm. This particle size was chosen for the rest of the study. 20%
Extraction of Copper
16%
12% w ith o u t s h a k in g 8%
w ith s h a k in g
4%
0% 0 0 9 4 9 0 5 .0 .0 .1 .8 .5 .3 .1 -4 -2 -1 -0 -0 -0 -0 0 9 4 9 0 5 7 0 1 8 5 3 1 0 2. 1. 0. 0. 0. 0. 0. M e s h s iz e ( m m )
Figure 1. Comparative diagram of copper extraction for the chosen particle sizes for shaken and unshaken samples 3.1.2 Effect of pH on copper extraction Using the data from the first test, one gram of ore with particle sizes between 0.15mm and 0.3mm was placed in the vials with 10mL of 1% rhamnolipid for 5 days at various pH values. The pH was adjusted to between 6 and 9.5 and the extraction efficiencies were found by the measurement of the copper concentration in the solutions. Fig. 2 shows how the pH affects extraction of copper from ore. The minimum extraction of copper occurred 257
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when the pH was around 8.8. This is similar to information in the solubility curve of copper hydroxide of Radhakrishnan [18] which show that minimum solubility is at pH 8.7. This can explain the occurrence of minimum extraction at the same pH. 20%
Copper Extraction
16%
12%
8%
4%
0% 6
6.5
7.5
8.5
9.5
pH
Figure 2. Variation of copper extraction with pH. Particles were between 0.15 to 0.3 mm in diameter. Each sample contains 1 g ore in 10 mL solution of 1% rhamnolipid. The samples were kept at 25°C for 5 days and left without shaking 3.1.3 Effect of the concentration of biosurfactant and NaOH The effect of the concentration of rhamnolipid on copper extraction was determined by using 0.05% to 5.0% of rhamnolipid. Higher concentrations were not tested since the solution would be considerably viscous and hard to work with. According to previous studies [6], adding 1% NaOH to the solution will improve the copper extraction process. To verify the effect of NaOH, another test was designed in which the same concentrations were used but with co-addition of 1% NaOH. A comparative graph is presented in Fig. 3. Controls are not shown since the controls did not demonstrate significant extraction. As the figure shows, the effect of NaOH is quite remarkable on copper extraction for concentrations of 2% rhamnolipid or less while it has a negative effect on the extraction for more concentrated rhamnolipid. In the later case, the extraction of copper in the solution of 1% NaOH and 2% rhamnolipid has almost the same value as of 4% rhamnolipid without NaOH. The added NaOH increases the pH up to 13.5. 3.2 Sediment washing studies A series of washings was performed on the sediments using 0.5%, 1% and 2% rhamnolipid solutions with and without 1% NaOH. The control was 1% NaOH. Copper was analyzed in the supernatants and is plotted in Fig. 4. Removal rates for copper were highest for 0.5% rhamnolipid with 1% NaOH. Increasing the biosurfactant concentration when NaOH was added did not enhance removal rates. Maximal removals were 15.6% copper. For example, for a 2% rhamnolipid concentration, approximately 20 mg/kg of copper (0.5% rhamnolipid with NaOH). Zinc removal on the other hand was higher without 1% NaOH and was maximal at 2% rhamnolipid. Although zinc removal at 5% does not seem significant, the actual amount of zinc is higher than for copper at 250 mg/kg of zinc removed. This could be one of the reasons why the removal for copper by the rhamnolipid is lower for the sediments than for the mining residues. In other words, there 258
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seems to be some preference for the certain heavy metals by the biosurfactant in the sediment. Another reason may also be that the organic content in the sediment is high and may lead to some adsorption of the biosurfactant. 70%
60%
Copper Extraction
50%
40%
30%
20%
10%
0% 0 .0 5
0 .1 0 .2 0 .4 0 .5 1 2 C o n c e n tr a tio n o f S u r fa c ta n t (% ) W ith o u t N a O H W ith 1 % N a O H
4
5
20 15
Cu
10
Zn
5
Su Su rf rf+ 1% 1% N aO Su H rf+ 1% 2% N aO Su H rf +1 % Na O H 1% Na O H
0. 5%
2%
Su rf 1%
0. 5%
Su rf
0
N on e
Metal removal (%)
Figure 3. The effect of 1% NaOH on the extraction process for various concentrations of rhamnolipid. Particles were between 0.15 to 0.3 mm in diameters. Each sample contains 1 g ore in 10 mL of rhamnolipid solution while the pH was adjusted to 6. The samples were kept at 25°C through the test and left without shaking for 6 days.
Washing agent
Figure 4. Percentage of copper and zinc removed from sediments by various washing agents 4.
CONCLUSIONS This research was performed to evaluate the application of rhamnolipid on extraction of copper from mining residue. The ore initially has a low concentration of copper that cannot be economically removed. From an environmental point of view, the existing copper in the ore is a heavy metal contaminant and should be removed so that it will not impact the environment. 259
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This study dealt with several chemical and physical factors involved in the extraction to find the best conditions for the extraction. These parameters were ore particle size, pH, concentration of biosurfactant, effect of NaOH addition, and shaking. The importance of particle size was in the contact surface area between the ore particle and the biosurfactant. Rhamnolipids were also evaluated for removing contaminants from sediments. Results were not as good as for the mining residues possibly due to competition from the elevated levels of zinc in the sediments or the high levels of organic matter in the sediments. ACKNOWLEDGEMENTS The authors would like to acknowledge the financial support from NSERC for this research. REFERENCES
1. R.E. Cameron, Guide to Site and Soil Description for Hazardous Waste Site Characterization. Volume 1: Metals. Environmental Protection Agency EPA(1992) 2. WRI Heavy Metals and Health. World Resource Institute. http:// www.wri.org/wr-9899/metals.htm (1999). 3. C.N. Mulligan, R.N. Yong and B.F. Gibbs, Eng. Geol. 60 (2001)193. 4. C.N Mulligan and R. Galvez-Cloutier, Water Science and Technology 41(2000) 255. 5. M. Chartier G., Mercier and J. F. Blais, Wat. Res. 35 (2001) 1435. 6. C.N. Mulligan, R.N. Yong, and B.F. Gibbs, Environ. Prog. 18 (1999) 50. 7. K. Tsujii, Surface Activity, Principles, Phenomena, and Applications. Academic Press, USA, 1998. 8. D. Myers, Surfaces, Interfaces, and Colloids. Principles and Applications. 2nd edition, Wiley, New York, NY, 1999. 9. M. J. Rosen, Surfactants and Interfacial Phenomena. John Wiley & Sons, New York, NY, 1978 10. C.N. Mulligan, and B. F. Gibbs, Biosurfactants: Production, Properties, and Application. Vol. 48 Kosaric N., ed., Marcel Dekker, New York, 447 (1993). 11. P. Becher, Emulsions, Theory and Practice. 2nd edition, Reinhold Publishing, NewYork, 1965. 12. R. Muller-Hurtig, F Wagner, R. Blaszezyk and N Kosaric. Biosurfactants: Production, Properties, and Application. Vol. 48, Kosaric, N. ed., Marcel Dekker, New York, 447469d (1993). 13. L. Ju, Efficiency of biosurfactants applied by means of electrokinetics. Ph.D.Thesis, Concordia University, Montreal, QC, Canada (1999). 14. Jeneil Biosurfactant Co. JBR215, Product Data Sheet (2001). 15. ASTM “Special procedure for testing soil and rocks for engineering purposes” D42254. American Society of Testing and Materials, West Conshohocken, PA, 101-103, 1970. 16. Environnement Canada, Inventaire national des rejets de pollutants. Rapport regional, Région du Québec, Direction de la protection de l’environnement, Section Enjeux atmosphériques et substances toxiques du Québec, Octobre (1996). 17. APHA, American Public Health Association, American Water Works Association, Water Pollution Control Federation Standard Methods for the Examination of Water and Wastewater, APHA, Washington, DC, 1989. 18. E. Radhakrishnan, Recovery of Metals from Sludges and Wastewater. Noyes Data Corp., Park Ridge, NJ, 1993. 260
15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Improving of film coating bioleaching using biorotor process A.R. Shahverdia, M. Oliazadehb, R. Rohia and M. Davodib a
Department of Biotechnology, Faculty of Pharmacy* and b Department of Mining, Faculty of Engineering, Tehran University Abstract A sample of gold bearing pyrite concentrate from Mouteh plant, Iran was partially leached in a biorotor at a level of ten percent, then it was used for coating the crushed ore. Coated particles were subjected to thirty days column bioleaching with a culture of Thiobacillus ferrooxidans named Tsho. This study showed that pyrite biooxidation could be increased about two times. Furthermore by using the pretreated concentrates the lag phase period decreased dramatically. This enhanced biooxidation caused the gold recovery by cyanidation to be increased of about seven percent. Keywords: biorotor, GeoBiotics process, pyrite, Thiobacillus ferrooxidans 1.
INTRODUCTION Film coating bioleaching (developed by GeoBiotics, Inc.) is an innovative and economic process that has been designed to sum the advantages of biooxidation in stirred tank reactors and heap leaching. At the same time this system is aimed at overcoming several drawbacks of both processes. This process is economical with low-grade concentrates and with mineral sulfides such as pyrite, whose biooxidation is slow [1]. This work presents results of column testing of the Mouteh gold concentrate coated on a support-crushed rock. Furthermore it was found that by using a pretreated concentrate the efficiency of film coating bioleaching could be considerably increased. The rotating drum was selected for preliminary biooxidation of the concentrate because this type of bioreactor can be operated at much higher pulp densities than conventional stirred tank reactors [2].
2.
MATERIALS AND METHODS
2.1 Mineral sample A representative sample of pyritic gold concentrate from Mouteh gold plant, Isfahan, Iran, was used in the bioleaching experiments. Table 1 shows the primary mineralogical species and the chemical composition of the concentrate. Particle size analysis revealed that 85 percent of the sample was finer than 150 microns.
* Correspondence address: Department of Biotechnology, Faculty of Pharmacy, Tehran University of Medical Sciences. Fax: (98) 21-6461178, E-mail:
[email protected]
261
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Table 1. Mineral and chemical composition of the Mouteh pyrite gold concentrate used in the bioleaching tests Minerals Identified
Wt. %
Chemical Composition
Wt. %
Pyrite Chalcopyrite Covellite Tennantite Sphalerite Galena
68.4 Trace Trace
Sulfide sulfur Silicate Aluminum Magnesium Iron Gold Calcium
30.09 24.46 3.32 0.55 32.08 30ppm 1.19
2.2 Basal medium HP medium with the following composition was used for the bioleaching experiments: 0.4 g/L (NH4)2SO4, 0.4 g/L MgSO4.7H2O and 0.1g /L K2HPO4. The solution pH was adjusted with H2SO4 [3]. 2.3 Microorganism The strain used in the experiments was Thiobacillus ferrooxidans Tsho [4]. The strain was previously adapted to HP medium containing 3% (by weight) concentrate at 30°C. 2.4 Biological pretreatment of concentrate The mineral sample was partially oxidized at a level of ten percent in a Plexiglas biorotor [2]. For monitoring the process in the biorotor, the pulp was sampled every 24 hours, and the samples were analyzed for pH, ferrous iron, total soluble iron, total iron. The samples were decanted and the solids were returned to the reactor. The volume of the pulp was maintained constant by periodical additions of acidified distilled water. During the leaching process, some portion of the released iron was precipitated. Therefore, total iron was measured after acid digestion with 6N HCl for 30 min at 65°C. The metal content was analyzed by titrimetric method using 0.06 N K2Cr2O7. Also the total number of cells in solution was analyzed by Neubauer improved counting chamber. 2.5 Column leaching of coated support rocks The concentrate was mixed with the bacterial inocula so as to form thick slurry, which was then sprayed onto a sized supporting rock 1 x 2.5 cm diameter (20% by weight of the support rock). The support rock was a refractory sulfide ore that had been crushed in a jaw crusher. The thickness of concentrate coating layer was 0.5 to 1 mm. The tests were performed at 20°C in a column 9 cm in diameter by 40 cm in height. The feed solution at pH 1.7 was applied to the column through glass wool pad at various flow rates. Sampling of the effluent solutions was performed every 4 days, and the samples were analyzed for pH, Eh, ferrous iron and total iron. Also, measurements of residual nutrients were performed periodically. Gold recovery, before and after biooxidation, was determined by conventional cyanide roll bottle tests. 3.
RESULTS AND DISCUSSION Figure 1 shows the effect of the different flow rates on the biooxidation of coated material. The experiment was carried out at room temperature and pH 1.7. Flow rate is an important factor in column leaching of coated particles, since it can affect the
262
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30
14 Q=4 (lit/m^2/h)
25
Q=8 (lit/m^2/h)
20
Q=9.5 (lit/m^2/h)
15 10 5 0
4 12 20 28 36 44 52 60 68 76 84 92
Time (days)
Oxidation Level %
Cumulative TSI (g/lit)
characteristics of the feed materials. The results showed that a flow rate of 4 dm3.m-2.h was not sufficient, compared to the higher flow rates, to oxidize the sample. Increasing the flow rate to 8 dm3.m-2.h significantly increased the oxidation rate. Higher flow rate increments did not improve the oxidation rate. Figure 2 shows the comparison of the biooxidation rate of the pre-treated sulfide concentrate with that of the untreated concentrate in column leaching. Pre-treatment of the concentrate resulted in increasing the oxidation of the concentrate more than twice after 28 days. The lag phase of bacterial growth has also been considerably decreased. This improvement of biooxidation produced an increase of about 7 per cent in recovery of gold by cyanidation.
12
Pyrite concentrate
10
Partially oxidized pyrite concentrate
8 6 4 2 0 4
8
12
16
20
24
28
32
Time (days)
Figure 1. Column bioleaching operation of pyrite coated particles at various flow rates
Figure 2. The effect of biologically pretreatment of concentrate on the efficiency of column bioleaching
REFERENCES
1. D. E. Rawlings (ed.), Biomining: Theory, Microbes and Industrial Process, SpringerVerlag and Landes Bioscience, Chapter 6, 117-127, 1997. 2. Vargas, T., Jerez, C.A., Wiertz, J.V. and Toledo, H. (eds.), Biohydrometallurgical Processing, University of Chile, Santiago, Chile, Vol. 1, 263-271, 1995. 3. H.M. Liazma and I. Suzuki, Biotech. and Bioeng., 32 (1988) 110. 4. A.R.Shahverdi, M. T. Yazdi, M. Oliazadeh and M.H. Darabidi, J.Sci.I.R.Iran, 12 (2001) 209. 5. M.N. Herrera, B. Escobar, N. Parra, C. Gorizalez and T.Vargas, Minerals and Materials Processing, 15(1998) 15.
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Isolation and evaluation of indigenous iron- and sulphuroxidising bacteria for heavy metal removal from sewage sludge R. Matlakowska and A. Sklodowska Laboratory of Environmental Pollution Analysis, Faculty of Biology, Warsaw University, CEMERA - Centre of Excellence, Miecznikowa 1, 02-096 Warsaw, Poland E-mail:
[email protected] Abstract The bioleaching of copper and cadmium from sewage sludge contaminated with heavy metals (obtained from a wastewater treatment plant in a large agglomeration) was optimised using indigenous bacteria. The enrichment and adaptation of naturally existing microflora was achieved after sludge supplementation with elemental sulphur (ferrous sulphate was added as a flocculent to crude wastewater), intensive aeration and four successive transfers. The application of enriched and adapted bacteria resulted in sludge acidification to pH 1.9 after 10 days of cultivation without the use of inorganic acid. The maximal extraction of copper and cadmium achieved in a laboratory-scale experiment was 70% and 100%, respectively. Two groups of sulphur and iron-oxidising bacteria were isolated from the sewage sludge using Starkey and 9K media. The morphology and physiology of the isolated bacteria grown in mineral medium as well as in sterilised sewage sludge were investigated. Examination of the isolates by scanning electron microscopy indicated the presence of rod–shaped bacteria and revealed a characteristic leaching pattern on sections of copper sulphide ore. The morphological and ultrastructural differences between cells grown in mineral medium and in the sewage sludge were clearly visible. Keywords: sewage sludge, iron- and sulphur-oxidising bacteria, TEM, SEM 1.
INTRODUCTION The growth of many species of chemolithotrophic bacteria is inhibited by some organic compounds. According to Tutle and Dugan [1] the acidophilic leaching bacteria Acidithiobacillus ferrooxidans and Acidithiobacillus thiooxidans are the most strongly affected by the presence of organic substances in a mineral medium. The iron- and sulphur-oxidising bacteria, however, are very often isolated from environments containing organic matter. According to Zagury et. al. [2] iron-oxidising bacteria are present in contaminated soil and can be easily enriched and adapted by supplementing the soil with ferrous sulphate. Additionally, the high pH of soil, high organic content, oil and grease does not inhibit the growth of those microorganisms [2]. Iron- and sulphur-oxidising bacteria identified as A. ferrooxidans and A. thiooxidans, were also successfully enriched by Gomez and Bosecker [3] from various soil samples contaminated with heavy metals 265
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and organic matter. The adapted bacteria were used for heavy metal removal from soil and sediments. Similarly, iron-oxidising bacteria were detected by Tyagi et. al. [4] in different types of sewage sludge and can be also adapted by supplementing the sludge with ferrous sulphate. Blaise et. al. [5] characterised the naturally occurring microorganisms responsible for metal leaching activity in different sewage sludges. According to them the initial acidification of the sludge in the bioleaching process is brought about by the growth of indigenous less-acidophilic thiobacilli, followed by the growth of acidophilic thiobacilli, resulting in pH reduction to approximately 2.0. The successive growth of lessacidophilic and acidophilic bacteria was observed in different types of sludge and with varying sludge solid concentration. Taking into account the chemical composition, the existence of chemolithotrophs in sewage sludge seems to be a very interesting phenomenon. The sewage sludge as an end product generated in a wastewater treatment plant besides heavy metals, accumulates many chemical substances that are not fully degraded during this treatment. The spectrum of xenobiotic organic compounds in sewage sludge is extremely wide and constantly changing. Typical representatives of those substances detected in sewage sludge are polycyclic aromatic hydrocarbons, polychlorinated biphenyls and chlorinated pesticides [6]. The influence of organic matter on speciation of heavy metals in sewage sludge is well known. However, the effect of these substances on microorganisms used in the bioleaching process is still unclear. The scope of our paper was the evaluation of indigenous bacteria for copper and cadmium bioleaching from sewage sludge. The possibility of enrichment and adaptation of naturally existing microflora was estimated and the growth, protein profile, morphology and ultrastructure of isolated bacterial cells grown in mineral medium and in sewage sludge were compared. 2.
MATERIALS AND METHODS
2.1 Samples Anaerobically digested and dehydrated sewage sludge was obtained from a wastewater treatment plant in Warsaw. 2.2 Enrichment and adaptation The process was carried out in 5.0-litre flask containing 500g of unsterilised, dehydrated, anaerobically digested sludge suspended in 2000 ml of mineral medium (9K medium without energy source). The sludge was supplemented with 10 g of elemental sulphur per litre. The cultures were aerated with pressurised air and were maintained at room temperature. 200 ml of culture was transferred to fresh sewage sludge after 10 days and cultivated under the same conditions. The experiment was repeated successively four times (series I, II, III, IV). 2.3 Evaluation for heavy metal bioleaching The pH of sludge in series I, II, III and IV was estimated every day. The cadmium and copper concentration in sewage sludge was determined in the beginning and at the end of every series. The metal was analysed with flame atomic absorption spectrometry according to FAAS protocol after acid digestion. 266
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2.4 Isolation The sludge obtained in series IV was used for the isolation of two groups of iron- and sulphur-oxidising bacteria. Media 9K [7] and Starkey [8] were used. 125 ml of sterile medium was inoculated with 25 ml of sewage sludge. The culture was cultivated on a rotary shaker at room temperature. 2.5 Growth characterisation The isolated iron- and sulphur-oxidising bacteria were cultivated in 9K mineral medium and in Starkey medium, respectively. Additionally, after six months of cultivation in mineral medium, the growth of isolates was tested in sterilised sewage sludge supplemented with 9K or Starkey medium without an energy source. The sludge was preacidified to pH of 2,5 (9K) or 4,0 (Starkey). The cultures were performed in 500 ml flask containing 50 mg of sludge and 200 ml of mineral medium. The cultures were incubated on rotary shaker at room temperature. The number of bacterial cells in a population was assessed by counting the cells in a sample using a counting chamber under a microscope. 2.6 Preparation of samples for SEM and TEM Thin polished sections of naturally occurring copper sulphide ore were introduced into actively growing bacterial cultures according to the method described by Ostrowski and Sklodowska [9]. Samples were incubated at room temperature without stirring for two months. After the bioleaching process thin sections were carefully rinsed with water, allowed to dry at room temperature and fixed with 3% glutaraldehyde. They were dehydrated in the increasing concentration of ethyl alcohol and propylene oxide. Samples were coated with gold and examined under a scanning electron microscope. For TEM cells were fixed with 3% glutaraldehyde in sodium cacodylate buffer and then treated with osmium tetroxide for 4h. An increasing concentration of the ethanol to 100% was used for dehydration. Ultrathin sections of epon-embedded cells were treated with uranium acetate and lead citrate. 2.7 Extraction of protein fraction The method of Neu and Heppel [10] was used for isolation of periplasmic proteins. Cytoplasmic and membrane proteins were isolated according to method of Witholt [11]. Proteins were isolated from bacterial cells cultivated in mineral medium and sewage sludge according to the description in section 2.5. 2.8 Protein analysis SDS-PAGE electrophoresis was performed on 8-25% polyacrylamide gradient gel using Phast System equipment (Amersham Pharmacia Biotech). Gels were stained with silver and analysed with ImageMaster 1D Elite (NonLinear Dynamics). The molecular mass of proteins was determined using DP-Soft (analySIS) software (Soft Imaging Systems for Olympus) according to a high (not presented) and low molecular weight calibration kit (Amersham Pharmacia Biotech). 3.
RESULTS AND DISCUSSION A mixture of indigenous bacteria, containing autotrophic as well as heterotrophic acidophili, was used for bioleaching of copper and cadmium from anaerobically digested and dehydrated sludge. The enrichment and adaptation of natural microflora was achieved 267
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pH
after the sludge supplementation with elemental sulphur as the energy source and after four successive subcultures (series I, II, III, IV). The sludge acidification in the course of bioleaching is presented in Fig. 1. The enrichment and adaptation of bacteria contributed to gradual sludge acidification. Finally, a decrease in pH to about 1.9 after 10 days of cultivation was observed in series IV. 8 7 6 5 4 3 2 1 0 0
1
2
3
4
7
8
9
Time (days) I
II
III
IV
Figure 1. The sludge acidification in the course of bioleaching process
The efficiency of copper and cadmium bioleaching in the successive series is presented in Table 1. The concentration of removed copper and cadmium was the lowest in series I and gradually increased in the following subcultures. The maximal extraction of cadmium was 100% and was obtained in series III and IV. Copper bioleaching was the highest in series III (70%). Table 1. Cadmium and copper bioleaching efficiency (%) Metal Cadmium Copper
Series I 36 8
II 6 14
III 100 70
IV 100 42
In the next part of the experiment a mixture of enriched and adapted indigenous bacteria (obtained in series IV) was used for isolation of autotrophic bacteria. The two groups of microorganisms were separated depending on the energy source. The ironoxidising bacteria were isolated in 9K medium with ferrous sulphate as the energy source. The Starkey medium with elemental sulphur was used for the isolation of sulphuroxidising bacteria. The existence of iron-oxidising bacteria in sewage sludge was explained by a high concentration of ferrous sulphate added as a flocculent to crude wastewater. As indicated later the isolated iron-oxidising bacteria were also able to oxidize sulphur. The growth of isolates was screened by acid production and cell number. Growth curves of bacteria (Fig. 2A, B) and pH decrease (data not presented) confirmed that the isolates were capable of lithotrophical growth with ferrous iron or sulphur as the energy source under laboratory conditions. In the culture of iron-oxidising bacteria in 9K medium the pH decreased to about 1.9 during 6 days. Liquid medium with ferrous sulphate changed from green to red-brown with ferric sulphate. At pH close to 2.0 precipitation and encrustation of jarosites was observed. 268
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Cells per ml (x106)
In Starkey medium the elemental sulphur was oxidised by isolated strains resulting in pH decrease to 1.25 after 14 days of cultivation. This process was also confirmed by the appearance of colloidal sulphur. The ability of the isolated bacteria to re-growth (after six months of cultivation in mineral medium) in sterilised sewage sludge without any additional energy sources was also proved (Fig. 2A, B). A 60 50 40 30 20 10 0 0 1 2 3 4 5 6 7 8 9 10 11 12 13 14 Time (days) 9K
Sludge
Cells per ml (x106)
B 60 50 40 30 20 10 0 0 1 2 3 4 5 6 7 8 9 10 11 12 13 14 Time (days) Starkey
Sludge
Figure 2. Growth of indigenous iron- (A) and sulphur-oxidising (B) bacteria in mineral medium and sewage sludge
The comparison of growth curves of two groups of bacteria grown in mineral medium or in sewage sludge reveals differences, especially in the beginning of cultivation. A lag phase of the culture of iron- as well as sulphur-oxidising bacteria grown in the sewage sludge was observed which lasted about 4-5 days. Exponential growth of bacteria cultivated in the mineral medium begins immediately after inoculation. In all cases an exponentially growing culture was inoculated into the same medium under the same growth conditions. As seen on electron micrographs of ultrathin sections of bacteria grown in sewage sludge (Fig. 3) many cells are completely destroyed, membranes and leaky spheroplasts are also observed. The lag phase could be also connected with the inhibitory effect of chemical substances present in the sewage sludge. According to Tutle and Dugan [1] the chemolithotrophic growth of some strains of A. ferrooxidans is inhibited by a wide variety of organic substances. The growth and particularly iron or sulphur oxidation is inhibited by yeast extracts, peptones, amino acids, carbohydrates, carboxylic acids, anionic detergents, cationic surfactants and ammonium compounds. Some of these compounds (carboxylic acids) were completely inhibitory at 269
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the concentration used, others (malate, succinate, fumarate) caused an extended lag phase [1, 12]. Finally, after 10 days the number of bacterial cells in the sewage sludge was the same (iron-oxidising bacteria) or even higher (sulphur-oxidising bacteria) than in the mineral medium. The chemical analysis of sewage sludge used with GC-MS showed the presence of such components as organic acids esters (benzodicarboxylic, hexanoic, propenoic), alcohols, benzene and derivatives, polycyclic aromatic hydrocarbons (phenanthrene, anthracene, naphthalene), phenols, aliphatic saturated and unsaturated hydrocarbons and terpens (data unpublished). The concentration of Cd, Cu and Pb in the sewage sludge was 12, 476 and 117 mg/kg d.w., respectively. Microscopic observation of isolated bacteria revealed the presence of single, rod– shaped, Gram-negative bacteria. No evidence for Leptospirillum ferrooxidans in the culture of iron-oxidising bacteria was showed on microscopic figures. Preliminary identification of isolates with FISH method using probe Thio820 specific for A. ferrooxidans and A. thiooxidans [13] revealed that isolated bacteria belong to the genus Acidithiobacillus (unpublished results).
Figure 3. Transmission electron micrographs of ultrathin section of sample containing cells, membrane envelopes and leaky spheroplasts of sulphur-oxidising bacteria cultivated in sewage sludge. Bar, 1µm
The bacterial cells immobilised on the surface of the ore section introduced into growing bacterial culture were observed under a scanning electron microscope. Indirectly, the ability of isolates to oxidize copper sulphide in ores was shown. Deep corrosion sponge-like pits after two months of the bioleaching process were visible (Fig. 4 A, B, C). The destruction of the ore was more visible in the case of the culture in the mineral medium than in the sewage sludge. The bacterial cells attached to the surface of the section and in the pits were visible at large magnifications (Fig. 5A, B, C). The surface of ore and bacterial cells was covered by a mucosal biofilm (Fig. 5B). However more colloidal and crystal-like particles were visible in sections from the mineral medium (Fig. 5A, C).
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A
B
C Figure 4 A, B, C. Corrosion of ore after two months of bioleaching with sulphur-oxidising bacteria in Starkey medium (A) and sewage sludge (B, C).
A
C
B
Figure 5 A, B, C. Iron-oxidising bacteria attached to surface of ore during bioleaching in 9K medium (A, C) and sewage sludge (B).
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A
D
B
E
C
F
Figure 6. Transmission electron micrographs of ultrathin sections of iron-oxidising bacteria cultivated in 9K medium (A, B, C) and in sewage sludge (D, E, F). Bar, 0.5 µm.
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Cross and longitudinal ultrathin sections of isolated bacteria were examined using a transmission electron microscope. The morphological and ultrastructural differences between cells grown in mineral medium and in the sewage sludge were clearly visible. Cells of iron- and sulphur-oxidising bacteria grown in mineral media showed typical shape and ultrastructure. Autotrophically-grown cells possessed a multilayered cell envelope with a cytoplasm containing a dispersed nucleus, ribosomes and the inclusion of polygonal profiles (carboxysomes) (Fig. 6A, B, C, 7A, B). In the thin section the carboxysomes have a granular structure of medium electron density and there are several (3-5) bodies per cell, which is shown in Fig. 6B and 7B. A
C
B
D
Figure 7. Transmission electron micrographs of ultrathin sections of sulphuroxidising bacteria cultivated in Starkey medium (A, B) and in sewage sludge (C, D). Bar, 0.5 µm
Sewage sludge grown cells were embedded with slime, slightly deformed and contained reserve material (Fig. 6D, E, F, 7C, D). It is likely that the electron transparent areas seen in the thin section are deposits of poly-β-hydroxybutyrate (PHB) (Fig. 6E, F). It is known that the polymer accumulates when carbon and energy sources are accessible and according to Shively [14] PHB is considered to be a cellular reserve of energy or carbon and energy. Tabita and Lungren [15] indicated that the growth of A. ferrooxidans in mineral medium with glucose resulted in poly-β-hydroxybutyrate accumulation. The section of iron- as well as sulphur-oxidising bacterial cells grown in sewage sludge also showed more peripherally localised electron dense bodies than cells grown in mineral 273
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medium (Fig. 6E, 7C]. Regular hexagons, with a solid, granular interior may be regarded as polyphosphate granules. Additionally, in either case, the nucleoplasm of cells cultivated in sewage sludge had a characteristic consistence. It seemed to be very dense and tightly arranged. Carboxysomes did not occur. As mentioned earlier cell envelope disruption and partial leakage of cellular material in samples of bacteria cultivated in sewage sludge were visible (Fig. 3). The comparison of the size of undistorted cells grown under two conditions revealed that the cells grown under autotrophic conditions (9K or Starkey) were slightly smaller. The deformation of cells of A. ferrooxidans after treatment with organic acids was reported earlier. Electron micrographs showed blebbing and waviness of the cell envelope and void spaces inside the cell. According to Tutle et. al. [16] the cell envelope is the site of the toxic action of organic acids. It was postulated that organic acids disrupt the cell envelope by dissolving in it and by complexing with cations that maintain its integrity [16]. Organic compounds in the growth medium had also a detectable effect on the ultrastructure of another chemolithotrophic bacterium, Nitrobacter agilis [17]. Most cells grown under heterotrophic conditions were irregular in shape, distorted and filled with PHB. Pope et. al. [17] reported that the cells were larger than cells grown under strictly autotrophic conditions. The comparative analysis of proteins extracted from iron- and sulphur-oxidising bacteria grown in mineral medium and in sewage sludge was performed. Preliminary protein profile characterisation showed visible differences. Fractions of iron-, as well as sulphur-oxidising bacteria, grown in sewage sludge were characterised by small number of protein bands. This phenomenon was observed in all analysed fractions. Additionally, in the case of cells grown in the sewage sludge difficulties in isolating and separating the proteins on the gel appeared. Unseparated bands were obtained in a few cases although the same methods of isolation and separation were used. Analysis of proteins synthesised by bacteria grown in sewage sludge revealed that only a few of them remained unchanged in comparison with proteins synthesised by bacteria cultivated in mineral medium. Generally, many proteins were completely absent or expressed in different amount. Only few new proteins appeared in sewage sludgecultivated bacteria. Iron-oxidising bacteria grown in sewage sludge (Fig. 8A) synthesised two proteins present only in the membrane fraction (36, 32.2 kDa) and three (15.2, 29.9, 68.3 kDa) in the cytoplasmic fraction. Two other proteins (56.6, 69.2 kDa) isolated from the periplasmic fraction were expressed at a higher concentration than in bacteria cultivated in mineral medium. In the case of sulphur-oxidising bacteria (Fig. 8B) grown in sewage sludge specific results were obtained. Only one new protein of molecular weight of approximately 81 kDa was identified in the periplasmic fraction and one (19.4 kDa) in the cytoplasmic fraction. There was also only one unseparated protein band in the membrane fraction detected. In the literature, many examples of adaptation of Gram-negative microorganisms to different environmental conditions can be found. For example, A. ferroxidans responds and adapts to changes in its external mining environment by the synthesis of an outer membrane protein [18, 19]. The 40-kDa protein (omp40) was found, whose synthesis is regulated by external pH and concentration of phosphorus in medium.
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Figure 8. Patterns of protein synthesis by iron- (A) and sulphur-oxidising bacteria (B) cultivated in mineral medium (9K or Starkey, St) and in sewage sludge (SS). Periplasmic (PP), membrane (MP) and cytoplasmic (CP) protein were analysed using SDS-PAGE. LMW – low molecular weight protein standards.
Taira et. al. [20] suggested that the enhanced resistance to cadmium in A. ferrooxidans is connected with the synthesis of some protein. The presence of cadmium in the growth medium also indicated several changes in the cytoplasmic and membrane fraction protein profile. The small amount of proteins bands isolated from cells cultivated in sewage sludge seems to be connected with the fact that many cells of indigenous bacteria were destroyed. Proteins, especially, of the cytoplasmic and periplasmic compartments could be released to supernatant. Earlier, Tabita and Lungren [15] reported that treatment of cell suspension of A. ferrooxidans with organic acids, such as oxaloacetic or hexanoic, led to cell envelope disruption and the release of various cell components, including proteins, DNA, RNA and reducing sugars. 4.
CONCLUSIONS The cultivation of chemolithotrophic bacteria under nutrient conditions prevailing in sewage sludge results in changes in physiology (growth), morphology, ultrastructure and protein synthesis of iron- and sulphur-oxidising bacteria. The organic compounds may directly inhibit the iron and sulphur oxidation resulting in slower decrease of the pH of the sludge and slower growth rate. The disruption of cells cultivated in the sewage sludge was observed, but many bacterial cells were able to survive in the presence of organic contaminants and heavy metals. The morphological and ultrastructural changes observed under the electron microscope were evidently connected with adaptation to those conditions. Moreover, the application of the mixture of naturally existing acidophilic 275
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microflora, containing among other the described strains, in the biotechnology of heavy metal removal from sewage sludge is possible and very effective. ACKNOWLEDGEMENTS This research was supported by grant from KBN (Committee of Scientific Research, Poland) No. 6PO4G 095 21. We express appreciation to A. Łuczak (MSc) for technical assistance. We wish to thank Dr. E. Lewandowska (Laboratory of Electron Microscopy, Intitute of Psychiatry and Neurology, Warsaw) and Dr. M. Sobolewska (Laboratory of Electron Microscopy, Warsaw University) for TEM and SEM observations. REFERENCES
1. J.H. Tuttle and P.R. Dugan, Can. J. Microbiol. 22 (1976) 719. 2. G.J. Zagury, K.S. Narasiah and R.D. Tyagi, Environ. Technol. 15 (1994) 517. 3. C. Gomez and K. Bosecker, Geomicrobiol. J. 16 (1999) 233. 4. R.D. Tyagi, J.F. Blais and J.C. Auclair, Environ. Pollut. 82 (1993) 9. 5. J.F. Blais, R.D. Tyagi and J.C. Auclair, Wat. Res. 27 (1993) 101. 6. H.R. Rogers, The Science of Total Environ. 185 (1996) 3. 7. M.P. Silverman, J. Bacteriol. 94 (1967) 1046. 8. R.L. Starkey, Soil Sci. 39 (1935) 197. 9. M. Ostrowski and A. Sklodowska, World J. Microbiol. Biotech. 9 (1993) 328. 10. H.C. Neu and L.A. Heppel, J. Biol. Chem. 240 (1965) 3685. 11. B. Witholt, H. Boekhout, M. Brook, J. Kingma, H.V. Herrikhuizen and L. Leij, Anal. Biochem. 74 (1976) 160. 12. D.P. Kelly, Ann. Rev. Microbiol. 25 (1971) 177. 13. J. Pecia, E.A. Marchand, J. Silverstein and M. Hernandez, Appl. Environ. Microbiol. 66 (2000) 3065. 14. J.M. Shively, Ann. Rev. Microbiol. 28 (1974) 167. 15. R. Tabita and D.G. Lundgren, J. Bacteriol. 108 (1971) 328. 16. J.H. Tuttle, P.R. Dugan and W.A. Apel, Appl. Environ. Microbiol. 33 (1977) 459. 17. L.M. Pope, D.S. Hoare and A.J. Smith, J. Bacteriol. 97 (1969) 936. 18. C.A. Jerez, M. Seeger and A.M. Amaro, FEMS Microbiol. Lett. 98 (1992) 29. 19. A.M. Amaro, D. Chamorro, M. Seeger, R. Arredondo, I. Peirano and C.A. Jerez, J. Bacteriol. 173 (1991) 910. 20. M.C. Taira, C. Reche, S. Porro and S. Alonso-Romanowski, in Biohydrometallurgy and the environment toward the mining of the 21st century, Elsevier, (1999) 115.
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Kinetics of ferrous iron oxidation with Sulfolobus metallicus at 70ºC Gabriel Meruane, Carlos Cárcamo and Tomás Vargas Centro de Hidrometalúrgia/Electrometalúrgia (CHEM-CHILE). Depto. de Ingeniería de Minas Depto. de Ingeniería Química. Universidad de Chile. Tupper 2069 Santiago, Chile. Abstract The kinetics of ferrous ion oxidation under the catalytic action of Sulfolobus metallicus at pH = 1.5 and 70ºC was characterized. Measurements were conducted using a two-chamber electrochemical cell, which enabled to obtain accurate measurements of the oxidative activity of Sulfolobus metallicus on ferrous iron at controlled solution redox potential, preventing the influence of ferric iron precipitates formation. The kinetic oxidation of ferrous iron with Sulfolobus metallicus was well described with a Monod expression with a ferric inhibition term (product competitive inhibition) given by the equation: V Fe2 +
VMax × [ Fe 2+ ] = [ Fe 2+ ] + K S × (1 + K I * [ Fe 3+ ])
with Vmax = 1.8 x 10-6 [µg Fe/(hr cell)], Ks = 203.6 [mg/l] and KI = 3.83 [-]. The maximum specific oxidation rate (Vmax) for this archaea is about 50% the value of those previously reported for mesophilic and moderate thermophilic microorganisms. On the other hand, the ferrous iron affinity constant (KS) and the inhibition constant (KI) are about 6 times larger than the respective values reported for those microorganisms. According to the present kinetic characterization the catalytic influence of Sulfolobus metallicus on ferrous iron oxidation can be optimized when operating with a solution with high ferrous iron concentration while keeping simultaneously a low ferric iron concentration. Keywords: bacterial activity; bioleaching; ferrous iron oxidation; indirect action; Sulfolobus metallicus 1.
INTRODUCTION Use of thermophile microorganism in the bioleaching of mineral sulphides is particularly attractive for treating sulphides such as chalcopyrite and molybdenite, which are quite refractory to dissolution in the temperature range where mesophiles operate (1040°C). The mechanism of catalysis in the bioleaching of sulfide minerals with thermophiles is partially substained by an indirect mechanism which is based on the catalytic influence of the microorganism in the oxidation of ferrous ion according to the reaction:
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1 1 ism ⎯⎯⎯⎯ → Fe3+ + H 2 O Fe2+ + H + + O2 ⎯microorgan 4 2
(1) The ferric ion produced then chemically attacks the sulfide according to the general reaction:
2 Fe 3+ + MS ⎯ ⎯→ M 2+ + S 0 + 2 Fe 2+
(2) where M correspond to a divalent metal cation. Thermophiles can effectively catalyze the oxidation of ferrous ion (reaction 1) up to temperatures between 60 – 90 °C which enables to bioleach refractory sulfides in a temperature range where its dissolution with ferric ion (reaction 2) is much faster that at room temperature. The high temperature biooxidation of ferrous ion with thermophiles has been studied by several authors. Nemati and Harrison [1] determined the ferrous ion oxidation activity of Acidianus brierleyi and concluded that was significativelly smaller than that of mesophiles like Acidithiobacillus ferrooxidans. Marsh et al. [2], Norris and Barr [3] and Mier et al. [4] reported studies on ferrous ion oxidation by Sulfolobus metallicus, an acidophillic chemolitothrophic archaea that can grow autotrophically at temperatures between 65 and 80°C. All this studies, however, report qualitative data which precludes the activity of this microorganism to be compared with other microoganisms or to model its behaviour in bioreactors. In the present work the kinetics parameters of ferrous ion oxidation under the catalytic action of Sulfolobus metallicus were determined pH = 1.5 and 70ºC. Measurements were conducted using a two-chamber electrochemical cell which enabled to obtain accurate measurements of the oxidative activity of Sulfolobus metallicus on ferrous iron at controlled solution redox potential, preventing the influence of ferric iron precipitates formation. 2.
MATERIALS AND METHODS
2.1 Archaea culture Sulfolobus metallicus was cultured in 250 ml Erlenmeyer flasks containing basal medium with the following composition: (NH4)2SO4 (0.4 g/l), MgSO4.7H2O (0.5 g/l), KH2PO4 (0.2 g/l). The culture was supplemented with a source of reduced sulfur, K2S4O6 (1.5 mmol), necessary for this microorganism to maintain an efficient ferrous iron oxidation activity [2, 3, 4]. The pH of the medium was adjusted to 1.8 with sulfuric acid. The culture was maintained at 70ºC in a rotary shaker, with periodically subculturing. The inoculum for the bioelectrochemical cell was prepared from 80 ml of culture sample that was filtered through a 0.22µm Millipore® membrane. Cells were washed three times with 20 ml of pH 1.5 sulfuric acid solution to remove iron and then resuspended in 20 ml of iron–free basal medium. The cell population in this inoculum, determined by direct counting using a Petroff-Hausser chamber, was typically in the range: 2-6 x 108 cells per ml. 2.2 Electrochemical cell The scheme of the experimental set up is shown in Figure 1. The electrochemical cell, made of Pyrex®, has two compartments separated by a 1 cm2 cation exchange membrane (Nafion® 90209). The working electrode is a platinum wire (diameter 0.5 mm, length 2 m), with a 16 cm2 total area. This electrode together with the reference electrode (Radiometer Analytical Model REF601 Hg/HgSO4 electrode) were placed in the cathodic 278
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compartment with 30 ml of basal medium containing the microorganisms to be characterized. In this compartment there was also a Cole-Palmer 5990-57 combination platinum electrode for Eh determination. The counter electrode, a platinum foil of 3 cm2 total area, was in the anodic compartment. The catholyte was stirred with a magnetic impeller and sparged with air. The electrochemical cell was placed in a thermostatic water bath maintained at 70.0 ± 0.1 ºC. The working, reference and counter electrodes were connected to a Model 363 Electrochemical Interface from EG&G (Princeton Applied Research). A 7150 Solartron Schlumberger digital voltmeter was used for current measurement. 2.3 Procedure The cathodic chamber was filled with 30 ml of basal medium which contained an initial concentration of ferrous iron varying in the range 50-200 mg/l and an initial archaea population varying in the range 2-20 x 107 cells / ml. The microorganisms were added as aliquots of the inoculum described above. A constant potential was then applied to the working electrode with values ranging between 0.2 and 0.6 V (SHE), which induced a cathodic current related to the reduction of ferric iron. The potential was mainatined until a final steady-state cathodic current was reached which indicated that the rate of ferrous generation at the cathode equaled the rate of biological ferrous iron oxidation in the solution. e
-
REFERENCE ELECT RODE
Potentiostat CE
WE
MEMBRANE O2
REDOX ELECT RODE
+
H
+
¼ O2 +H
½ H2 O
+
¼ O2 +H
½ H2 O
Fe
3+
Fe
2+
Archaea COUNT ER ELECT RODE
WORKING ELECT RODE
Figure 1. Schematic diagram of the experimental apparatus
Successive amounts of ferrous iron were added into the cathodic chamber of the cell, each one inducing a further increase in the steady-state cathodic current related to the increase in the rate of biological ferrous iron oxidation. Each time that a steady – state current was established, the value of this current and the Eh were recorded and the solution was sampled to determine the microorganisms and total iron concentrations. Total iron was determined by the o-phenantroline method [5]. Ferrous iron was determined from total iron concentration and Eh determinations, using a Nernst type equation obtained experimentally [6 - 9]. The Nernst equation specifically determined, at 70ºC in the same basal medium for a total iron range 0.05 – 1 g/l and the [Fe3+]/[Fe2+] ratio range 0.01-100, was:
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⎛ [ Fe 3+ ] ⎞ ⎟ Eh = 0.689 + 0.0675 × log⎜⎜ 2+ ⎟ ⎝ [ Fe ] ⎠
(3) where Eh is the solution potential in Volts versus the Standard Hydrogen Electrode (SHE). From each steady–state cathodic current the specific rate of bacterial ferrous iron v oxidation, Fe 2 + , was determined from the expression: I VFe2 + = z×F ×N (4) V where Fe2 + is expressed in (mol Fe2+ / ( s x cell), I is the steady–state cathodic current (A), z is the number of electrons in the ferric reduction (z=1), F is the Faraday constant (96496 C/eq) and N is the number of bacteria in the cathodic compartment.
3.
RESULTS AND DISCUSSION Figure 2 shows the specific rate of ferrous iron oxidation by Sulfolobus metallicus as a function of ferrous iron concentration determined at different applied potentials (Va vs SHE). Plotted values correspond to the average of two duplicate experiments. Ferrous iron concentration in this figure is the one established in the catholyte when the cathodic current reaches a steady value at each applied potential. It can be observed that rate of ferrous iron oxidation increases with the concentration of ferrous iron, but for a given ferrous iron concentration the rate of oxidation decreases when the applied potential becomes less cathodic. The decrease of Va was paralleled by an increase in the solution Eh, which evidenced an increase of the ferric/ferrous ratio established in solution. Therefore, the decrease in bacterial oxidative activity obtained when Va was less cathodic can be associated to the inhibiting influence of ferric iron, which increases its concentration when Va decreases. The inhibiting influence of ferric iron can be more clearly visualized in Figure 3 which show now measured ferrous iron oxidation rates as a function of the ferric/ferrous ratio established in solution at each applied potential Va. 1,40E-06
V Fe2+ [ug Fe/(hr cell)]
1,20E-06 1,00E-06 8,00E-07 6,00E-07 4,00E-07 2,00E-07 0,00E+00 0
200
400
600
800
1000
2+
[Fe ], mg/L Va = 200
Va = 400
Va = 600
Figure 2. Effect of the ferrous iron concentration and the applied potential at the cathode on the biooxidation rate. Va represents the working electrode potential 280
Bioleaching Applications 1,60E-06 1,40E-06
V Fe2+ [ug Fe/(hr cell)]
1,20E-06 1,00E-06 8,00E-07 6,00E-07 4,00E-07 2,00E-07 0,00E+00 500
550
600
650
700
750
800
Solution Eh, mV vs SHE
Figure 3. Effect of solution Eh on the ferrous iron biooxidation rate
Considering the previous experimental results the kinetics of ferrous iron oxidation with Sulfolobus metallicus was described in terms of a Monod equation with ferrous iron as limiting substrate with a ferric inhibition term (product competitive inhibiton), of the following form: V Fe2 + =
VMax × [ Fe 2+ ] [ Fe 2+ ] + K S × (1 + K I * [ Fe 3+ ])
(5) This equation has been found to describe well the kinetics of ferrous iorn oxidation with mesophilc iorn oxidizers such as Acidithiobacillus ferrooxidans [6, 10, 11, 12]. Equation (5) was applied to experimental data shown in Figures 2 and 3 using a nonlinear least square regression and a good correlation was found (92%) with the values of parameters shown in Table 1. Kinetics parameters for equation 5 previously reported for the oxidation of ferrous iron with Acidithiobacillus ferrooxidans [7] are also included in Table 1, for comparing purposes. Values in Table 1 show that the maximum specific oxidation rate (Vmax) for Sulfolobus metallicus, the most determinat kinetic parameter, is only about about a half the value of that reported for mesophilic microorganisms. On the other hand, the ferrous iron affinity constant (Ks) of S. metallicus is about 6 times smaller than the respective value for A. ferrooxidans which indicates that ferrous iron oxidation with this thermophile will be much more effective at high ferrous iron concentration. Finally, the inhibition constant (KI) of S. metallicus is about 6 times larger than the respective value of A. ferrooxidans which indicates a strong inhibiting influence of ferric iron on the activity of this thermophile. The influence of the parameters shown in Table 1 on the kinetics of ferrous iron oxidation with S. metallicus and A. ferrooxidans can be visualized in Figure 4. Curves in that figure show ferrous iron oxidation rate as a function of ferrous iron and ferric iron concentration for both types of microorganisms, calculated from equation 5 with parameters in Table 1, at 2 different ferric iron concentrations. According to this figure the oxidative activity of S. metallicus can be comparable to that of A. ferrooxidans if the ferrous iron concentrations is, for a similar ferric iron concentration, six times higher than the one used with the mesophile. Similarly, the oxidative activity of S. metallicus can be comparable to that of A. ferrooxidans in the whole range of ferrous iron concentration, if 281
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the ferric iron concentration used by the thermophile is 1/15 of the ferric iron concentration used by the mesophile. According to the present kinetic characterization one can conlcude then that the catalytic influence of Sulfolobus metallicus on ferrous iron oxidation can be optimized when operating with a solution with high ferrous iron concentration while keeping simultaneously a low ferric iron concentration. Table 1. Adjusted parameters values for Sulfolobus metallicus and reported values for Acidithiobacillus ferrooxidans
VMax KS KI
Value for Acidithiobacillus ferrooxidans 2.817 x 10-6 73.128 0.641
Value for Sulfolobus metallicus 1.836 x 10-6 203.55 3.831
Parameter [µg Fe/(hr cell)] [mg/L] [-]
V Fe2+ [ug Fe/(hr cell)]
2,50E-06
2,00E-06
1,50E-06
1,00E-06
5,00E-07
0,00E+00 0
500
1000
1500
2000
2500
3000
2+
[Fe ], mg/L S. metallicus 0,1 g/l Fe3+ S. metallicus 1,5 g/l Fe3+
At. ferrooxidans 0,1 g/l Fe3+ At. ferrooxidans 1,5 g/l Fe3+
Figure 4. Ferrous iron oxidation rates with Acidithiobacillus ferrooxidans and Sulfolobus metallicus calculated according to equation (5)
The strong inhibiting effect of ferric iron on the oxidative activity of S. metallicus is probably linked to the formation jarosite-type precipitates which is usually triggered in the presence of this ion. In fact, the formation of jarosites has been proved to play an inhibiting effect on the ferrous iron oxidative activity of A. ferrooxidans [13]. The inhibiting influence of these precipitates is expected to be much important in the case of thermophiles because the solubility of jarosites decreases and the kinetics of jarosites formation increases in the temperature range where these microorganisms operate (7090°C). Bioleaching of sulfides with thermophiles usually operate with solutions containing large concentrations of ferric iron (5-17 g/l) in the presence of important concentrations of suspended jarosite-like precipitates. As in this type of reactors normally operate with a consortia of thermophilic microorganisms it very likely than the demand for ferrous iron oxidation would be relying on the oxidative activity of microorganisms different than S. metallicus. 4.
CONCLUSIONS The kinetics of ferrous ion oxidation under the catalytic action of Sulfolobus metallicus at pH = 1.5 and 70ºC can be well described with a Monod expression with a ferric inhibition term (product competitive inhibition) given by the equation:
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V Fe2 + =
VMax × [ Fe 2+ ] [ Fe 2+ ] + K S × (1 + K I * [ Fe 3+ ])
(5) The values of the parameters in this equation are: Vmax = 1.8 x 10 [µg Fe/(hr cell)], Ks = 203.6 [mg/l] and KI = 3.83 [-]. According to this kinetic characterization, in order to optimize the catalytic influence of Sulfolobus metallicus on ferrous iron oxidation it is necessary to operate in solutions with high ferrous iron concentration while simultaneously minimizing the concentration of ferric iron. -6
ACKNOWLEDGMENTS This work was supported by BHP-Billiton and Conicyt under FONDEF D00I-1050 project. REFERENCES
1. M.Nemati and S. Harrison, Minerals Engineering, 13 (2000) 19. 2. R.M. Marsh, P.R. Norris and N.W. Le Roux, In Progress in Biohydrometallurgy, Cagliari, Italy (1983) 71. 3. P.R. Norris and D.W. Barr, FEMS Microbiol Lett., 28 (1985) 221. 4. J.L. Mier, A. Ballester, F. Gonzalez, M.L. Blázquez and E. Gomez, J. Chem. Tech. Biotechnol. 65 (1996) 272. 5. L. Herrera, P. Ruiz, J.C Aguillon, and A. Fehrmann, J Chem Technol Biotechnol 44 (1989) 171. 6. G. Meruane, Ph.D. Thesis (2002). 7. G. Meruane, C. Salhe, J. Wiertz and T. Vargas, Biotechnol. Bioeng. 80(3) (2002) 280. 8. P. Graindorge, S. Charbonnier, J.P. Magnin, C. Mauvy and A. Cheruy, J Biotechnol 35 (1994) 87. 9. B. Pesic, D.J. Oliver and P. Wichlacz, Biotechnol Bioeng 33 (1989) 428. 10. C.A. Jones and D.P. Kelly, J Chem Technol Biotechnol 33B (1983) 241. 11. M.S. Liu, R.M.R. Branion and D.W. Duncan, Can J Chem Eng 66 (1988) 445. 12. M. Boon, G.S. Hansford and J.J. Heijnen, In Biohydrometallurgical Processing, proceedings of the International Biohydrometallurgy Symposium IBS-95, Eds. T. Vargas, C.A. Jerez, J.V. Wiertz and H. Toledo, (1995) Vol. 1 p 153. 13. G. Meruane and T. Vargas, Hydrometallurgy (2003, in press).
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Kinetics of sulphur oxidation: pH and temperature influence on bioleaching E. Patinoa, R. Sandovala and J. Frenayb a
b
Engineering Department, University Arturo Prat, Iquique-Chile Metallurgie et Traitement des Minerais, Geomac-Faculty of Applied Sciences, Université de Liège-Belgium
Abstract The oxidation of sulphur constitutes a key process in mineral biotechnology from the point of view chemical and biological. The sulphur oxidation rate was determined in one strain of Acidithiobacillus ferrooxidans at 35°C and in one strain of Sulfolobus metallicus at 70°C, grown in different media. Because of the eventual liberation of protons upon sulphate formation, the sulphur oxidation was monitored by following the pH variations. 1.
INTRODUCTION Bioleaching is playing an increasingly important role in the extraction of the metals from low-grade ores and refractory sulphide minerals. Two mechanisms have been proposed for bioleaching of sulphide minerals, the direct and the indirect. The direct dissolution of minerals is caused by the attack on sulphide by the enzymatic system of the microorganism situated at the mineral surface. In the indirect mechanism the primary attack on the sulphide mineral is assumed to be a ferric iron chemical leaching with the role of micro-organism, whether at the mineral surface or not, being to oxidise ferrous iron to ferric iron, maintain a high redox potential and also to oxidise the sulphur produced to sulphate. The chemical oxidation can be complete, in which case ferrous iron and sulphate are produced. The role of the bacteria is to regenerate ferric iron by oxidizing the ferrous iron. This does not require that the bacteria are closely associated with the mineral surface. The chemical oxidation reaction can also be incomplete, in which case ferrous iron and elemental sulphur are produced. The role of the bacteria is to oxidize the sulphur to sulphate and so preventing the mass transfer limitation caused by sulphur layer [1]. Acidithiobacillus ferrooxidans, Acidithiobacillus thiooxidans and Leptospirillum ferrooxidans are conventionally implemented in bioleaching processes at temperatures ranging from 30 to 45°C and as such the mesophilic biooxidation of ferrous iron has been studied extensively [2,3]. In recent years, however, there has been interest in application of high temperature processes (65 to 80°C) utilising thermophilic archaea such as Sulfolobus acidocaldarius, Sulfolobus metallicus, Acidianus brierleyi and Metallosphaera sedula [4-6]. These earlier studies focused mainly on the improvement of bioleaching rate through verification of optimum particle size and pulp density mineral, as well as the bioreactor configuration.
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Bioleaching Applications
In Table 1 the simplified stoichiometric equations of the incomplete and complete chemical oxidation and direct bacterial oxidation of sphalerite, pyrite, covellite, chalcocite and chalcopyrite are listed. Table 1. Stoichiometric equations for the direct and indirect bacterial oxidation of different minerals [3]
Stoichiometric
Mechanism
ZnS + 2 Fe3+ → Zn 2 + 2 Fe 2 + + S 0 3+
ZnS + 8 Fe + 4 H 2O → ZnSO4 + 8 Fe ZnS + 2 O2 → ZnSO4
Incomplete 2+
+8 H
+
Complete Biological
FeS2 + 2 Fe3+ → 3 Fe 2 + + 2 S 0 3+
Incomplete 2−
2+
FeS 2 + 14 Fe + 8 H 2O → 15 Fe + 2 SO4 + 16 H 15 1 FeS2 + O2 + H 2O → Fe3+ + 2 SO4 2 − + H + 4 2 3+ CuS + 2 Fe → Cu 2 + 2 Fe2 + + S 0
+
Complete Biological Incomplete
CuS + 8 Fe3+ + 4 H 2O → CuSO4 + 8 Fe 2 + + 8 H + CuS + 2 O2 → CuSO4
Complete Biological Incomplete
Cu2 S + 2 Fe3+ → CuS + Cu 2 + 2 Fe 2 + 2−
Cu2 S + 10 Fe3+ + 4 H 2O → 2 Cu 2 + + SO4 + 10 Fe 2 + + 8 H + 5 Cu2 S + O2 +2 H + → 2 Cu 2 + + SO4 2 − + H 2O 2 CuFeS2 + 4 Fe3+ → Cu 2 + + 5 Fe2 + + 2 S 0 2−
CuFeS 2 + 16 Fe3+ + 8 H 2O → Cu 2 + + 17 Fe 2 + + 2 SO4 + 16 H + 1 17 CuFeS 2 + O2 + H + → Cu 2 + + Fe3+ + 2 SO4 2 − + H 2O 2 4
Complete Biological Incomplete Complete Biological
As can see from Table 1, the ferrous iron and sulphur formation constitutes two key processes in mineral biotechnology due at their further required chemical and/or biological oxidation. Biological oxidation of ferrous iron by mesophile have been demonstrated to be 106 times faster than chemical and has received great attention [2,3]. As thermophilic bioleaching is applied increasing in the mining industry, there is a growing need for an understanding of the principles governing the high temperature bioleaching of the sulphide mineral. Fundamental kinetics studies indicate that the rate of chemical reactions approximately doubles with every 10°C rise in temperature. Not surpringly, biological oxidation of elemental sulphur has received as much attention. This paper shows the comparative kinetics of the biological sulphur oxidation, by mesophilic and thermophilic biooxidation and a biotechnology strategy is recommended. 2.
EXPERIMENTAL, MATERIALS AND METHODS Both mesophilic and thermophilic bacteria have been described which can utilize inorganic compounds of Fe and S as electron donor [7,8]. At present, although thermophiles are characterized by higher oxidation rates, only mesophilic iron-and sulphur-oxidizing acidophiles are used in large-scale leaching processes for metal recovery from sulphides ores.
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Bioleaching Applications
The chemical oxidation of sulphur can be expressed by the following reaction: S 0 + 6 Fe 3+ + 2 H 2 O → H 2 SO4 + 6 Fe 2+ + 6 H +
(1) Previous studies have showed that chemical oxidation of sulphur by ferric iron is very low compared with that mesophile biooxidation. However, at higher temperature, i.e 70°C, chemical oxidation contribution is very important reaching up 20% of the global oxidation. The ultimate end product of the bacterial sulphur oxidation is the sulphate ion. Therefore, the elemental sulphur oxidation can be presented with the following net equations: S 0 + 3 O2 + 2 H 2 O → 2 SO4
2−
S 0 + 3 O2 + 2 H 2 O → 2 HSO4
+ 4H+
2−
(2)
+ 2H +
(3) Because of the eventual liberation of protons upon sulphate formation, the sulphur oxidation can be monitored by following the increase in hydrogen ion activity. Under normal culture conditions, bacterial sulphur oxidation is coupled with growth and the measurement of oxidation based on pH changes (dpH/ dt) can therefore also be used to monitor bacterial growth. Exponential growth coupled with sulphur oxidation hence should yield a linear decrease in pH values. Measurement of pH was the method of choice in the present work because homogeneous sampling and direct quantitative measurement of growth by determination of biomass concentration by either cell counts or chemical methods have not been developed for cultures growing with flowers of sulphur. Acidithiobacillus ferrooxidans from Cerro Colorado Mining, Iquique-Chile, was grown in MC medium at pH 2.6 and temperature 35°C and Sulfolobus metallicus obtained from the German Collection of Microorganism and Cell Cultures-Germany, was growing in Norris medium at pH 2.6 and temperature 70°C. Composition of utilized medium are described in Table 2. Shake flasks cultures consisted in 100-ml of nutrient medium in 250ml shake flask at 180 rpm and suplemented with 1 g of elemental sulphur. Enriched experiments under shake flask cultures were incubated at 35 and 70°C in mineral salts media. Growth of the cultures with elemental sulphur was monitored by pH measurements. Table 2. Nutrient medium utilized in cultures (NH4)2SO4 .
MgSO4 7H2O .
K2HPO4 3H2O KH2PO4
MC medium (g/l)
Norris medium (g/l)
0.4
0.4
0.4
0.5
0.056
--
--
0.2
3.
RESULTS AND DISCUSSION Figure 1 shows the measurement of variation pH during biological sulphur oxidation at 35 and 70°C. Results can be adjusted by the following equations: pH = 2.6938 - 0.0775*t at 35°C (4) pH = 2.7 - 0.12*t at 70°C (5) where t is time in hours. 287
Bioleaching Applications
3 p H variat io n at 3 5°C
2 .5
p H variat io n at 70 °C
p
2 1.5 1 0 .5 0 0
10
20
30
T ime(hours)
Figure 1. Variation of pH during cultures at 35°C (mesophiles) and 70°C (thermophile) 0 .59
0 .14 Hyd ro g en co ncentratio n M at 3 5°C
Rate o f H p ro d uctio n M /h at 70 °C
0 .3 9 0 .2 9 0 .19
0 .1 0 .0 8 0 .0 6 0 .0 4
0 .0 9 -0 .0 1
Rate o f H p ro d uctio n M /h at 3 5°C
0 .12
Hyd ro g en co ncentratio n M at 70 °C
Rate producti
H concentrati
0 .4 9
0 .0 2 0 0
10
20
30
T ime(hours)
Figure 2. Evolution of hydrogen concentration during cultures at 35 and 70°C
0
5
10
15
20
T ime(hours)
Figure 3. Rate of acid production during the experiments, calculated from results obtained in Fig. 1
Extreme thermoacidophiles, mostly members of the archaea order Sulfolobus, represent a unique type of extremophile in that they must deal with high temperature and low pH simultaneously. In agreement with this, the pH variation rate in thermophilic culture is 1,5 times greater than mesophilic culture, as shows the results presented in Figure 1. From this, the variation of molar concentration and the hydrogen production rate can be calculated, as is shown by Figure 2 and Figure 3. From Fig 3, it can see that at the beginning of the culture, acid generation rate is almost near to zero, due principally to the period of bacterial attachment on the sulphur surface. After this, extreme culture reachs faster acid production rate compared to mesophile as result of both, the required acidic environment by thermophilic microorganism and thermodynamical considerations. On the other hand, a stoichiometric equation for bacterial growth on elemental sulphur can be derived from the elemental balances on C, H, O, N, S and the charge balance: ⎛ 1.5 − 1.05 Ysx CO2 + 0.2 NH 4 + + ⎜⎜ ⎝ 1.05 Ysx
⎛ 0.2 Ysx + 2 ⎞ + ⎛ 0.6 Ysx + 1 ⎞ ⎞ 1 1 ⎟⎟ H ⎟⎟ H 2 O → CH 1.8 O0.5 N 0.2 + ⎟⎟ O2 + SO4 2− + ⎜⎜ S + ⎜⎜ Ysx Y Y Y sx sx sx ⎝ ⎠ ⎝ ⎠ ⎠
As cells grow there is, as a general approximation, a linear relationship between the amount of biomass produced and the amount of substrate consumed (in this case elemntal sulphur). This relationship is expressed quantitatively using the biomass yield, Ysx, which must be experimentally determined. From stoichiometric relationships, it’s possible to 288
Bioleaching Applications
evaluate the sulphur and oxygen consumption rates, rSo and rO2 respectively, in function of the biomass production rate, rx , as: − rSo =
rx YSX
⎛ 1.5 − 1.05YSX − rO2 = ⎜ ⎝ 1.5YSX
(6) ⎞ ⎟ rX ⎠
(7)
Eliminating Ysx from equations (6) and (7) yields the degree of reduction balance (relationship between the oxidation rate of substrate and oxygen and carbon dioxide consumption rates): −1.5rSo = −1.05rO2 − 1.05rCO2
(8)
Due that at higher temperatures, the stress conditions (chemical and biological) are more important, and the microorganism uses a major amount of substrate to gain energy for maintaince. As strategy, this biotechnological tool can be used to evaluate the biokinetic and bioenergetic parameters of sulphur biooxidation. Further studies will be carried out, to continuously measurement of oxygen and carbon dioxide consumption rates. 4.
CONCLUSIONS The following conclusions could be made: - It is possible to adapt the mesophile and thermophile bacterial culture using elemental sulphur as substrate. - Production of acid is faster in thermophilic culture than mesophilic - From measurement of pH evolution, it is possible to calculate the sulphur consumption rate using biomass yield.
REFERENCES
1. 2. 3. 4. 5. 6. 7. 8.
E. Torma. 1988. Leaching of metals. In: Biotechnology. 6b:367-399,Rehm H. J. (ed) M. Nemati. 1996. PhD Thesis, UMIST, Manchester, UK M. Boon, 1988. PhD Thesis, Technical University of Delft, The Netherlands. D. E. Rawlings, Biomining: theory, microbes and industrial processes (1997) SpringerVerlag, Berlin. Y. Konishi, K. Kogasaki, S. Asai. 1997.Chemical Engineering Science, 52(24):45254532. J. Han, M. R. Kelly.1998. Biotechnology and Bioengineering 58(6):617-24. L. Ahonen and O.H. Tuovinen. Appl. Environ. Microbiol. 55 (1989) 312. J.A. Brierley and C.L. Brierley. In T.D. Brock (ed) Thermophile: general, molecular and applied microbiology (1986)
289
15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Leaching of iron from China clay with oxalic acid: effect of acid concentration, pH, temperature, solids concentration and shaking S. K. Mandal and P. C. Banerjee* Indian Institute of Chemical Biology, 4 Raja S. C. Mullick Road, Kolkata-700032, India
Abstract China clay is an important mineral, which is used in the manufacture of ceramics and refractory, as also in other industries. Mined China clay contains iron oxides and silicates as impurity; if present in excess of a threshold level, the impurities affect the commercial value of the products. The currently available processes for lowering iron content in China clay to the desired level (< 0.8%) are energy and cost intensive, not sufficiently flexible, and may cause environmental pollution. An alternative approach for iron removal consists in the development of a biotechnological process which is expected to be cost-effective, less complex and eco-friendly. We reported earlier that several fungi, especially Aspergillus niger, and their culture filtrates could leach sufficient amount of iron from a China clay sample; oxalic acid was found to be the most active component of the culture filtrate. We now report the rates of iron leaching from another China clay sample by oxalic acid and by the culture filtrate of A. niger NCIM 548 that was found to be the most active strain in our earlier study. The rates increased with temperature (T) and followed biphasic kinetics. The effect of oxalic acid concentration (C), pH (H), solids concentration or pulp density (P), time, as also rate and mode of agitation on the rate of iron leaching is described. The rate of leaching with oxalic acid (Ro) can be calculated theoretically from the following relationship: Ro ~ (C)0.76 (T)1.76 (H)0.80 (P)0.20 under the specified set of conditions. Using the same concentration of oxalic acid in A. niger culture filtrate, the relationship of the rate differed; this may be due to the influence of other metabolites present in the culture filtrate on the rate. Keywords: iron leaching, oxalic acid, reaction rate, activation energy, Aspergillus niger
1.
INTRODUCTION The importance of China clay in the manufacture of pottery, ceramics and refractory, as also in other industries is well known [1]. The common impurities of the natural mineral are iron oxides and silicates, which impart poor quality to the finished products and cause other problems, if present in excess. For the production of high quality materials, the iron content in China clay should be lower than 0.8% (w/w). Many methods,
* Corresponding author: Fax: 91-33-24733967/24730284; E-mail:
[email protected]
291
Bioleaching Applications
such as froth floatation, gravity separation, acid treatment, reductive roasting and magnetic separation are used for the beneficiation of China clay [2, 3]. Based on these physico-chemical methods, several industrial processes (and related patents) have been developed [4-7]. But these operations are expensive, energy-intensive and not sufficiently flexible, and give rise to environmental pollution. They also are unable to lower the iron content to the desired level. On the other hand, suitable biotechnological methods are expected to produce low-iron clay at lower cost under environmentally safe and relatively less complex conditions [8, 9]. Iron is an essential element for growth. Though it is abundant in nature, it remains mostly in the insoluble state. Microorganisms have therefore evolved special mechanisms for extracting it from nature. One of them is the production of metabolites like organic acids and siderophores [10-12]. Compared to abiotic processes, these mechanisms help microorganisms in extracting more iron from several minerals [13]. Commercial feasibility of iron removal by bioleaching was first studied in 1980’s [8, 14]. A process was developed for the removal of iron from quartz sands, kaolins and clays using the culture filtrates as leaching solution of acid producing fungi, mainly Aspergillus niger, at high temperature (90°C) [15]. Recently, a bacterial consortium occurring in kaolins and kaolin-containing rocks was utilized for removing iron from such materials. Microbial treatment followed by iron removal through subsequent magnetic separation resulted enrichment of kaolins and other minerals [16]. The percentage of iron in most of the China clay deposits in India is high and this cannot be lowered by conventional processes to the level required for the production of high quality materials [3, 17]. We observed that several fungi and their culture filtrates could leach iron from an iron-rich China clay sample [18]. At ambient condition, the highest iron-leaching activity was observed with the culture filtrate of an oxalic acid producing A. niger strain. It is well established that oxalic acid is a potential leaching agent for dissolving heavy metals from various minerals including clay and kaolin, and biohydrometallurgy groups are now considering processes for removal of heavy metals depending on this property of oxalic acid [4, 12, 19-22]. It is therefore worth while to study the effects of different environmental parameters on the rate of iron leaching from iron-rich China clay. The use of culture filtrate of an oxalic acid producing A. niger strain for removing iron from a China clay containing only 0.11% iron has been reported previously [23]. But the rate equations derived for a very low iron containing China clay may not be applicable for one having high content of iron, particularly because the limits of percentage of iron, within which the rate remains linear, is not known. Moreover, the rates calculated with culture filtrates containing other components besides oxalic acid, which might have influenced the iron-leaching rate, are expected to be different from those for pure oxalic acid solution. Therefore, it is necessary to evaluate the rate equations of iron dissolution from clay using oxalic acid or A. niger culture filtrate with respect to variables like temperature, pH, solids concentration and oxalic acid concentration. This report describes the results of such a study. 2.
MATERIALS AND METHODS
2.1 Fungal strain and growth conditions The Aspergillus niger NCIM 548 strain was obtained from Prof. A. K. Guha, Indian Association for the Cultivation of Science, Kolkata. The strain was grown for 7 days at 30°C on a rotary shaker (215 rpm) in a modified medium [9] of the following composition (in g/l): glucose, 105.5; NaNO3, 1.5; KH2PO4, 0.5; MgSO4.7H2O, 0.025; KCl, 0.025; yeast 292
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extract, 1.6; and universal indicator solution, 2% (v/v). The pH of the medium was adjusted to 6 initially and maintained within 5.5-6.0 throughout the culture period with addition of 4M NaOH at regular intervals. The strain was maintained in Czapek-Dox medium [18]. 2.2 China clay The China clay sample (particle size –300 mesh BSS) used was mined from a deposit near Mukdumnagar, Birbhum district, West Bengal, India. It had the following elemental composition (w/w as oxide): SiO2, 45.72%; TiO2, 1.52%; Al2O3, 35.96%; Fe2O3, 1.87%; CaO, 0.33%; MgO, trace; Na2O, 0.18%; and K2O, 0.19%; and the loss on ignition (LOI) was 13.79%. The clay was reddish white in colour suggesting that the iron was partly present as free iron oxyhydroxide. 2.3 Treatment of China clay with oxalic acid and culture filtrate The culture filtrate was made 100 mM with respect to oxalic acid by adding the requisite amount of the acid. The clay sample was taken either in screw-capped bottles (30-ml capacity) or in Erlenmeyer flasks (100-ml capacity). The bottles containing 5 ml slurry of oxalic acid solution or A. niger culture filtrate withclay were rotated (cyclic) in a hybridization oven (Stuart, UK) at 60 rpm, and the flasks containing 25 ml of the slurry were shaken (orbital or reciprocating) at 215 rpm in an environmental incubator (Rosi 1000; Thermolyne, USA). Oxalic acid concentration (C), pH (H), solids concentration (P) and temperature (T) were varied from 10–300 mM, 0.75-4.0, 5-50% (w/v) and 40-80°C, respectively. After a specified time, either an aliquot from the containers or the whole content of a vessel was centrifuged, and the supernatant was collected for iron estimation. 2.4 Analytical methods Iron content was measured following a modified method of May and Fish [24, 25] described in detail previously [15]. In practice, a small volume of sample solution containing less than 5 µg of iron was diluted to 1ml with 0.5 ml of 0.02 N HCl and water. This solution was incubated at 60°C for 2 h after adding 0.5 ml of freshly prepared Reagent A (0.6 N in HCl and 0.142 M in KMnO4). The temperature was brought down to ambient, and 0.1 ml of Reagent B (5M in ammonium acetate, 2 M in ascorbic acid, 6.5 mM in ferrozine, and 13.1 mM in neocuproine) was added to the mixture. Absorbance at 562 nm was measured after 20 min but before 20 hr. Oxalate in the culture filtrate was precipitated with CaCl2 and estimated by KMnO4 titration. 2.5 Scanning electron microscopy A scanning electron microscope (Model No. S440) of Leo Electron Microscopy Limited (UK) was used for this purpose. 2.6 Determination of iron-leaching rates Plots were drawn with the values of log (rate of iron dissolution) on the y-axis versus log of the variable parameter X (viz. concentration, pH, temperature or solids concentration) on the x-axis. Rates (R) were determined from the slope (m) in each case and the relation between the leaching rate and the variable parameter was expressed as R α Xm.
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3.
RESULTS AND DISCUSSIONS The leaching experiments were normally conducted under the following conditions: concentration of oxalic acid, 100 mM; temperature, 60°C; pH, 1.5; pulp density, 10% (w/v); and time, 4 h. The rates of iron leaching were calculated with change in oxalic acid concentration, temperature, pH, pulp density, reaction time and shaking conditions and are described in the respective sections. All the experiments and analyses were replicated. The rate of iron dissolution was observed to be very fast during the first hour. As this might be due to the dissolution of freely available iron oxyhydroxide, the rates were calculated from the data obtained after the first hour of reaction.
Log (reaction rate)
3.1 Effect of oxalic acid concentration on the iron dissolution rate The experiment was carried out in 100-ml Erlenmeyer flask on an orbital shaker (16mm throw) at 215 rpm using concentrations of oxalic acid varying from 10 to 300 mM. Iron dissolution rate (R) was slow up to 40 mM, but increased above this concentration. From the plot of log (rate of iron dissolution) versus log (concentration), the slope was calculated as 0.76 above 40 mM (Fig. 1). Therefore, the rate equation for oxalic acid, i. e. amount (%) of iron leached/minute can be written as R~ (C)0.76, where C is expressed in mM. 0 -0,4 0
1
2
3
-0,8 -1,2 -1,6 -2 Log (conc. in mM of oxalic acid)
Figure 1. Correlation between oxalic acid concentration and iron leaching rate
The culture filtrate of A. niger was estimated to be 89 mM in oxalic acid. Since the rates for pure oxalic acid were determined at 100 mM concentration, the culture filtrate was enriched with additional oxalic acid to the same concentration (100 mM). It was noted that at the same oxalate concentration, much less iron was leached with culture filtrate than oxalic acid (5.4% vs. 24.9%). The result suggests that the culture filtrate of A. niger strain contains such materials (other than the medium components) that strongly inhibit the iron dissolution process under the conditions (short period leaching at temperature higher than ambient). It may be mentioned that on prolonged incubation for 15 days at ambient temperature (37°C), more iron was leached by the culture filtrate compared to oxalic acid [18]. 3.2 Temperature effect on the iron dissolution rate The experiments were conducted in screw-capped bottles in a hybridisation oven (cyclic rotation) at 40 to 80°C. After the first hour, iron dissolution occurred at a constant rate, which increased with temperature (Fig 2). During short-period leaching, little iron dissolution occurred with culture filtrate even at 50°C. The rates of iron dissolution for different temperatures were calculated after the first hour of leaching. From the plot of log (rate of iron dissolution) versus log of temperature 294
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(in absolute scale as T), the slope was derived as 1.76 for oxalic acid and 3.72 for culture filtrate (Fig. 3A). Therefore, the rate equations can be written as Ro ~ (T)1.76 and Rcf ~ (T)3.72 for oxalic acid and culture filtrate, respectively. When log values of the reaction rates at different temperatures were drawn against 1/T, the slope was derived as –1.76 and –3.72 for oxalic acid and culture filtrate, respectively. With the help of Arrhenius rate equation, the activation energy of iron dissolution was calculated as 8.1 and 17.2 kcal / T / mol for oxalic acid and culture filtrate, respectively (Fig. 3B) suggesting a faster reaction rate for oxalic acid than the culture filtrate. The result further indicated that the rate of oxalic acid reaction with iron was inhibited by one or more compounds present in the culture filtrate; these may be either medium components or products of the fungus. In a previous report [23], where culture filtrate of another A. niger strain and a low-iron China clay sample were used, the rate equation was derived as R ~ (T)1.25 and the activation energy was calculated as 2.31 kcal / T / mol. Low activation energy in this case might be due to the low iron content of the mineral.
Figure 2. Leaching of iron from China clay at (■) 80°C, (×) 70°C, (○) 60°C, ( ) 50°C and (c) 40°C by 100 mM oxalic acid (A) and A. niger culture filtrate (B)
Figure 3. Correlation between iron leaching rate and temperature (°C) [A], and evaluation of activation energy [B] by 100 mM oxalic acid ( ) and A. niger culture filtrate (c) 3.3 Effect of pH on the iron dissolution rate Initial pH of the oxalic acid solution was allowed to vary from 0.75 to 4.0 in this experiment that was carried out in Erlenmeyer flasks placed in a reciprocal shaker (25.5mm throw). It was observed that above pH 2.0, the amount of dissolved iron in the leached solution decreased rapidly; this was probably due to precipitation of iron at pH >2 (Fig. 4A). The highest rate of iron leaching was noted at pH 1.75, which is higher than the pK1 of oxalic acid. From the plot of log (rate of iron dissolution) versus log (pH), the slope was calculated to be 0.8 (Fig.4B). Therefore, the rate equation can be written as R ~ (H)0.8, H 295
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being the initial pH of oxalic acid solution. With culture filtrate as the leaching solution, iron leaching was highest at pH 1.25 and dropped above this pH. An almost similar observation was reported previously with culture filtrate where iron dissolution was highest at the lowest pH (0.5) tested; the rate equation was presented as R ~ (H)0.4 [23]. 30
0 -0,2
Log(reaction rate)
Leached iron (% )
25 20 15 10 5
0
0,2
0,4
-0,4 -0,8 -1,2
0 0
0.5
1 pH
1.5
-1,6
2
Log (pH)
Figure 4. Effect of pH on the iron leaching rate by oxalic acid ( ) and A. niger culture filtrate (c) [A], and correlation of iron dissolution rate with initial pH of the oxalic acid solution [B] 3.4 Effect of solids concentration on the iron dissolution rate This experiment was conducted by varying the pulp density from 5 to 50% (w/v), and was performed in the hybridisation oven. The rate of iron dissolution was observed to increase with increase in pulp density, reaching a maximum at 15%, and then declining slowly (Fig. 5). With culture filtrate, the rate of iron dissolution also increased up to 15% (w/v) solids concentration. At higher concentrations (>15%, w/v), the suspension probably becomes sufficiently thick, restricting free mixing of clay particles with the leaching solution, and thus decreasing the rate. From the plot of log (rate of iron dissolution) versus log (solids concentration) (Fig. 6), the slope was calculated as 0.2 for oxalic acid and 0.9 for culture filtrate; therefore, the rate equations are Ro ~ (P)0.2 and Rcf ~ (P)0.9, respectively. It was derived as Rcf ~ (P)0.27 in a previous report [23]. 0 0.6 Log (reaction rate)
Leached iron (%)
30
20
10 0
20 40 solids concentration
60
Figure 5. Effect of solids concentration variation on the rate of iron leaching with oxalic acid
296
0.8
1
1.2
1.4
-0.4 -0.8 -1.2 -1.6 -2 Log (solids concentration)
Figure 6. Relationship on rate of iron removal and solids concentration by 100 mM of oxalic acid solution (◊) and A. niger culture filtrate (U)
Bioleaching Applications
3.5 Effect of shaking condition on the iron dissolution rate Leaching of iron was slightly better when China clay suspension was agitated in a reciprocal rather than an orbital shaker. For example, under a set of conditions, 26% and 27.5% of iron was leached under orbital and reciprocal motion, respectively. But, from the various results obtained so far, it may be suggested that leaching must be much better under cyclic rotation if other parameters are kept constant. Variation of the shaking speed within the range of 60-250 rpm did not have much effect on iron dissolution. 4.
CONCLUSION This study was conducted with a China clay sample from which at least 40% (w/w) of total iron had to be removed to make it suitable for the production of quality materials. It was observed that the reddish colour of the clay was almost completely removed after treatment on a rotary shaker (60 rpm) with 100 mM oxalic acid (pH 1.5) for 6 h at 80°C (Fig. 7). Electron micrographs of the materials indicated reduction in the number of clumps in the treated sample (Fig. 8A) compared to the untreated one (Fig. 8B). It may therefore be concluded that the desired beneficiation of China clay could be achieved using oxalic acid. The leaching rate can be calculated theoretically from the relationship: Ro ~ (C)0.76 (T)1.76 (H)0.8 (P)0.2 under a set of conditions. The rate equations were derived at pH 1.5 where the iron-leaching rate was higher in case of culture filtrate compared to oxalic acid (Fig. 3A). This is reflected in rate equations for temperature (Rcf ~ T3.72) and pulp density (Rcf ~ P0.9). Thus, in spite of the higher activation energy, iron leaching from China clay continued at pH >1.25 when culture filtrate containing other component(s) was used.
Treated
Original
Figure 7. Photograph of China clay before (right) and after (left) leaching with oxalic acid
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Treated (A)
Original (B)
Figure 8. Scanning electron micrograph of China clay after (A) and before (B) oxalic acid treatment ACKNOWLEDGEMENTS This work was supported by the Department of Science & Technology and NES, Government of West Bengal, Kolkata. Authors are indebted to The West Bengal Projects Ltd., Kolkata, for the China clay along with its elemental composition. Sincere thanks are also due to Mr. S. Shome and other persons of the Geological Survey of India, Kolkata for the scanning electron micrographs. Authors thank Dr. B. Achari for revising the manuscript and other staff members of the institute who helped them in various ways. REFERENCES
1. Goetz, P.W., Kaolin. In: The New Encyclopaedia Britannica – Micropaedia/ Ready Reference, fifteenth ed. Encyclopaedia Britannica Inc, Chicago, vol. 6, (1985) 730. 2. Grim, R.E., Clay Mineralogy, second ed., International Series in the Earth and Planetary Sciences. McGraw-Hill, New York, (1968). 3. Kumar, S., Hand Book of Ceramics, vol. 1., Kumar & Associates, Calcutta, (1994) 5966. 298
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4. Chiarizia, R., Horwitz, E.P., New formulations for iron oxide dissolution. Hydrometallurgy, 27, (1991) 339-360. 5. Veglio, F., Factorial experiments in the development of a kaolin bleaching process using thiourea in sulphuric acid solutions. Hydrometallurgy, 45, (1997) 181-197. 6. Danilova, D.A., Tkacheva, L.P., Lapin, V.V. and Ermolaeva, Z.I., Bleaching of kaolin. U.S.S.R. Pat. No. 937410 (1982). 7. http//www.apexis.co.uk/htdocs/summaries/rc67s.htm. Removal of iron from industrial minerals. Apex Technical Report, European Community Project, E C Contract No. MA2M-CT19-0014, Project No. RC 67. 8. Groudev, S.N. and V.I. Groudeva, Biological removal of iron from quartz sands, kaolins and clay, in 15th Int Min Proc Congr, Kannes, France, vol. 2, (1985) 378-387. 9. Strasser, H., Pumpel, T., Brunner, H., Schinner, F., Improvement of quality of quartz sand by means of microbial leaching of iron oxide. Arch. F. Largerst. Forsch. Geol, B, -A, 16, (1993) 103-107. 10. Winkelmann, G., vander Helm, D., Neilands, J.B., 1987. Iron Transport in Microbes. VCH, New York. 11. Powell, P.E., Cline, G.R,, Reid, C.P.P., Szaniszlo, P.J., Occurrence of hydroxamate siderophore iron chelators in solids. Nature 287, (1980) 833-834. 12. Strasser, H., Burgstaller, W., Schinner, F., High yield production of oxalic acid for metal leaching processes by Aspergillus niger. FEMS Microbiol Lett. 119, (1994) 365370. 13. Barker, W.W., Weleh, S.A., Chu, S., Banfield, J.F., Experimental observation of the effects of bacteria on aluminosilicate weathering. American Mineral 83, (1983) 15511563. 14. Groudev, S.N., Groudeva, V.I., Iron from quartz sands- a microbial approach. Industrial Minerals (March), (1986) 81-84. 15. Groudev, S.N., Use of heterotrophic microorganisms in mineral biotechnology. Acta Biotechnol. 7, (1987) 299-306. 16. Shelobolina, E.S., Parfenova, E.Y., Avakyan, Z.A., Microorganisms of kaolins and their role in the processes of iron solubilization and transformation. In: Process Metallurgy 9A: Biohydrometallurgy and the Environment Toward the Mining of the 21st Century. Amils, R., Ballester, A. (Eds.), Part A. Elsevier, Amsterdam, (1999) 559568. 17. Wadia, M.D.N., Clays. In: Wadia, D.N. (Ed.), Minerals of India, fifth ed. National Book Trust, New Delhi, (1994) 22-33. 18. Mandal, S.K., Roy, A., Banerjee, P.C., Iron leaching from China clay by different fungal strains. Trans. Indian Inst. Met. 55, (2002) 1-7. 19. Taxiarchou, M., Panias, D., Douni, I., Paspaliaris, I. and Kontopoulos, A., Dissolution of hematite in acidic oxalate solutions. Hydrometallurgy, 44, (1997) 287-299. 20. Taxiarchou, M., Panias, D., Douni, I., Paspaliaris, I. and Kontopoulos, A., Removal of iron from silica sand by leaching with oxalic acid. Hydrometallurgy, 46, (1997) 215227. 21. Veglio, F., Passariello, B. and Abbruzzese, C., Iron Removal Process for High-Purity Silica sands Production by Oxalic Acid Leaching. Ind. Eng. Chem. Res., 38, (1999) 4443-4448. 22. Dudeney, A.W.L., Narayanan, A., Tarasova, I.I., Removal of iron from silica sands: Integrated effluent treatment by sulphate reduction, photochemical and reverse osmosis. In: Amils, R., Ballester, A. (Eds.), Process Metallurgy 9B: Biohydrometallurgy and the Environment Toward the Mining of the 21st Century, Part B. Elsevier, Amsterdam, (1999) 617-625. 299
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23. RoyChaudhury, G., Das, R.P., Biological removal of iron from China clay. Erzmetall, 43, (1990) 210-212. 24. Fish, W.W., Rapid calorimetric micro method for the quantitation of the complexed iron in biological samples. Methods Enzymol. 158, (1988) 357-364. 25. May, M.E., Fish, W.W., UV and visible spectral properties of ferritin. Arch Biochem Biophys. 190, (1978) 720-725.
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Mathematical modeling of the chemical and bacterial leaching of copper ores in stack F. Zeballosa, O. Barbosa Filhob and R. José de Carvalhoc a
Department of Engineering, ATNAS Engenharia Ltda, Rua Mariz e Barros 383, Icaraí Niterói, ZIP: 24220-120, Rio de Janeiro, Brazil b, c Department of Materials Science and Metallurgy, Catholic University of Rio de Janeiro, Rua Marquês de São Vicente 225, Gávea, ZIP: 22453-900, Rio de Janeiro, Brazil
Abstract The objective of the work is to develop a system of mathematical models to allow improving the control of the process of leaching of ores through the simulation of this in the industrial level. For this purpose, were used the continuity equation in the materials balance, the kinetic model of the shrinking-core for the dissolution of the solid, the equation of Michaelis-Menten for the enzymatic activity of the bacteria and some values of variables showed in the bibliography with empiric models to relate variables of the ones that doesn’t have experimental data or specific operationals It joins part of the work were developed in a Peruvian mining company using the database of the industrial leaching operations. Keywords: bioleaching, modeling, copper sulphides, ferrooxidans, shrinking-core
1.
INTRODUCTION In the present work, a mathematical model was developed starting from the continuity equation supposing a plug-flow of the liquid phase inside the heap. The model of the shrinking-core it was applied to represent the reaction of the ore with the acid solution. In the production of the ferric sulfate it was used the model of Michaeles-Menten. It is supposed that the predominant controlling mechanism in the leaching of the copper ores is the transport of mass, and for the oxides doesn’t exist any restriction in the application of this concept. However, in the case of the leaching of the sulfides the intermediate phenomenon of the biochemical reactions does exist (which have chemical controlling mechanism). Due to that the data of the sequential analysis of copper of a database of industrial operations of leaching it was not achieved to determine experimentally some variables (i.e. difusivity), for the ones which or it was taken available values in the literature or it was applied empiric relationships giving as result a hybrid model empiricfenomenological.
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2.
OVERVIEW The solubility of the copper minerals in sulfuric acid solution is variable, being almost complete in the case of the oxided mineral (70% for the cuprite, 100% for the azurite, the chrysocolla and the malachite) [1, 2, 3], very low in the secondary sulfides and almost null in the case of the primary sulfides. In the leaching in moderate temperature of ores that contain a mixture of oxides and secondary and primary sulfides is necessary the application of an acid solution containing ferric íons and/or quimiolitotrofic bacteria of the gender Thiobacillus or Leptospirillum among others. The most usual (temperature between 10 and 40ºC) are Acidithiobacillus Ferrooxidans and Leptospirillum Ferrooxidans. During the leaching is formed a layer of bacteria cells under the sulfide particle, these interact with the surface of the sulfide modifying her and provoking the dissolution of particle [4]. In the leaching systems, the relationship among the current generated by the flow of electrons and the electrode potential various in function of the transformations that happens in the structure and in the composition of the surface of the mineral, caused by reactions of dissolution and precipitation [4]. Thus, the ratio Fe2+/Fe3+ place should be maintained in a maximum that results in a great potential redox. The potential maximized redox provides a better use of the chemical force resulting in maxims oxidation ratios [4, 5, 6]. The presence of bacteria in the system of leaching of sulfides improvement the galvanic interaction, optimizing the selective leaching of some sulfidic minerals. The improvement is attributed to the bacterial oxidation of Fe2+ and of So produced in the anodic reaction [5]. In the primary minerals, the leaching can be so much for direct mechanism as indirect, however in the secondary sulfides, the leaching is accomplished by an oxidation that follows the indirect mechanism [4, 7, 9]: 2Cu 2 S + 2 Fe2 ( SO4 )3 ⎯ ⎯→ 2CuSO4 + 4 FeSO4 + 2CuS (1)
2CuS + 2 Fe2 ( SO4 )3 ⎯ ⎯→ 2CuSO4 + 4 FeSO4 + 2 S 0 3.
(2)
DESCRIPTION OF THE PROCESS The ore is extracted of the mine in a flow of 26.000 tons per day with a total grade of copper of 0.89%. Following, the ore is reduced to -3/8 of inch after passing for three crushing and screening combined circuits. Once reduced of size, the ore is agglomerated using 9.02 kg on average of sulfuric acid for ton of ore, being transported later to stacking. The ore in the heap is submitted a cure process for later to be leached by an average period of 256 days. Each layer of the heap is divided in cells of 85 meters of width, with variable length according to the topography of the heap and an average height of 5 meters. After the leaching the pregnant liquor solution is processed in the plant of solvent extraction and electrowinning (SX-EW) where is obtained as final product 188 tons of cathodes of copper degree A - LME.
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4.
DATA ANALYSIS
4.1 Sequential analysis It is a laboratorial procedure of chemical analysis applied to the ore to know the lixiviability of this in function of mineralogy. This procedure is also used during the industrial process to determine the metallurgy losses, what leads to an auditing efficient metal works [1, 2, 3, 10]. Solubility in Acid
Solubility in Cyanid
Chalcocite
Chalcopyrite
Covelite Bornite
Cuprite Tenorite
Covelite
Malachite
Chalcocite
Chrysocolla Cuprite
Azurite 0
20
40
60
Solubility, %
80
100
0
20
40
60
80
Solubility, %
Figure 1. Maximum solubility of the copper minerals in sulfuric acid and cyanide of sodium solution
The solubility of the copper minerals in the sulfuric acid and in the cyanide of sodium it is showed in the Figure 1 [1, 2]. The result is presented in three groups: the total copper (CuT) obtained by analysis of a complete sample, the soluble copper in sulfuric acid solution (CuSAc) and the soluble copper in solution of cyanide of sodium (CuSCN). Of these values, it can be deduced the insoluble copper, Cuinsoluble = CuT – (CuSAc + CuSCN), presumably composed for calcopyrite mainly. The soluble copper in sulfuric acid (constituted mainly by oxided mineral of copper) it will be the easiest copper percentage to leach, needing a leaching strictly acid. Already the soluble copper in cyanide (composed by secondary sulfides of copper, mainly chalcocite, cc and covelita, cv) it requests of larger times of leaching and the presence of an effective oxidizer, as the ion Fe3+. The leaching of the calcopyrite (cpy) represented by the insolubles in the sequential analysis is still more complex. The values of the sequential analysis of the first layer of the heap* are presented in the Table 1. The sequential analysis is used as base for the modelling, once this procedure supplies implicit information of the behavior of the copper minerals in function of the time and of the leaching solution.
* The sequencial analysis was applied to a sample of the ore to grinding to -100 mesh Tyler, leached in a beaker with sulfuric acid solution to 5% during one hour to 25ºC. Following the washed solid is cyanided with solution of NaCN to 10% for 30 minutes to 25ºC. The residual solid is taken to digestion and then do the analyzes of the solutions for atomic absorption.
303
100
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4.2 Statistical analysis Once clear the concept of the sequential analysis as medium to make the mathematical modelling, the data of the Table 1 are analyzed through statistical tools to know the homogeneity, the existent correlation with the operational variables and the relationship with the ore of deeper layers of the mine in the leaching. The confidence interval was calculated for a degree of confidence of 99%, it was done a dispersion analysis for all the values of the Table 1, was determined the correlation between the copper grade and the process variables, besides the significant difference among the layers composed of ore of deeper zones in the mine. After that can be comment the following: Table 1. Applied sequential analysis to the ore loaded in the first layer of the heap and made use by cells Cell
Ore, ton
CuT, %
101 102 103 104 105 106 107 108 109 110 111 112 113 114 115 116 117 118 119 120 121
237,704.00 294,839.24 350,100.43 452,339.00 292,182.80 291,211.00 139,198.00 212,302.00 195,111.00 216,465.00 226,907.99 436,985.17 288,565.22 416,516.00 368,344.00 616,632.99 436,785.00 394,887.00 293,210.00 144,089.00 137,586.00
0.90 1.06 1.20 1.01 0.90 0.89 0.93 0.79 0.76 0.75 0.87 0.84 0.79 0.78 0.98 0.85 0.90 0.95 0.83 0.80 0.76
Sequential analysis of the copper CuSAc,% CuSCN,% Cuinsoluble,% 0.24 0.55 0.11 0.16 0.74 0.16 0.21 0.85 0.14 0.19 0.68 0.13 0.17 0.59 0.14 0.16 0.58 0.14 0.18 0.61 0.15 0.15 0.53 0.12 0.17 0.52 0.07 0.27 0.37 0.11 0.16 0.51 0.20 0.15 0.51 0.17 0.16 0.54 0.09 0.11 0.53 0.14 0.15 0.71 0.12 0.16 0.59 0.10 0.12 0.58 0.20 0.19 0.54 0.22 0.19 0.48 0.16 0.20 0.51 0.09 0.26 0.38 0.13
a. The values obtained in the calculus of the confidence interval, CuT = 0.07; CuSAc = 0.03; CuSCN = 0.07 e Cuinsoluble = 0.02; are sufficiently small indicating that exists a distribution uniform of the copper grade more or less in the industrial heap; b. It can be said that the data of the cell 103 are not valid for the modelling, they are values very dispersed respect to the general average; c. Analyzing the values obtained in the determination of the significant differences among the layers, it can be said that the model obtained in function of the data of the first layer can be applied with good certainty degree the superior layers or future layers overlapped in the heap that have physical and mineralogical similar characteristics.
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5.
DEVELOPMENT OF THE MODEL They are determined two mathematical models separated, being supposed that the leaching in the heap it carries through three phases formed by the minerals of oxided copper, secondary sulfides and primary sulfides. These phases can happen in parallel inside of the heap, but with kinetics and leaching mechanisms differentiated.
5.1 Strategy of the mathematical development Will be taken in consideration the following hypotheses: a. The growth of the microorganisms is controlled by the oxidation of Fe2+ (limiting nutrient); b. The enzymatic activity behaves according to the model of Michaelis-Menten; c. The production of bacteria for unit of consumed substrate is constant; d. The drainage of the solution at the bed of the heap happens without appreciable axial dispersion (plug-flow); e. The half-time of residence of the solution inside the layer doesn't vary with the time of leaching; f. The heap presents a granulometric and grade copper species distributions homogeneous; g. The reactions of the minerals of oxided copper are controlled by diffusion of the solution inside the particles very little porous of the ore; h. The leaching of the calcopyrite is insignificantly. A leaching differentiated is produced among the copper species in the heap, or either, exist the action separate from the íons H+ as leaching agent of CuSAc and the ion Fe3+ as leaching agent of CuSCN. 5.2 Mathematical model for the soluble copper in acid The diffusion of the solution in the pores of the ore is a process usually slow, much more than the diffusion that feels in the surrounding liquid to the particle. The process of chemical diffusion is composed for five successive stages: a. Diffusion of the ion H+ in the surrounding solution to the particle until the interface liquid-solid; b. Migration of the ion H+ of the interface liquido-sólido to the particles of copper mineral for the pores to the particle of the ore; c. Chemical reaction of the ion H+ with the copper inside of the ore particle; d. Migration of the ion Cu2+ among the pores of the particle to the interface solidliquid; e. Diffusion of the ion Cu2+ of the interface solid-liquid to the surrounding solution of the ore particle. For the calculus of the copper it is used the equation (3), it describes the diffusion in state of pseudo-stationary and the fast dissolution of the oxided mineral of copper [11]:
(
2 1 − Ft , r0 − 1 − Ft , r0 3
)
2
3
⎛ 2VCu Def A0 ⎞ ⎟⎟ t = ⎜⎜ B r02 ⎝ ⎠
(3) 305
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Dε s . τ The difusivity for the species in water doesn't vary a lot and a reasonable approach to room temperature (25ºC) is D = 1.5 cm3.day-1 [11]. Comparing the experimental data with the results calculated for the diffusion in porous media, the tortuosity factor is usually τ = 2 [11] and, for a volumetric fraction of pores in the heap, εs = 0.1; results in a value of Def = 0.075 cm2.day-1. With the equation (4), can be calculated the molar volume of copper in the ore: M (4) VCu = Cu ρ ore G The effective coefficient of diffusion (Def) it can be made calculations of: Def =
Being ΜCu the molecular weight of the copper and, with a specific weight of the ore of ρore = 1.65 g.cm-3, and the fraction average of the grade of oxided copper (CuSAc) except for the cell 103 (very dispersed value), G = 1.79x10-3; then: VCu = 22.65 l.mol-1. The initial concentration of sulfuric acid in the leaching solution is: Ao = 4.21x10-2 mol.l-1. Therefore, with a initial radius average of particle, ro = 0.0922 cm and supposing that all the CuSAc reacts with the sulfuric acid and making the correction corresponding to the consumption of acid for the gangue mineral (fator ≈ 1,3), the consumption of the acid for each kilogram of extracted copper is: B = 1.30 mol.kg-1. 2 ⎤ ⎡ 2 Relating graphically the factor F and ⎢1 − F − (1 − F ) 3 ⎥ to make easier the calculus ⎣ 3 ⎦ of Ft,γο (fraction of leached copper), can be determined the value of the extraction of copper of the ore oxided in function of the time (t).
⎡ 2 Making: X = ⎢1 − Ft,γο − 1 − Ft,γο ⎣ 3 equation no lineal of the type:
(
)
2
3
⎤ ⎥⎦ , can be represented the relationship for an
⎧ ⎡ 0,0119 + 0,001 exp(− 0,0129 X ) − Ft ,r0 = −3,44 x10 −5 + 4,31⎨1 − exp(− 0,001X ) − 0,0099 ⎢ 15,34 x10 −5 ⎣ ⎩ 0,0129 exp(− 0,001X ) ⎤ ⎫ ⎥⎬ 15,34 x10 −5 ⎦⎭ (5) 2 With: r = 0.999 Calculating the value of X in Mathcad 2000 Professional®, MathSoft, Inc. for a period of 300 days (being the leaching period average of the first layer of the heap, 256 days), it can be said that the leaching of the soluble copper in acid is very fast and efficient in this stage, becoming theoretically according to the model of the shrinking-core in a 99% of reduction of the volume of the nucleus in the particle and 98,50% of the copper extraction in 300 days.
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Bioleaching Applications
5.3 Mathematical model for the soluble copper in cyanide The chalcocite is the predominant mineralogical specie in the processed ore, 81% of content of the sulfides of secondary copper. The general mineralogical distribution of the copper sulfides is the following: chalcopyrite, 24.19%; chalcocite, 60.99%; covelite, 14.04 and other sulfides, 0.78%. It can be assumed as global reactions of this process the following estequiometric relationships: 5 Cu 2 S + O2 + H 2 SO4 ⎯ ⎯→ 2CuSO4 + H 2O 2 (6)
7 FeS 2+ O 2 + H 2O ⎯ ⎯→ FeSO4 + H 2 SO 4 2
(7) If the controlling factor the leaching reaction of the secondary sulfide is the bacterial activity in function of the production or generation of Fe3+, Fe2+ should be maintained in excess and the limiting factor it starts to be the necessary oxygen for the oxidation reactions, as well as, for the metabolism of the bacteria. Once the indirect mechanism of biochemical reaction in the phase hard-liquid is much faster than the direct mechanism, is considered that the leaching of the sulfides is strictly ferric being the restricted bacterial action to the conversion of the ion Fe2+. Theoretically, it is known that for a maximum population of 1011 bacteria.m-2 (≈1010 −1 ), the dissolution of the mineral is controlled by the bacterial activity for bacteria. kg ore particles of diameter less than 2 cm [12]. Therefore, considering an effective diffusion coefficient for the Fe3+ of 5x10-11 m2.s-1 and supposing that the enzymatic behavior can be represented by the equation of Michaelis-Menten, the following relationship for the rate of copper dissolution in the stationary state is used [12]:
⎛ Bss dFss = X Vm ⎜⎜ dt ⎝ ρ bed G ss
⎞ ⎛ CL ⎟⎟ ⎜⎜ ⎠ ⎝ K m + CL
⎞ ⎟⎟ ⎠
(8) In function of the equations (6) and (7) the estequiometric factor can be calculated the following way:
Bss =
M cc M py 5 M M + 7 (RPC ) M M o cc 2 o py 2
(9) where, RPC it is the ratio among the pyrite (py) and chalcocite leaching amounts: RPC = −1 , Mcc, Mpy, Mo, they are the molecular weight of the chalcocite, pyrite and 0.80 kgφγ. kg cc oxygen respectively, Βss = 0.844 kgcc. kg O−12 . The relationship among the specific maximum ratio of breathing of the bacteria, Vm, and the temperature was obtained starting from the data of oxidation of iron by Acidithiobacillus Ferrooxidans:
⎛ 7000 ⎞ 6,8 x10−13 T exp⎜ − ⎟ T ⎠ ⎝ Vm = 74000 ⎞ ⎛ 1 + exp⎜ 236 − ⎟ T ⎠ ⎝
(10) 307
Bioleaching Applications
For an average temperature (T) of the heap to 2.500 meters on the level of the sea is approximately 293 K, there are: Vm = 8.389x10-21 kg O2 .bacteria −1.s −1 . According to the equation (10) the largest consumption of oxygen and, consequently the maxim capacity of oxidation of the substrate (Fe2+), is produced around 35°C. At this temperature, the maximum activity of Acidithiobacillus Ferrooxidans takes place, at pH = 2, km = 1x10-3 kg.m-3, and considering a concentration of oxygen dissolved in the solution of 6.5x10-3 g.l-1. The concentration of dissolved oxygen is calculated using Henry's law, supposing its equilibrium between the liquid and the gaseous phase. Henry's constant was obtained in function of the temperature using the solubility of the oxygen in the leaching solution for different temperatures, CL = Cg.He, e He express in function of the temperature: H e = 21.312 + 0.784T − 0.00383T 2 = 35.46 . The concentration of oxygen in the solution to 298 K and 101 kPa in equilibrium with the concentration of oxygen in the air, Cg = 0.24 kg.m-3, is: 0.0084 kg.m-3. Being adopted an apparent density or density of the bed of the layer of ρbed = 1800 kg.m-3 and a average bacterial population, X = 5x1013 bacteria.m-3, in the expression (8), the average fraction of soluble copper in cyanide for the heap is Gss= 5.8x10-3. Integrating the equation (8) for a period of 300 days of leaching it could be calculated the extraction of secondary copper. 5.4 General model Adding the results obtained so much for the soluble copper in acid as for the soluble copper in cyanide of sodium according to the previous models, the percentage of extracted copper referred to the copper contained in the oxides and secondary sulfides is calculated. The respective curve of extraction of the copper in function of the time and the comparison with the industrial curve of copper extraction is presented in the Figure 2. 100
Extraction, %
80 60 40 20 0 0
60
120
180
240
300
Time, day Calculated
Industrial
Figure 2. Graphic comparison of the calculated and industrial data of the copper extraction for chemical and bacterial leaching in mixed ores of copper 6.
VALIDATION Using the program EMPV®, Effective Management of Process Variability - version 1.06, it was calculated the mistake in all of the points of the curve being obtained an average mistake of 7.77%.
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It is also made a sensibility analysis for two process variables in order to visualize the performance of these variables in the model. In the Figure 3A the effect of the volumetric fraction of pores in the heap, εs, can be observed. The model simulations indicate that, for the range of εs values typically found in the industrial practice, there is no appreciable variation in the total recovery of copper by leaching. The same happens with regard to the sulfuric acid concentration in the leaching solution, Ao. On the other hand, Figure 3B shows that the most influential variable in the model is the ratio of the pyrite mass leached in relation to the mass of leached chalcocite, RPC. It can be said that, while the potential of the leaching solution doesn't reach very high values, the selective leaching of the secondary copper minerals will be comparatively more efficient. The ratio of the pyrite mass leached in relation to the mass of leached chalcocite cannot be obtained directly, but it can be dear and used as parameter of calibration of the model and as parameter of the process for the control of the electrochemical potential and the concentration of Fe3+ in the solution. Like this, this variable can be correlated empirically with another varied as the concentration of Fe3+, the time of residence of the solution in the heap and the electrochemical potential of the leaching solution. B
100 80
Total extraction, %
Total extraction, %
A
60 40 20
100 80 60 40 20 0
0 0
60
120
180
240
300
0
60
120
0,08
0.10
180
240
Time, day
Time, day 0.12
0,5
0,8
1,1
Figure 3. Effect of the volumetric pore fraction in the heap, εs (A) and of the ratio of the pyrite mass leached in relation to the mass of leached chalcocite, RPC, (B) 7.
CONCLUSIONS In industrial practice, the sequential analysis of the ore is very useful not only to support decisions regarding the crushing degree, sulfuric acid addition in the agglomeration and irrigation rate and time in the heap leaching, but also to supply valuable data to be used in the estimation of copper extraction rates through phenomenological mathematical models, with a good degree of accuracy, 92.23% in the current case. The copper oxides present a very fast leaching kinetics and the metal extraction is completed in the first leaching cycle (240 days). After this time, a sampling of the leaching residues still shows a reasonable amount of copper which would be dissolved in acid; however, this copper should belong to the secondary covelite formed during the oxidation of the chalcocite. This covelite is much more porous than the natural covelite and its leaching kinetics is very much faster. The factor of major importance in the leaching of copper sulfides is the concentration of Fe3+. 309
300
Bioleaching Applications
REFERENCES
1. J. C. Cárdenas. Lixiviación Bacteriana de sulfuros de Cobre, Sociedad Minera Cerro Verde S.A.A., Arequipa, Internal Manual of the Company (1999). 2. G. A. Parkison, R. B. Bhappu. The sequential Copper Analysis Method Geological, Mineralogical, and Metallurgical Implications, SME Preprint Presented at SME Annual Meeting, Denver – Colorado (1995). 3. S. L. Brown, J. D. Sullivan. Dissolution of Various Copper Minerals, Íon Studies in the Metallurgy of Copper, U.S. Bureau of Mines RI 3228, pp. 37 – 51. 4. G. I. Karavaiko, G. Rossi, A. D. Agata, S. N. Groudev, Z. A. Avakyan. Biogeotechnology of Metals – Manual, United Nations Environment Programme – USSR Commission for UNEP, Centre for International projects GKNT – Moscow, 1988. 5. M. Riekkola–Vanhanen, S. Heimala. Electrochemical Control in the Biological Leaching of Sulfidic Ores, Biohydrometallurgical Technologies, (1993) 561. 6. A.W. Breed, G. S. Hansford. Studies on the Mechanism and Kinetics of Bioleaching, Minerals Engineering, v. 12, n. 4, (1999) 383. 7. L. M. Yañez. Pautas para la Utilización de Microorganismos en Lixiviación, INCITEMI – Arequipa, (1985) 53. 8. W. Sand, T. Gehrke, P. Jozsa, A. Schippers. (Bio) Chemistry of bacterial Leaching – Direct vs. Indirect Bioleaching, Hydrometallurgy, n. 59, (2001) 159. 9. H. Tributsch. Direct versus Indirect Bioleaching, Hydrometallurgy, n. 59, (2001) 177. 10. W. Baum. The Use of a Mineralogical Data Base for Production Forecasting and Troubleshooting in Copper Leach Operations, Hydrometallurgy of Copper – Copper 99, International Conference, v. IV, (1999) 393. 11. R. W. Bartlett. Solution Mining – Leaching and Fluid recovery of Materials, Gordon and Breach Science Publishers, 1992. 12. J. M. Casas, J. Martinez, L. Moreno, T. Vargas. Bioleaching Model of a Copper– Sulfide Ore Bed in Heap and Dump Configurations, Metallurgical and Materials Transactions B, v. 29B, (1998) 899.
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Model for bacterial leaching of copper sulphides by forced aeration M. Sidborn and L. Moreno* Department of Chemical Engineering and Technology, Royal Institute of Technology, S-10044 Stockholm, Sweden Abstract A two-dimensional model for bacterial leaching of copper minerals was developed, which can handle aeration by natural and forced convection with and without aeration channels. The use of forced aeration increases the reaction rate in the pile and may significantly decrease the operation time, however the costs are increased. When only natural convection is used, the leaching rate is commonly limited by the oxygen availability, and factors such as bacteria population and particle size have less significance. If forced aeration is used, the oxygen availability may be arbitrarily varied. Two aspects are studied; the distance between aeration channels and the adequate aeration rate. Regarding the distance between aeration channels, it was found that for a large separation between channels the aeration is deficient in zones located between the channels. When the distance between channels is decreased, the aeration is improved, however, this improvement is less significant by further decrease of the distance between channels. The most advantageous aeration rate is a function of several factors such as particle size, type of minerals, number of bacteria, and dimensions of the pile. For example, if the pile consists of small particles, the consumption of oxygen increases and a greater aeration is required. The number of channels and the aeration rate are important when costs are considered.
Keywords: copper leaching, modelling, forced aeration, bacterial leaching, numerical modelling 1.
INTRODUCTION Bacterial leaching of copper minerals has become an important process in the mining industry. At the beginning, bacterial leaching was applied to low-grade ores, however, at present, due to exigencies of a better environment, the process is also applied to highgrade ores. Since the main reaction occurring in the heap is the oxidation of the sulphide minerals, aeration, which may take place either by natural or forced convection, is a critical issue. When natural convection is used, the leaching rate is commonly limited by the oxygen availability, and factors such as bacteria population, particle size and intrinsic reaction rate have less significance. The use of forced aeration, through channels in the bottom of the heap, may secure the adequate supply of oxygen to the greater part of the
* Corresponding author: E-mail:
[email protected]
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heap with increasing reaction rate as a result. This may significantly decrease the leaching time, but the costs are increased. The bio-leaching process is complex and involves several steps: a) oxygen is introduced into the heap by natural or forced aeration, b) oxygen oxidises the ferrous ions by means of a reaction mediated by bacteria, c) ferric ions diffuse into the ore particles to reach the sulphides, d) the ferric ions oxidise the sulphides and e) the ferrous ions emigrate to the surface of the particle to re-initiate the cycle. The bio-leaching rate is controlled by the slowest of these steps and they may coexist in different zones in the ore bed (Ritchie, 1994; Casas et al., 1998). Modelling of the leaching operation for sulphide ore beds has received considerable attention in the last years (Ritchie, 1994; Casas et al., 1998; Bartlett, 1997, Coderre, F., Dixon, D.G., 1999; Dixon, D.G., 2000; Orr, 2002; Orr and Vesselinov, 2002). The macroscopic models developed by Cathles (1994), Ritchie (1994), Casas et al., (1998), and Sidborn et al., (2003) showed that in the case of natural convection, the aeration is commonly not sufficient and in large zones of the heap the oxygen is almost totally depleted. Under these circumstances, bacterial leaching of the sulphide minerals is too low. Today, the tendency is to use forced ventilation through channels at the bottom of the heap in order to improve the air supply and to obtain high bio-oxidation rates. Lizama (2001) studied the impact of different rates of forced aeration on copper recovery. This mode of operation involves higher capital and operating costs, but they are compensated for by the faster metal recovery (Bartlett, 1997). The rate of sulphide mineral dissolution is modelled according to the unreacted core model. The transport of ferric ions from the particle surface to the reaction zone is calculated considering film diffusion, diffusion within the particle and reaction kinetics. The rate of oxidation of the ferrous ion by bacteria attached to the ore surface is modelled using the Michaelis-Menten relationship. The influences of temperature, dissolved ferric iron and dissolved oxygen in the leaching solution are considered in the kinetic formulation. In order to show the capabilities of the model, the impact on the copper recovery of the distance used between the aeration channels and the aeration rate are studied. Modelling of the process is a useful tool to aid the design and optimisation of industrial operations. 2.
THEORY
2.1 Leaching of copper ores The copper-oxide mineral dissolution kinetics is rapid and oxide minerals are readily dissolved by sulphuric acid applied on top of the heap. Copper sulphide minerals are however much more stable and can only be dissolved under oxidising conditions. In copper leaching, the oxidising agent present may be the ferric ion that diffuses into the ore particle and reacts with the metal sulphide (MS) according to
MS + 2Fe 3+ → M 2+ + 2Fe 2+ + S
(1) Microorganisms catalyse the reverse oxidation of ferrous ions to ferric ions. The most important bacteria for this purpose is Acidithiobacillus ferrooxidans. The bacterial oxidation takes place according to 4Fe 2 + + O 2(aq ) + 4H + ⎯⎯ ⎯⎯→ 4Fe3 + + 2H 2 O bacteria
312
(2)
Bioleaching Applications
This reaction consumes oxygen that has to be transferred to the leaching solution from air in the ore pile. The rate of consumption of oxygen by bacteria can be described in terms of the Michaelis-Menten equation: ⎛ [O 2 ]L ⎞ ⎟ R O2 = X Vm ⎜⎜ ⎟ [ ] + K O 2 L⎠ ⎝ m
(3)
R where O 2 is the rate of consumption of oxygen by bacteria, X is the number of bacteria per volume of bed, Vm is the maximum specific respiration rate of bacteria, O2,L is the oxygen concentration in the liquid solution and Km is the Michaelis constant for the system. For Acidithiobacillus ferrooxidans, the maximum specific respiration rate, Vm, is dependent on the temperature according to (Casas et al., 1998) Vm =
6.8 × 10 −13 T e 1+ e
236 −
−
7000 T
74000 T
(4) where T is the temperature in Kelvin. The rate-determining step in the bioleaching of small sulphide mineral particles is generally the slow intrinsic dissolution of sulphide minerals, provided that oxygen is available in the bed. For larger ore particles, however, a mixed-kinetics model has to be used to describe the leaching rate. Such a mixed kinetics model is the shrinking core model that includes resistances due to the intrinsic dissolution kinetics of the mineral, the diffusion resistance of ferric iron through an inert porous layer of reacted material, and the diffusion of ferric iron through the liquid film around the ore particle surface. The rate of decrease of the unreacted core radius for a given mineral species can be written as: MS − drC = dt ρG φ
[Fe ] 3+
1 ⎛ σ +⎜ G β ⎜⎝ D eff
⎛ 1 ⎞ ⎛ rC ⎞ ⎟⎟ ⎜ ⎟ (R − rC ) + ⎜⎜ ⎝ KC ⎠⎝ R ⎠
⎞ ⎛ rC ⎞ ⎟⎟ ⎜ ⎟ ⎠⎝ R ⎠
2
(5)
where rC is the unreacted core radius, M S is the ore molecular weight, ρ is the mineral particle density, φ is the particle shape factor, G is the copper ore grade, β is the global specific kinetics factor, σ is the stoichiometric factor, Deff is the effective diffusion coefficient, R is the mineral particle radius and KC is the mass transfer coefficient in the liquid-solid film. [Fe3+] is the concentration of ferric ions in the leaching solution. The leaching rate of a given mineral species is related to the rate of decrease of the unreacted core radius through the mass balance.
2.2 Air transport through the bed Air is mainly transported through the ore bed by convection (forced or natural). Oxygen may, however, be transported by diffusion in some zones of the heap. The density of the gas varies through the ore bed due to temperature changes caused by the exothermic reactions and due to changes in the air composition. Oxygen is consumed by the oxidation of the ferrous ions and the air humidity varies with the local temperature. The local velocity of air, qg, can be expressed as: qg = −
ρ g k rg k µ
∇P
(7) 313
Bioleaching Applications
ρ where k and krg are the intrinsic and relative gas permeabilities of the bed, g is the gas density, µ is the fluid viscosity and ∇P denotes the fluid pressure gradient. The pressure applied in the aeration channels and the air density variation cause the airflow in the bed. The transport of oxygen in the gaseous phase is described by the AdvectionDispersion (AD) equation: εg
∂O 2,g ∂t
= ε g D g ∇ 2 O 2 ,g − q g ∇O 2 , g − R O 2
(8)
O where 2, g is the concentration of oxygen, Dg is the dispersion coefficient in the gas phase and εg the volume fraction of air. 2.3 Liquid flow During leaching a solution of sulphuric acid is applied at the top of the bed at a given rate. The liquid flow is constant and is given by the irrigation rate and, in general, is independent on the permeability of the bed. The transport of solutes in the liquid phase is described by: ∂C i εL = D L ε L ∇ 2 C i − q L ∇C i + R i ∂t (6) where Ci is the concentration of species i, DL is the dispersion coefficient, εL the volume fraction of liquid, qL the liquid flow rate, and Ri the reaction rate of species i. This equation may be applied to both ferric and ferrous ions and copper ions.
2.4 Energy balance Due to the exothermic reactions, the temperature of the bed is increased. Energy is transported through the ore bed by conduction and convection. The energy transported by convection depends on both the liquid flow and the gas flow, which have opposite directions in the bed. C p,B ρ B
∂T = k B ∇ 2 T − ρ L q L ∇H L − ρ g q g ∇H g − ∆H R R ch ∂t
(9) where CpB is the mean heat capacity of the ore bed, ρΒ is the bed density, kB is the thermal conductivity of the ore bed, ρL is the liquid density, qL and qg denote the liquid and the gas flow respectively, HL and Hg denote the liquid and gas enthalpy respectively. ∆ΗR denotes the heat of reaction per mineral dissolved.
3.
CASES MODELLED A mineral consisting of chalcocite and pyrite was considered in these simulations. The pile was assumed to have a flat top and to slope downwards at the edges at an angle of 45°. In the model, for the sake of simplicity, it is assumed that the ferrous ions are present in excess and that they are not a limiting factor for the oxidation reaction. Bacteria, the population of which is assumed to be constant, mediate the oxidation of ferrous ions. The ferric ions produced diffuse into the ore particle and react with the copper mineral and copper ions are leached to the solution. The ferric ion diffusion resistances, both in the film and the particle, are taken into account. The air in the bed is assumed to be watersaturated and local equilibrium is assumed between the temperature of the solid, the liquid and the gas. Two cases are modelled:
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Bioleaching Applications
a.
The central part of the heap, where the aeration takes place through a number of channels equidistantly separated. For reason of symmetry, only one channel needed be modelled, as shown in Figure 1. b. The edge of the heap, where air is introduced by natural convection through the slope of the heap and by aeration channels at the bottom. Five differential equations are solved in addition to the equation for the airflow. These equations consider the balances of ferric ions, copper ions, oxygen, and energy and the variation in the unreacted core diameter of the particle as a result of the leaching process. The conditions inside of the bed vary with time, since the diffusion-resistance is increased as copper is dissolved. For sake of simplicity, the model results will be shown only for a short time. Modelled zone
Aeration Channel
Figure 1. Schematic picture of the heap showing the aeration channels and the modelled zone The parameters needed for these simulations are shown in Table 1. The simulation software FEMLAB® (http://www.femlab.com/) was used to solve the system of differential equations.
5.
RESULTS AND DISCUSSION This two-dimensional transient model was simulated only for short times, since our aim is to show the impact of the distance between the channels and the aeration rate on the leaching process. In the initial period the necessity of aeration is the largest. When the directly accessible copper has been depleted, the dissolution of copper will be slower and the needed aeration will be less. In the calculations, aeration is assumed to take place through a channel 0.2 m in width. The simulations were done for a heap with a height of 10 metres. Figure 2a shows the relative oxygen concentration profiles for channels separated in 20 m and over-pressures of 1000 Pa. In spite of forced aeration, the results shown zones between the aeration channels where the oxygen concentration in the air is almost zero, with insignificant dissolution of copper in these zones, as consequence of the bad aeration. For a channel distance of 12 m (Figure 2b), the aeration is significantly improved, the relative oxygen concentration at the bottom is at all the locations higher that 0.30 (6-7 vol-%). Better aeration is, of course, obtained with shorter distances, but the ventilation costs are increased.
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Figure 2. Relative concentration profiles for oxygen in the air for channel distances of 20m (on the left) and 12m (on the right) Calculations were also done for other airflow rates by changing the air pressure at the aeration channel. Figure 3 shows the case where no over-pressure is applied on the channels and the air is introduced only by the effect of natural convection. As may be observed in the figure, the zones with deficient aeration are strongly increased.
Figure 3. Relative concentration profiles for oxygen in the air for channel distances of 20 m for natural convection The results are summarised in Figure 4, where a relative measure of the extent of the leaching process is plotted for different channel distances for a heap height of 10m. The leaching process is estimated by integrating the copper concentration leaving through the bottom of the heap and dividing by the channel distance. The results show that forced aeration has a beneficial effect on the leaching process; the amount of copper that is dissolved is increased. However, also the cost of inversion and operation are increased.
Table 1. Important parameters used in the simulations Parameters Heap height Number of bacteria Michaelis constant Liquid flow Gas permeability Ambient temperature Ore copper concentration FPY parameter
316
Unit m Bacteria/m3 bed Kg/m3 L/m2,s m2 °C % Kg pyrite/Kg Cu2S
Value 10 1.0.1014 0.001 2.8.10-3 2.5.10-9 20 0.63 2.0
Bioleaching Applications 1.3 1.2
Forced aeration, over-pressure 1000 Pa
Copper leaving per metre
1.1 1 0.9 Forced aeration, over-pressure 500 Pa
0.8 0.7
Natural convection
0.6 0.5 0.4
4
6
8
10 12 14 Channel distance (m)
16
18
20
Figure 4. Dissolved copper per metre of bottom for a 10m high heap for different distances between channels and different aeration rates
Figure 5. Relative concentration of oxygen at the edge of the heap, for natural convection The situation is somewhat different in the edge of the heap. Without aeration channels at the bottom the slope of the heap is well aerated, but the oxygen concentration decreases to the interior part of it. If the heap is equipped with aeration channels, the aeration is significantly improved. Figure 5 shows the situation for natural convection. For overpressure, the effect is greater. The pressure applied to the air in the aeration channel should not be too large, since, in this case the natural aeration is not utilised.
6.
CONCLUSIONS A two-dimensional model bioleaching of copper sulphide mineral in heap was developed. The model includes aeration by natural and forced convection. The emphasis was set in the impact of the aeration (distance between the aeration channels and the aeration rate) on the leaching process. The results show that the copper-leaching rate is increased when aeration is improved by using channels at the bottom of the heap. Even aeration channels without overpressure, natural convection, helps to enhance the leaching process. It was found that a large separation between the channels cause a deficient aeration in zones located between the channels. When the distance between channels is decreased, the aeration is improved, however, this improvement is less significant if one continues to decrease the distance 317
Bioleaching Applications
between channels. The optimal aeration rate is a function of several factors such as particle size, type of minerals, number of bacteria, and dimensions of the pile. However, the installation of aeration channels increases the equipment cost and the forced aeration increases the operation costs. The results may be used to determine the most favourable distance between channels and aeration rate for the leaching process.
REFERENCES 1. Bartlett, R.W., 1997. Metal extraction from ores by heap leaching, Metallurgical and Materials Transactions B-Process Metallurgy and Materials Processing Science, 28 (4), 529-545 2. Cathles, L.M., 1994. Attempts to model the industrial-scale leaching of copper-bearing mine waste, Environmental geochemistry of sulfide oxidation, ACS Symposium Series, 550, 123-131 3. Casas, J.M., Martinez, J., Moreno, L., Vargas, T., 1998. Bioleaching model of a copper-sulfide ore bed in heap and dump configurations, Metallurgical and Materials Transactions B-Process Metallurgy and Materials Processing Science, 29 (4), 899-909. 4. Coderre, F., Dixon, D.G., 1999, Modeling the cyanide heap leaching of cupriferous gold ores - Part 1: Introduction and interpretation of laboratory column leaching data, Hydrometallurgy, 52 (2), 151-175 5. Dixon, D.G., 2000, Analysis of heat conservation during copper sulphide heap leaching, Hydrometallurgy, 58 (1), 27-41 6. Lizama, H.M., 2001, Copper bioleaching behaviour in an aerated heap, Int. J. Miner. Process., 62 (1-4) 257-269. 7. Orr, S., 2002, Enhanced heap leaching – Part 1: insights, Mining Engineering, 54 (9) 49-56 8. Orr, S., Vesselinov, V., 2002, Enhanced heap leaching – Part 2: applications, Mining Engineering, 54 (10) 33-38 9. Ritchie, A.I.M., 1994. Rates of mechanisms that govern pollutant generation from pyritic wastes, Environmental geochemistry of sulfide oxidation ACS Symposium Series, 550, 108-122. 10. Sidborn, M., Casas, J., Martínez, J., Moreno, L., Two-dimensional dynamic model of a copper sulfide ore bed, Hydrometallurgy, In Press.
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Optimal oxygen and carbon dioxide concentrations for thermophilic bioleaching archaea S.H. de Kocka, K. Naldretta and C.A. du Plessisa a
BHP Billiton, Johannesburg Technology Centre, Private/Bag X10014, Randburg, 2125, South Africa Abstract Tank bioleaching reactors are currently sparged with air to satisfy both oxygen and carbon dioxide requirements in the reactors. Under high sulphide loading conditions, as is the case with high-grade concentrates, the microbial and chemical demand for oxygen is significantly increased during the bioleaching process in agitated tank reactors. Sparging with enriched oxygen gas may offer a potential solution in order to overcome the mass transfer difficulties at elevated temperatures. In the case of air sparging, the dissolved oxygen (DO) concentration in tank reactors could not be increased to a point where it would become inhibitory. The use of enriched oxygen in such reactors at large scale does, however, pose its own set of process risks. The first aim of this investigation was therefore, to determine the effects of various DO concentrations, in both the limiting and inhibitory ranges, on the microbial activity of Sulfolobus sp. U40813. Secondly, the effect of carbon dioxide concentration on the rate of ferrous iron oxidation was investigated. Both the oxygen and CO2 kinetics were examined in controlled batch cultures at 78°C, using ferrous sulphate and potassium tetrathionate as energy sources. The optimal DO concentration for iron oxidation was found to be between 1.5 to 4.1 mg.L-1. The use of elevated DO concentrations (above 4.1 mg.L-1) inhibited the ferrous oxidation rates. This inhibition effect increased progressively as the DO was increased above 4.1 mg.L-1. Due to the sensitivity of Sulfolobus to elevated dissolved oxygen, the used of oxygen-enriched air to overcome low solubilities in tank bioleaching reactors at high temperatures will have to be strictly controlled. The optimal CO2 concentration for ferrous iron oxidation is predicted to be between 7% and 17%. The iron oxidation rates were however, severely limited with CO2 concentrations less than 7%, indicating that the CO2 supply was limiting in this range and inhibited the microbial growth rate.
Keywords: Sulfolobus, oxygen, carbon dioxide, thermophilic, bioleaching, archaea 1.
INTRODUCTION Mineral bioleaching, the process by which metals are dissolved from sulphide orebearing rocks by microorganisms, is an established technology for metal recovery (Rawlings, 1997; Marsh et al., 1983). Although these technologies offer process and environmental advantages in certain instances, they are more prone to certain process upsets than non-biological hydrometallurgical extraction processes. One such important bioprocess condition is the requirement for a suitable dissolved oxygen concentration. 319
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Tank bioleaching reactors are currently sparged with air to satisfy both oxygen and carbon dioxide requirements in the reactors. Under high sulphide loading conditions, as is the case with high-grade concentrates, the microbial and chemical demand for oxygen is significantly increased during the bioleaching process in agitated tank reactors. The increased demand for oxygen under these conditions is facilitated by increased aeration rates, higher impeller agitation rates and improved agitator designs. Agitation speeds cannot however be indefinitely increased to improve mass transfer limitations, as cell damage to bioleaching microbes becomes a limiting factor at high agitation speeds and power inputs in the presence of high pulp densities. Mass transfer limitations are not routinely encountered in commercial tank leaching operations at mesophilic (35-40°C) conditions when sparging with air. At thermophilic (65-80°C) bioleaching conditions, as are required to leach primary copper sulphide minerals such as chalcopyrite, mass transfer limitation due to the reduced solubility of oxygen is, however, a significant process challenge that cannot be overcome by simply increasing agitation speeds and aeration rates. Sparging with enriched oxygen gas offers a potential solution in order to overcome the mass transfer difficulties at elevated temperatures. The use of enriched oxygen in such reactors at large scale does, however, pose its own set of process risks. In the case of air sparging, the dissolved oxygen concentration in tank reactors could not be increased to a point where it would become inhibitory. The use of oxygen-enriched air, however, could potentially result in the increase of dissolved oxygen concentrations as high as 15 mg.L-1. Such elevated dissolved oxygen concentrations could have a detrimental effect on microbial cells as the capacity to dissipate oxygen derived free radicals enzymatically could become overloaded. The concentration at which severe effects are encountered depends both on the species of microorganism as well as process and medium factors (Onken and Liefke, 1989). Su and Kelly (1987) investigated the effect of hyperbaric oxygen on the heterotrophic growth of Sulfolobus acidicaldarius. Elevated dissolved oxygen tensions were created through higher levels of air added to the gas phase while the total pressure was maintained at 50 bar. With agitation at 480 rpm, increasing the air partial pressure from one to two bar reduced the cell growth rates as well as the final cell density. The inhibition was even more severe at higher stirring rates. At 75°C, three bar air, or an oxygen partial pressure of 0.6 bar (estimated by Su and Kelly (1987) to be a dissolved oxygen tension of approximately 0.01 mg.L-1) resulted in the complete inhibition of growth. Besides oxygen, carbon dioxide is a critical component for growth of the bacteria to meet their carbon assimilation requirements. In contrast to bacteria (mesophiles and moderate thermophiles) where CO2 fixation occurs by means of the Benson Calvin cycle (Holuigue et al., 1987; Beudeker et al., 1980; Wood and Kelly, 1985), the high temperature mineral-oxidizing archaea such as Sulfolobus do not assimilate CO2 via the Calvin cycle. Kandler and Stetter (1981) showed that CO2 fixation in Sulfolobus brierleyi occurs via a reductive tricarboxylic acid pathway in which malic acid, aspartic acid, glutamic acid and citric acids play an important role. Enzyme assays by Wood et al. (1987) confirmed the absence of a Calvin cycle in Sulfolobus brierleyi and showed that a small proportion of CO2 fixation could occur through the carboxylation of pyruvate and phosphoenolpyruvate, but provided no further evidence for the proposed reductive carboxylic acid pathway believed to operate in Sulfolobus. Subsequent work has found enzyme activities that support the operation of a modified 3-hydroxypropionate pathway in Sulfolobus (now Acidianus) brierleyi and in other thermoacidophilic archaea 320
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(Metallosphaera sedula and Sulfolobus metallicus) (Burton et al., 1999; Ishii et al., 1997; Menendez et al., 1999). Nixon and Norris (1992) found that the yield in an air-sparged thiosulphate-limited continuous Sulfolobus culture (0.03% v/v CO2) was only 60% of that of a culture grown with 5% CO2 supplementation. In both cases however, the thiosulphate (20 mM in the feed) was completely oxidized by the cultures. Both Norris (1989) and Norris et al. (1989) reported that the rate of iron solubilization by Sulfolobus, in a pyritic medium, was reduced if the culture aeration was not enriched with CO2. Various studies focused on the effect of CO2 on the growth of mesophilic and moderate thermophilic bioleaching bacteria. Holuigue et al. (1987) found that an increase in the CO2 concentration from 0.03% to 0.1% caused a reduction of more than half in the doubling time of A. ferrooxidans. A further increase to 5.4% CO2 did not result in a further increase in the growth rate, although it seemed to enhance the total yield of bacteria as compared with 0.1% CO2. Hazeu et al. (1986) supplemented the air supply to a thiosulphate chemostat culture with A. ferrooxidans with 2% CO2 (v/v) and also observed a slight increase in the yield. The optimal CO2 concentration for growth of A. ferrooxidans during the bioleaching of a pyrite-arsenopyrite ore concentrate was predicted to be 3 to 7 mg.L-1 (estimated to be a about 0.23 to 0.53% CO2). The study also showed that CO2 concentrations below the optimal levels lead to sharply reduced bacterial growth rates, whereas CO2 concentrations in excess of 10 mg.L-1 (estimated to be about 0.73% CO2) were inhibitory to the growth of A. ferrooxidans (Nagpal et al., 1993). Torma et al. (1972) observed that zinc extraction rates from a zinc sulphide mineral increased from 360 mg.L-1.h-1 to 640 mg.L-1.h-1 when the CO2 concentration in air was increased to 0.23%. However, increasing the CO2 further made no difference to extraction rates. In contrast, Norris (1989) and Norris et al. (1989) both found that the rate of pyrite dissolution of A. ferrooxidans was only slightly reduced by the use of air without additional CO2. At moderate thermophilic temperatures, both the growth rate on ferrous iron (Wood and Kelly, 1985) and pyrite dissolution rate (Norris an Owen, 1993) were enhanced with CO2 supplementation. A number of significant problems exist when interpreting literature results in which CO2 supplementation was tested as CO2 supplementation cannot directly or even indirectly be related to dissolved CO2 concentrations. Unlike for dissolved oxygen, no reliable dissolved CO2 measuring instruments are available, particularly at high temperatures. Microbial cells in solution are only exposed to dissolved CO2 concentrations. From existing literature no such measurements or comparisons can therefore be made. Furthermore, the interpretation of whether CO2 supplementation was beneficial is often skewed by the fact that it did not result in improved bioleaching rates of the mineral under consideration. In cases where mineral dissolution effects are ratelimiting, CO2 supplementation, and the resultant increase microbial growth rates and yields will not have an improved bioleaching effect. For this reason, the effect of CO2 supplementation on microbial growth per se (and the potential use of CO2 supplementation in bioleaching scenarios where mineral dissolution rates are not the limiting factor), has probably been overlooked in many instances. Another important factor that may influence the interpretation of the results is that some bacteria, for example, have been found to exhibit supplementary carbon fixing ability. According to this mechanism, some bacteria would switch to a different, more efficient, carbon fixing mechanism in environments where dissolved CO2 is limiting (Norris, 1989). The aim of this investigation was to determine the effects of various dissolved oxygen and CO2 concentrations, in both the limiting and inhibitory ranges, on the microbial 321
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activity of a representative thermophilic bioleaching archaea. Analyses of numerous continuous culture pilot scale tank reactors at laboratory facilities of BHP Billiton Johannesburg Technology Centre had revealed Sulfolobus sp. U40813 to be the most common mineral oxidizer across a wide range of mineral types at 78°C. This investigation, therefore, focuses on determining oxygen and CO2 kinetics effects for this particular archaea at typical thermophilic tank bioleaching conditions.
2.
MATERIALS AND METHODS
2.1 Microbial inoculum The Sulfolobus sp. U40813 microbial inoculum used was maintained in an 8-L "fedbatch type" bioleaching mini-plant treating a chalcopyrite concentrate at 78°C under nonsterile conditions. Two litres of the reactor slurry was removed daily and replaced with fresh solids and nutrients. Mineral particles in the samples taken from the mini-plant were allowed to settle gravitationally for approximately 30 minutes. The resulting supernatant, with suspended cells, was used as inoculum. The microbial cell concentration in the inoculum were determined using a CellFacts biological particle analyzer (Microbial Systems Limited, UK), and were in the range of 4.4 x 108 cells.mL-1 to 5.2 x 108 cells.mL1 . A 100 mL microbial inoculum was added to 900 mL culture medium at the start of each run. The microbial population in the bioreactor was identified using molecular microbiology techniques, including PCR amplification and denaturing gradient gel electrophoresis (DGGE). The sample was submitted to a series of low- and high-speed centrifugation steps using saline (0.856% m/v NaCl, pH 1.2) to remove debris and precipitates. DNA was extracted using the High Pure PCR template Preparation kit (ROCHE, Johannesburg, South Africa). A fragment of the 16S rRNA gene was amplified by PCR using primers annealing on either side of a variable region on the 16S rRNA gene of thermophilic archaea associated with the bioleaching of minerals. The PCR fragment was submitted to DGGE together with PCR fragments of known thermophilic archaea, which served as standards. The PCR product migrated in the same way as a Sulfolobus sp. (GenBank accession no. U40813). 2.2 Culture medium and growth conditions Batch cultivations were carried out in a 1.5-L glass vessel, placed on a stirring hotplate with temperature feedback control and mixed with a magnetic stirrer bar at approximately 1000 rpm. The nutrient medium used in all the batch cultivations was the 9K medium supplemented with 3 g.L-1 potassium tetrathionate (Fluka, Steinheim, Switzerland) and 1 mL.L-1 of a trace element solution. The 9K medium comprised (per litre tap water): (NH4)2SO4, 3 g; K2HPO4, 0.5 g; KCl, 0.1 g; Ca(NO3)2, 0.01 g; FeSO4.7H2O, 50 g (all ACE, Glenvista, South Africa) and MgSO4.7H2O, 0.5 g (Saarchem, Johannesburg, South Africa). The final concentration of the trace elements in the nutrient medium comprised (per litre); MnCl2.4H2O, 1.80 mg; Na2B4O7.10H2O, 4.50 mg; ZnSO4.7H2O, 0.22 mg; CuCl2.2H2O, 0.05 mg; Na2MoO4.2H2O, 0.03 mg; VOSO4.2H2O, 0.03 mg and CoSO4, 0.01 mg. The pH of the medium was initially adjusted between 1.5 and 1.55 with sulfuric acid and controlled at 1.5 during the cultivation with 4 N NaOH. Temperature was maintained throughout the experiments at 78°C. The redox potential was determined at 10 minute intervals using a combined electrode (Pt-Ag/AgCl in 3N KCl) connected to a 718 pH STAT Titrino (Swiss Lab Ltd, Rivonia, South Africa).
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Dissolved oxygen and CO2 was controlled by varying the proportions of nitrogen, carbon dioxide and oxygen in the influent gas through three 58505 S Brooks mass flow controllers (Alpret Control Specialists Ltd, Florida, South Africa). The gas mixture was sparged at a constant total flow rate of 1.5 L.min-1. The dissolved oxygen (DO, mg.L-1) concentration was measured with a polarographic oxygen probe (Knick, Berlin, Germany). An inlet CO2 concentration of 5% was used in the oxygen cultivations, whereas the DO was maintained at 3 mg.L-1 in the CO2 experiments.
3.
RESULTS AND DISCUSSION Although no specific precautions were taken to ensure aseptic conditions inside the bioreactor, the microbial population did not change during the duration of the experiments, due to the very selective culture conditions within the bioreactor. Prior shake-flask cultivations in the laboratory (data not shown) indicated that Sulfolobus sp. U40813 was not able to grow using sulphur in the form of potassium tetrathionate as sole energy source. Complete oxidation was, however, achieved using ferrous sulphate as sole energy source. The effect of increasing dissolved oxygen concentrations on the rate of iron oxidation was investigated in a series of batch cultivations, each controlled at a different dissolved oxygen concentration. The microbial activity in each test was evaluated by monitoring the increase in redox potential and Fe3+ (g.L-1) over a 25 to 30 h period. Due to rapid precipitation on the redox probes at the high temperatures, the redox potential readings were not very reliable and, therefore, excluded from comparative interpretation of the effect of DO on microbial activity. Although the conditions inside the mini-plant (supplying the inoculum) were kept as constant as possible by daily additions of fresh solids and nutrients, the microbial activity of the inoculum was not identical for each batch cultivation. In order to overcome variable inoculum activity, batch cultivations were conducted in sets of three, with a reference test (at a DO of 2.4 mg.L-1) included in every set of experiments. For every set of experiments, the maximum iron oxidation rate of the reference test was given a relative activity value of one, and the relative activity of the remaining two runs were expressed as a fraction of the maximum iron oxidation rate of the reference test. A relative activity of 1 is, therefore, equal to an iron oxidation rate of 0.524 (±0.081) g.L-1.h-1. An example of a typical set of batch cultivations results is shown in Figure 1. The relative activities of the Fe3+ production curves at different dissolved oxygen concentrations are given in Figure 2. The relative activity data points were fitted using a four-parameter log normal fit function (Equation 1), where y and x are the relative activity and DO (mg.L-1), respectively. 2 ⎡ ⎛ ⎛ x ⎞⎞ ⎤ ⎜ ln⎜ ⎢ ⎟⎟ ⎥ 2.5077 ⎠ ⎟ ⎥ ⎝ ⎜ ⎢ y = 0.1072 + 0.9029l − 0.5 ⎜ 0.9908 ⎟ ⎥ ⎢ ⎜ ⎟ ⎥ ⎢ ⎝ ⎠ ⎦⎥ ⎣⎢
(1)
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10 2.4 mg.L-1
Fe3+ (g.L-1)
8
6
-1
8.2 mg.L
9.0 mg.L-1 4
2
0 0
5
10
15 Time (h)
20
25
30
Figure 1. The Effect of a Dissolved Oxygen Concentration of 2.4 mg.L-1 (●), 8.2 mg.L1 (○) and 9.0 mg.L-1 (▼) on the Fe3+ Concentration During Autotrophic Growth of Sulfolobus sp. U40813 1.4
Inlet Oxygen Concentration (%, v/v) 20 40 60 80
0
R2=0.92
1.2
100
R2=0.92
Relative activity
1.0 0.8 0.6 0.4 0.2 0.0 0
1
2
3 4 5 6 7 Dissolved oxygen (mg.L-1)
8
9
10
Figure 2. The Effect of Dissolved Oxygen Concentrations on the Iron Oxidation Activity During Autotrophic Growth of Sulfolobus sp. Approximated by the fitted curve (Figure 2), a dissolved oxygen concentration between 1.5 and 4.1 mg O2.L-1 is required for optimal microbial activity. Within these ranges, the microbial activity was between 90% and 100% of the maximum activity obtained in the experiments. With an increase or decrease in the dissolved oxygen concentration outside this range, the microbial activity was affected negatively. The inhibitory effect was more severe at the very low dissolved oxygen concentrations (Figure 2) with a microbial activity of less than 50% of the maximum microbial activity resulting from a dissolved oxygen concentration of 0.7 mg.L-1. However, although the inhibition at 324
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the high oxygen concentrations (at and above a DO of 7.2 mg.L-1) (Figure 2) was not initially as severe as at the low oxygen concentrations, prolonged exposure to high oxygen concentrations resulted in a continual decrease in activity, which eventually resulted in a cessation of activity. To investigate the effect of increased CO2 concentrations of the microbial activity, a series of batch cultivations in sets of three (similar as for the oxygen work) was conducted with increasing CO2 concentrations. An example of a typical set of batch cultivations results is shown in Figure 3. The relative activities for the individual experiments were determined as for the oxygen work, and are shown in Figure 4. A relative activity of 1 in this case is equal to 0.434 (±0.086) g Fe3+.L-1.h-1.
Fe3+ (g.L-1)
12 10
11%
8
5%
6 0.03%
4 2 0 0
5
10
15 Time (h)
20
25
30
Figure 3. The Effect of an inlet CO2 Concentration of 0.03% (▼), 5% (●) and 11% (○) on the Fe3+ Concentration During Autotrophic Growth of Sulfolobus sp. U40813 1.4
R2=0.96
Relative activity
1.2 1.0 0.8 0.6 0.4 0.2 0
5
10 15 20 Inlet CO2 Concentration (%, v/v)
25
30
Figure 4. The Effect of CO2 on the Iron Oxidation Activity During Autotrophic Growth of Sulfolobus sp. U40813 325
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The relative activity data points were fitted using a four-parameter log normal fit function (Equation 2), where y and x are the relative activity and inlet CO2 concentration, respectively. 2 ⎡ ⎛ ⎛ x ⎞⎞ ⎤ ⎜ ln⎜ ⎢ ⎟⎟ ⎥ 11.02 ⎠ ⎟ ⎥ ⎝ ⎜ ⎢ y = 0.3450 + 0.8298l − 0.5 ⎜ 1.147 ⎟ ⎥ ⎢ ⎜ ⎟ ⎥ ⎢ ⎝ ⎠ ⎦⎥ ⎣⎢
(2) Approximated by the fitted curve, 95% or more of the maximum relative activity was achieved with inlet CO2 concentrations between 7% and 17%. With a further increase in the CO2 concentration (above a concentration of 17%), the relative activities decreased slightly, but remained above 85% of maximum relative activity. The iron oxidation rates were severely limited with a CO2 concentration less than 7%, indicating that the CO2 supply was limiting in this range and inhibited the microbial growth rate.
4.
CONCLUSIONS Oxygen and carbon dioxide are essential nutrients for the growth of mineral-leaching bacteria. Maintaining optimal dissolved oxygen and CO2 concentrations will lead to improved mineral oxidation process and bacterial growth rates. This study indicated that the optimal dissolved oxygen concentration for growth was between 1.5 to 4.1 mg.L-1. At dissolved oxygen concentrations below 1.5 mg.L-1, oxygen was the growth-limiting nutrient and inhibited the ferrous utilization rate. The use of elevated dissolved oxygen concentration (above 4.1 mg.L-1) also reduced the ferrous oxidation rates. This inhibition effect progressively increased as the DO was increased above 4.1 mg.L-1. Due to the sensitivity of Sulfolobus to elevated dissolved oxygen, the used of oxygen-enriched air to overcome low solubilities in tank bioleaching reactors at high temperatures will have to be strictly controlled in the narrow range. The optimal inlet CO2 concentration for ferrous iron oxidation is predicted to be between 7% and 17%. Above 95% of the maximum relative activity was obtained in these ranges. At elevated CO2 levels the ferrous oxidation rate was slightly lower, but remained above 85% of maximum relative activity. The iron oxidation rates were however, severely limited with a CO2 concentration less than 7%, indicating that the CO2 supply was limiting in this range and inhibited the microbial growth rate. Although this is much higher than is generally considered adequate for bioleaching, it is nevertheless the concentration that will allow for optimal microbial growth rates. The use of such a high concentration of CO2 would only be beneficial in cases where mineral dissolution is not a limiting factor, and where the additional cost could be justified against process improvement. Based on the results from this investigation, it is suggested that tank bioleaching conditions should be maintained at a dissolved oxygen concentration of 2.8 mg.L-1 and that strict process control mechanisms should be in place to ensure that DO concentrations in tanks are maintained in the narrow range of 1.5 to 4.1 mg.L-1. REFERENCES 1. Beudeker, R.F., Cannon, G.C., Kuenen, J.G. and Shively, J.M. (1980). Relations between D-Ribulose-1,5-Bisphosphate carboxylase, carboxysomes and CO2 fixing capacity in the obligate chemolithotroph Thiobacillus neapolitanus grown under different limitations in the chemostat. Arch Microbiol 124: 185-189. 326
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2. Burton, N.P., Williams, T.D. and Norris, P.R. (1999). Carboxylase genes of Sulfolobus metallicus. Arch Microbiol 172: 349-353. 3. Hazeu, W., Bijleveld, W., Grotenhuis, J.T.C., Kakes, E. and Kuenen, J.G. (1986). Kinetics and energetics of reduced sulfur oxidation by chemostat cultures of Thiobacillus ferrooxidans. Antonie van Leeuwenhoek 52: 507-518. 4. Holuigue, L., Herrera, L., Phillips, O., Young, M. and Allende, J. (1987). CO2 fixation by mineral-leaching bacteria: characteristics of the ribulose bisphosphate carboxylaseoxygenase of Thiobacillus ferrooxidans. Biotechnol Appl Biochem 9: 497-505. 5. Ishii, M., Miyake, T., Satoh, T., Sugiyama, H., Oshima, Y., Kodama, T. and Igarashi, Y. (1997). Autotrophic carbon dioxide fixation in Acidianus brierleyi. Arch Microbiol 166: 368-371. 6. Kandler, O. and Stetter, K.O. (1981). Evidence for autotrophic CO2 assimilation in Sulfolobus brierleyi via a reductive carboxylic acid pathway. Zbl. Bakt. Hyg. I. Abt. Orig C2: 111-121. 7. Marsh, R.M., Norris, P.R. and Le Roux, N.W. (1983). Growth and mineral oxidation studies with Sulfolobus. Progress in Biohydrometallurgy May: 71-81. 8. Menendez, C., Bauer, Z., Huber, H., Gad’on, N., Stetter, K-O. and Fuchs, G. (1999). Presence of acetyl coenzyme A (CoA) carboxylase and propionyl-CoA carboxylase in autotrophic Crenarchaeota and indication for operation of a 3-hydroxypropionate cycle in autotrophic carbon fixation. J Bacteriol 181: 1088-1098. 9. Nagpal, S., Dahlstrom, D. and Oolman, T. (1993). Effect of carbon dioxide concentration on the bioleaching of a pyrite-arsenopyrite ore concentrate. Biotechnol Bioeng 41: 459-464. 10. Nixon, A. and Norris, P.R. (1992). Autotrophic growth and inorganic sulphur compound oxidation by Sulfolobus sp. in chemostat culture. Arch Microbiol 157: 155160. 11. Norris, P.R. (1989). Factors affecting bacterial mineral oxidation: the example of carbon dioxide in the context of bacterial diversity. Biohydrometallurgy, 3-14. 12. Norris, P., Nixon, A. and Hart, A. (1989). Acidophilic, mineral-oxidizing bacteria: the utilization of carbon dioxide with particular reference to autotrophy in Sulfolobus. In: Microbiology of Extreme Environments and its potential for biotechnology (ed. M.S. da Costa, J.C. Duarte and R.A.D. Williams). Elsevier, London, pp. 24-43 13. Norris, P.R. and Owen, J.P. (1993). Mineral sulphide oxidation by enrichment cultures of novel thermoacidophilic bacteria. FEMS Microbiol Reviews 11: 51-56. 14. Onken, U. and Liefke, E. (1989). Effect of total and partial pressure (oxygen and carbon dioxide) on aerobic microbial processes. Adv Biochem Eng Biotechnol 40: 137-169. 15. Rawlings, D.E. (1997). Biomining. Theory, microbes and industrial processes. New York: Springer-Verlag. 16. Su, W-W. and Kelly, R.M. (1987). Effect of hyperbaric oxygen and carbon dioxide on heterotrophic growth of the extreme thermophile Sulfolobus acidocaldarius. Biotechnol Bioeng 31: 750-754. 17. Torma, A.E., Walden, D.D., Duncan, D.W. and Branion, R.M.R. (1972). The effect of carbon dioxide and particle surface area on the microbiological leaching of a zinc sulfide concentrate. Biotechnol Bioeng 14: 777-786. 18. Wood, A.P. and Kelly, D.P. (1985). Autotrophic and mixotrophic growth and metabolism of some moderately thermoacidophilic iron-oxidizing bacteria. In: Planetary Ecology (ed. D.E. Caldwell, C.L. Brierley and J.A. Brierley), Van Nostrand Rheinhold, New York, pp. 251-262.
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19. Wood, A.P., Kelly, D.P. and Norris, P.R. (1987). Autotrophic growth of four Sulfolobus strains on tetrathionate and the effect of organic nutrients. Arch Microbiol 146: 382-389.
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Optimization study on bioleaching of municipal solid waste (MSW) incineration fly ash by Aspergillus niger T.J. Xu and Y.P. Ting* Department of Chemical & Environmental Engineering, National University of Singapore Kent Ridge Crescent, Singapore 119260
Abstract The bioleaching efficiency of heavy metals from MSW incineration fly ash depends on physical, chemical and biological factors as well as the leaching environment. Our research objective is to determine the factors that have a greater impact on the bioleaching process, as well as to determine the optimal bioleaching conditions and maximum leached metal concentrations. Four factors were investigated in this study: sucrose concentration, inoculum spore concentration, fly ash pulp density, and the time of addition of fly ash to the fungus Aspergillus niger. The Central Composite Design (CCD) was used in order to determine the co-optimum level of the factors, as well as to provide an insight into the interactions amongst these factors during bioleaching. Empirical models obtained through 2nd order (Taylor series approximation) regression provided the optimal bioleaching conditions. Results showed that sucrose concentration and pulp density were more important factors than spore concentration and the time of addition of the fly ash. 1.
INTRODUCTION Bioleaching processes are based on the ability of microorganisms to transform solid compounds, and result in soluble and extractable elements which can be recovered [1]. Bioleaching is affected by a number of parameters; its effectiveness is highly dependent on physical, chemical and biological factors such as (i) nutrient, (ii) oxygen and carbon dioxide supply, (iii) pH, (iv) temperature of leaching environment, (v) pre-culture period and inoculum used, (vi) resistance of microorganisms to metal ions, (vii) physical and chemical states of the solid residue, (viii) liquid-solid ratio, and (ix) bioleaching period [2]. Maximum yield of metal leaching can be achieved when these parameters have been optimized. Most of the previous work on fungal bioleaching has been conducted using a "onefactor-at-a-time" technique. Unfortunately, this method fails to locate the region of optimum response, since the "one-factor-at-a-time" technique does not take into account any joint factor interactions on the bioleaching process. An alternative approach is the Response Surface Method (RSM), which simultaneously considers several factors at many different levels and determines the corresponding interactions among these factors using a smaller number of experimental observations. RSM has been employed to solve multivariate problems and optimize several responses in many types of investigations [3]. * Corresponding author:
[email protected]
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In this study the RSM approach is adopted to locate the co-optimum levels of sucrose concentration, fly ash pulp density, spore concentration (i.e. inoculum concentration), and the time of fly ash addition, and to gain an insight into the interactions among these factors during the bioleaching process using Aspergillusn niger. Using such an approach, the optimal of these four factors and the maximum concentration of the metal leached may be obtained.
2.
MATERIAL AND METHOD
2.1 Characteristics of local fly ash A locally-produced municipal solid wastes (MSW) incineration fly ash (at the Tuas South Incineration Plant, Singapore) was used in this optimization study. Some physical and chemical characteristics of the fly ash are shown in Tables 1 and 2 respectively. Table 1. Physical characteristics of MSW fly ash Particle size Mean (um) Particle size Median (um) Bulk density (g/cm3) Specific gravity (g/cm3) Surface area (m2/g) Total pore volume (cm3/g) Porosity (cm3/cm3)
26.33 15.62 0.4 3.1 5.75 0.01927 0.0062
Table 2. Chemical characteristics of MSW fly ash Concentration (mg/kg) Al 1,860 Ca 404,000 Cu 326 Fe 2,200 K 11,500 RSD: relative standard deviation Element
RSD
Element
0.17% 0.14% 1.3% 0.65% 2.29%
Mg Mn Pb Sr Zn
Concentration (mg/kg) 9,110 71 2,070 185 7,890
RSD 0.19% 3.4% 2.89% 0.39% 1.38%
2.2 Fungi inoculum preparation Aspergillus niger was obtained from Dr H. Brandl (University of Zürich, Switzerland) and was cultured according to the protocol in Bosshard et al. (1996) [4]. The number of spores was enumerated under a microscope (Olympus CX40) at 400x magnification using 0.1 mm depth haemocytometer (SUPERIOR MARIENFELD). To culture in liquid medium, 1ml of spore suspension was added to 100 ml of standard sucrose medium [4] with composition (g/L): 100 sucrose (Biorad), 1.5 NaNO3 (Merck), 0.5 KH2PO4 (Merck), 0.025 MgSO4.7H2O (Merck), 0.025 KCl (Merck), 1.6 yeast extract (Difco), and incubated in an incubator at 30ºC with rotary shaking at 120 rpm. All reagents were of analytical grade. The liquid medium was autoclaved at 121°C for 15 minutes prior to inoculation. Bioleaching was performed in 250ml Erlenmeyer flasks with 100ml of sucrose medium.
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2.3 Analytical method Inductive coupled plasma-atomic emission spectrometry (ICP-AES) was used to analyze the metals in the bioleaching process. The concentration of glucose was measured using YSI 2700 biochemistry analyzer. Fructose, sucrose, citrate, oxalate and gluconate were analyzed using HP 1100 series high performance liquid chromatography (HPLC) with variable wavelengths detector (VWD) at 210 nm for organic acids detection and refractive index detector (RID) for fructose and sucrose detection. Operation conditions for HPLC consisted of a 30 cm x 7.8 mm i.d. Biorad Aminex HPX- 87H hydrogen resin ionic form analytical column (9 µm particle size). The analysis was carried out at temperature of 30°C. The mobile phase used was 5mM sulfuric acid (Merck, analytical grade), at flow rate of 0.5 ml/min. 2.4 Biomass determination In the pure fungal culture, the mycelium was dried at 80°C for 24 hours. For biomass determination in the bioleaching tests, the mycelium together with fly ash, after filtration, was dried at 80°C for 24 hours, followed by ashing at 500°C for 4 hours in a Carbolite CWF 1100 furnace. Biomass was calculated gravimetrically after cooling in a dessicator [5].
2.5 Software for experimental design, statistical analysis, and optimization The Central Composite Design, statistical analysis of the data, and the development of the regression equations were performed using the MINITAB package (Minitab Inc.). Generalized Algebraic Modeling System (GAMS, GAMS Development Corp.) was used to optimize the 2nd order statistical empirical models (see Equation (2)-(7) later). 3.
RESULTS AND DISCUSSION In bioleaching, the optimal solid-liquid ratio (pulp density) has to be determined. In fly ash fungal bioleaching, the fly ash is toxic to the fungus. Although an increase in pulp density may lead to a decrease in leaching efficiency, the metal concentration in the leachate may still increase. Too high a pulp density however will lead to toxicity and result in poor bioleaching. Hence, there must be a critical (optimal) value for the pulp density, which will result in the maximum metal concentration. For this reason, the pulp density and time of addition of fly ash are considered in this study. Sucrose was selected as another factor since it is the precursor of organic acid, which is the most important leaching agent in the bioleaching process. The choice of the inoculum has an important impact on bioleaching efficiency. Hence, the spore concentration was investigated in this study.
Table 3. Parameter levels of central composite design (coded value) Run Order
Sucrose conc.
Spore conc.
Pulp Time of density addition
Run Order
Sucrose conc.
Spore conc.
1
1
-1
-1
2
0
0
3
1
4 5
Pulp Time of density addition
1
17
2
0
0
0
0
0
18
1
-1
1
-1
1
-1
1
19
0
0
0
0
1
1
1
-1
20
-1
1
1
-1
1
1
-1
-1
21
-1
-1
1
1 331
Bioleaching Applications Run Order
Sucrose conc.
Spore conc.
Pulp Time of density addition
Run Order
Sucrose conc.
Spore conc.
6
0
0
0
7
0
0
8
1
9
Pulp Time of density addition
0
22
-1
-1
-1
1
0
-2
23
-1
1
1
1
1
1
1
24
0
0
0
2
1
-1
1
1
25
0
2
0
0
10
0
0
0
0
26
0
0
2
0
11
-1
-1
-1
-1
27
-1
-1
1
-1
12
-1
1
-1
-1
28
-2
0
0
0
13
0
0
0
0
29
0
0
-2
0
14
1
-1
-1
-1
30
0
-2
0
0
15
0
0
0
0
31
-1
1
-1
1
16
0
0
0
0
Table 4. Quantitative value of the coded parameter levels Coded factor
Parameter
-2
-1
0
1
2
X1
Sucrose concentration, g/l
60
90
120
150
180
7
X2 X3
Spores concentration, *10
0.3
0.8
1.3
1.8
2.3
Pulp density, w/v%
0.1
0.3
0.9
2.7
8.1
X4
Time of addition, days
0
4
8
12
16
Bioleaching were conducted for duration of 26 days.
3.1 Central Composite Design (CCD) The CCD design for this study consists of a 24 (i.e. 4 factors) full factorial design, with 2*4 axial points at (±α, 0, 0,…, 0), (0, 0, ±α,…, 0),…, (0, 0, 0,…, ±α), and 7 central points at (0, 0,…, 0), where α is the distance of the axial point from the center [3]. Random error (standard deviation) can be estimated from the 7 central points. For a fourfactor design, α is usually set at 2.0 [3], since the distance of the axial points from the center point is given by α=2n/4 [6]. Therefore a total of thirty-one (16 full factorial tests + 8 axial points + 7 central points) batch bioleaching tests were performed to satisfy a central composite design. Table 3 and Table 4 describe how each parameter was varied in the batch tests. The data collected from the 31-batch runs were used to develop empirical models describing the experimental results. The models were generated using the method of least squares. The technique involved the estimation of model parameters for second order models of the form: k
k
k
E (Y ) = β 0 + ∑ β i X i + ∑∑ β ij X i X j
(1) where E(Y) is the expected value of the response variable, βο, βi, βj are the model parameters, Xi and Xj are the coded factors being studied and k is the number of factors being studied [3]. In this study, E(Y) represents biomass concentration, metal concentration, and organic acid concentration for the different empirical model. i =1
332
i =1 j =1
Bioleaching Applications
Numerous factors have varying impact on the bioleaching process. The relative importance of the four factors considered in this study, i.e. sucrose concentration, inoculum spore concentration, fly ash pulp density, and the time of addition of fly ash is manifested through the magnitude of the model coefficients, i.e. βi, βj.
3.2 Data analysis It is necessary to perform an analysis of the residual from the model in order to determine the adequacy of the least squares fit. A normal probability of the residuals, and the residuals versus predicted values of the response variable are constructed to verify that the data follow a normal distribution [3]. The analysis of variance (ANOVA) table was used to analyze these data. 3.3 Models for Biomass, Al, Fe, Zn, citric acid and gluconic acid concentration Equations (2)-(7) were obtained from the 31-batch runs using the MINITAB software. The data followed a straight line in the normal probability plot of the residual, thus representing a normal distribution and supporting the assumptions of the empirical model. The plot of the residuals versus predicted values of the response variable also supported the assumption of a normal distribution. 3.3.1 Biomass concentration model Equation (2) shows that all the four factors, i.e. sucrose concentration (X1), spore concentration (X2), pulp density (X3), and the time of addition of the fly ash (X4) have an important impact on the biomass concentration in bioleaching process. Equation (2) also shows an interaction between sucrose concentration (X1) and spores concentration (X2), sucrose concentration (X1) and pulp density (X3), sucrose concentration (X1) and the time of addition (X4), spores concentration (X2) and pulp density (X3), and pulp density (X3) and the time of addition (X4). Biomass = 2.45 + 0.0131X1 − 0.2312X2 − 0.758X3 + 0.1305X4 − 0.3304X12 − 0.1002X22 −0.1254X32 − 0.1485X24 − 0.1741X1X2 − 0.2637X1X3 + 0.2801X1X4 − 0.1049X2X3 − 0.1882X3X4
(2) It can be concluded that the sucrose concentration had a significant interaction with the spore concentration, pulp density and time of addition in the production of biomass, since sucrose is the only source of carbon in the process (and which is a limiting nutrient in the biomass production).
3.3.2 Al concentration model Equation (3) shows that all the four factors, i.e. sucrose concentration (X1), spore concentration (X2), pulp density (X3), and time of addition (X4) exert an important influence on Al concentration in the bioleaching process. Equation (3) also shows an interaction between sucrose concentration (X1) and pulp density (X3), spores concentration (X2) and pulp density (X3), and spores concentration (X2) and time of addition (X4).
Al = 5.58 + 1.56X 1 + 0.0564X 2 + 3.34X 3 + 0.115X 4 − 0.496X 12 − 0.181X 22 + 0.436X 32 − 0.182X 42 + 1.80X 1 X 3 + 0.155X 2 X 3 − 0.127X 2 X 4
(3) The model showed that pulp density had a significant interaction with the sucrose concentration and spore concentration in the leaching of Al. 333
Bioleaching Applications
3.3.3 Fe concentration model Equation (4) shows that the spore concentration (X2) over the range investigated was not a factor in the leaching of Fe. The equation also shows that the pulp density (X3) had a significant interaction with the sucrose concentration (X1), and that there was only one interaction amongst the four factors, i.e. between sucrose concentration (X1) and the pulp density (X3). Fe = 5.73 + 2.06 X 1 + 3.24 X 3 − 0.317 X 4 − 0.89 X 12 − 0.387 X 42 + 2.05 X 1 X 3
(4)
3.3.4 Zn concentration model Equation (5) shows that in the leaching of zinc, there was an interaction between sucrose concentration (X1) and pulp density (X3), between sucrose concentration (X1) and the time of addition (X4), and between spores concentration (X2) and pulp density (X3).
Zn = 35.69 + 11.89 X 1 + 1.27 X 2 + 18.64 X 3 + 0.488X 4 − 1.89 X 12 − 1.58 X 22 − 2.7 X 42 + 11.13X 1 X 3 + 1.239 X 1 X 4 + 2.15 X 2 X 3
(5) The pulp density had a significant interaction with the sucrose concentration in the Zn leaching, as the only source of Zn in the process is the fly ash. Sucrose is converted to citric acid and gluconic acid, which are the most important leaching agents in bioleaching process.
3.3.5 Citric acid concentration model Equation (6) shows that there was no interaction between any of the four factors. Indeed, it is surprising that no interaction between sucrose concentration and pulp density was observed. Our results are similar to that of Crolla et al. [3], where there were no interactions between the various factors, and only linear and square terms were obtained [16].
Citric= 47.0 + 9.08X1 − 0.928X 2 − 2.97X 3 + 3.34X 4 − 7.85X12 − 8.39X 22 − 4.58X 32 − 6.73X 42
(6)
3.3.6 Gluconic acid concentration model Equation (7) shows that there was an interaction between sucrose concentration (X1) and pulp density (X3), and sucrose concentration (X1) and time of addition (X4). Sucrose concentration had a significant interaction with the pulp density. Low pH (<3) inhibits the activity of the enzyme glucose oxidase. It is known that at pH > 3, the presence of fly ash presence activates glucose oxidase and results in the greater production of gluconic acid [7]. Gluconic= 27.3 + 69.4X1 −1.5X 2 + 37.8X 3 −10.3X 4 + 28.7 X12 + 3.84X 22 + 20.6X 32 + 6.37X 42 + 20.8X1 X 3 − 5.82X1 X 4
(7)
3.4 Summary Equations (2)-(7) are objective functions. In general, it can be seen from these equations that the coefficients of sucrose concentration (X1) and pulp density (X3) were larger than those of spore concentration (X2) and time of addition (X4). Since a higher coefficient represents greater importance, this optimization study showed that sucrose concentration and pulp density were more important than spore concentration and the time of addition of the fly ash. 334
Bioleaching Applications
The optimized values, obtained by the maximization of these objective functions are summarized in Table 5. The table also shows that gluconic acid (at 281mM) is the main bioleaching agent and is produced under similar optimal condition as the optimised values of the leached metals. The optimal time for the addition of fly ash for gluconic acid production is 0 days. This is consistent with the fact gluconic production is optimal at a high pH [7] (since the earlier the time of addition, the higher the pH). The optimal the time of addition of the fly ash for metal leaching varied from 6 to 9 days. The optimal sucrose concentration for metal leaching and gluconic acid production is about 150 g/l, a value 50% higher than Bosshard’s medium. The corresponding optimal pulp density is 2.7%.
Table 5. Optimal value for bioleaching process Parameter
Sucrose Conc (g/l)
Spores Conc. (*107/ml)
Pulp density (%)
Time of addition (days)
Maximum Value
Biomass
153
0.8
0.1
16
50.3 g/l
Al
150
1.6
2.7
8.5
12.3 mg/l
Fe
150
2.7
6.6
12.3 mg/l
Zn
150
1.8
2.7
9.3
77.6 mg/l
Citric
108
1.3
0.6
9
51mM
Gluconic
150
0.3
2.7
0
281mM
0.3-2.3
4.
CONCLUSSIONS Four factors (i.e. sucrose concentration, spore concentration, pulp density and the time of fly ash addition) were investigated in this bioleaching study. The Central Composite Design (CCD) was used in order to determine the co-optimum level of the factors, as well as to provide an insight into the interactions amongst these factors during bioleaching. Empirical models obtained through 2nd order (Taylor series approximation) regression provided the optimal bioleaching conditions. Results showed that sucrose concentration and pulp density were more important factors than spore concentration and the time of addition of the fly ash. ACKNOWLEDGEMENTS This work was funded by the National University of Singapore research grant R-279000-059-112. The authors thank Dr H. Brandl (University of Zürich, Switzerland) for providing the A. niger. REFERENCES 1. W. Krebs, C. Brombacher, P.P. Bosshard, R. Bachofen and H. Brandl, Microbial Recovery of Metals from Solids, FEMS Microbiol. Rev., 20, 605-617, 1997. 2. K. Bosecker, Bioleaching: Metal Solubilization by Microorganisms, FEMS Microbiol. Rev., 20, 591-604, 1997. 3. Crolla, K. J. Kennedy, Optimization of citric acid production from Candida lipolytica Y-1095 using n-paraffin, J. Biotechnol., 89, 27-40, 2001. 4. P.P. Bosshard, R. Bachofen and H. Brandl, Metal Leaching of Fly Ash from Municipal Waste Incineration by Aspergillus niger, Environ. Sci. Technol., 30, 3066-3070, 1996.
335
Bioleaching Applications
5. W. Burgstaller, H. Strasser, H. Woebking and F. Schinner, Solubilization of Zinc Oxide from Filter Dust with Penicillium simplicissimum: Bioreactor Leaching and Stoichiometry, Environ. Sci. Technol., 26, 340-346, 1992. 6. L. Shih, Y. T. Van, Y. N. Chang, Application of statistical methods to optimize production of poly (γ-glutamic acid) by Bacillus licheniformis CCRC 12826, Enzyme Microb. Technol., 31, 213-220, 2002. 7. H. Dellweg, (ed.). Biotechnology, 3, 419-465, Weinheim: Verlag Chemie. 1983.
336
15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Production of an Acidithiobacillus ferrooxidans biomass using electrochemical regeneration of energetic substrate C. Morra, N. Gondrexon*, J.-P. Magnin, J. Deseure, P. Ozil Laboratoire d'Electrochimie et de Physico-chimie des Matériaux et des Interfaces, UMR 5631 CNRS-INPG-UJF, ENSEEG BP75, 38402 St Martin d'Hères, France
Abstract The aim of this study was to develop and to optimise a process for the production of an Acidithiobacillus ferrooxidans biomass. This process includes a classical bioreactor associated with two working loops. The first loop involves a gas/liquid contactor, providing the dissolved dioxygen amount required for bacterial growth. A preliminary study of the biomass respiration has allowed to select static mixers because of their efficiency as gas-liquid transfer device [1]. The second loop is devoted to the regeneration of the bacteria substrate (Fe2+) and includes an electrochemical reactor (E3P©: Pulsating Porous Percolated Electrode). Two bacterial culture modes have then been considered: a batch mode i.e. without substrate regeneration and a continuous culture mode involving the electrochemical regeneration of the substrate. With this latter, the bacterial growth is maintained in its exponential phase during more than 100 hours by applying a maximum intensity of 40 A. This operating configuration resulted in a protein concentration up to 110 mg.L-1, i.e. a production yield of Acidithiobacillus ferrooxidans eight times greater than that achieved under the batch mode. This process appears to be among the most efficient existing ones when comparing both its productivity and operating scale. Keywords: Acidithiobacillus ferrooxidans, continuous electrochemical regeneration, static mixers, high-cell density
1.
INTRODUCTION Acidithiobacillus ferrooxidans is a well-known acidophilic, aerobic and chemolithotrophic bacteria which generates its own energy from the oxidation of iron and reduced sulphur compounds. Moreover, this organism is able to grow optimally at acidic pH. Therefore, it is of a great interest for industrial processes such as coal desulfurisation or core bioleaching. On the other hand, this bacteria can be used in wastewater treatment processes and specially for an efficient removal of metallic ions [2,3]. However such an application at a large scale requires important bacteria amounts. This is the reason why a part of our research works is focused on designing and optimising a process for the production of a high-concentrated Acidithiobacillus ferrooxidans biomass. This original process (Figure 1) is composed of a classical biological reactor associated with an electrochemical reactor and an aeration system.
*
[email protected]
337
Bioleaching Applications
Static mixers
- + Power supply Electrochemical reactor
air Biological reactor
Figure 1. Scheme of the high-density biomass production process 2.
MATERIALS AND METHODS
2.1 Strains and culture medium The high-density biomass cultivation is performed with the DSM 583 strain in the socalled 9K Grenoble medium [2] derived from that proposed by Silverman [4]. Ferrous ions (Fe2+) are the major compounds of this culture medium (see Table 1) and act as the single source of mineral energy for the A. ferrooxidans biomass growth. Table 1. 9K Grenoble Culture medium composition Composition FeSO4.7H2O MgSO4.7H2O (NH4)2SO4 K2HPO4 H2SO4
Concentration (g.L-1) 33.3 0.4 0.4 0.4 pH 1.4
2.2 Electrochemical reactor The electrochemical unit is a E3P© reactor (Pulsating Porous Percolated Electrode) [5], classically used to remove heavy metals from dilute aqueous solutions. Its role is to regenerate the substrate (Fe2+) through Fe3+ cathodic reduction, the ferric ions being issued from the Fe2+ oxidation due to bacteria activity. This electrochemical reactor involves a porous granular carbon cathode and a circular titanium anode separated by an anion exchange membrane (IONAC®). The culture medium containing A. ferrooxidans biomass flows throughout the cell under a pulsating mode. This flow mode prevents biomass attachment and biofilm formation onto the granular carbon matrix surface. Moreover, the three-dimension electrode offers a high active surface owing to the high values of both granular specific surface area and mass transfer coefficient [3].
338
Bioleaching Applications
2.3 Biological reactor and gas/liquid contactor The bioreactor is a 50 L spherical pyrex tank. Its temperature is controlled at 30°C that is the optimal bacterial growth temperature. The reactor is aerated and continuously stirred by a closed recirculation loop involving static mixers (SMV Sulzer® DN20). This loop ensures the gas-liquid transfer for a suitable supply of dioxygen and carbon dioxide with respect to the growth and production of a high-density A. ferrooxidans culture. 3.
MATERIALS AND METHODS
3.1 Culture conditions Bacterial cultures in the batch mode were precultured at 30°C with the 9K Grenoble medium in a 4 L bioreactor. This laboratory-scale reactor was well stirred and aerated by a simple air flow. At the end of the preculture, the biomass was harvested to inoculate the 40 L reactor. The initial protein concentration in the culture medium was close to 1 mg.L-1. 3.2 Analytical method for the bacterial growth The cultivated biomass was quantified by determining the protein concentration with the Lowry method [6]. The bacterial growth was estimated by measuring the bacterial metabolism characterised by the ferrous oxidation associated to the Fe2+/Fe3+ ratio. The determination of the substrate oxidation degree depends on the culture mode. For the batch mode, the Fe2+ concentration was determined directly by the phenanthroline method [7]. Total Fe concentration was estimated after prior reduction of Fe3+ ions by hydroxylamin. For biomass cultures in continuous mode, the Fe3+/Fe2+ ratio was deduced, via the Nernst law, from the redox potential given by a redox probe placed directly within the culture medium. 4.
RESULTS AND DISCUSSION
4.1 Experimental configuration for the determination of the growth model parameters A bacterial culture was carried out within a 40 L reactor working in batch conditions. Static mixers provided the dissolved dioxygen necessary for A. ferrooxidans growing. Measuring the remaining substrate and biomass concentrations versus time allowed to determine the growth rate. 4.2 Growth model The postulated model for the A. ferrooxidans growth in the 9K Grenoble medium is characterised by the following ordinary non-linear differential equations: ⎧ dX = µX ⎪ ⎪ dt ⎪ dS −µX = ⎨ YXS ⎪ dt ⎪ dP µX = ⎪ YXP ⎩ dt
(1) (2) (3) 339
Bioleaching Applications
The substrate/product yield can be considered as equal to unity because of the stoechiometric coefficient related to ferrous iron oxidation by A. ferrooxidans according to: 4 Fe2+ + O2 + 4 H+→ 4 Fe3+ + 2H2O (4) and a simple relation between concentrations is thus obtained: YXS=YXP (5) Assuming the bacterial growth kinetics obeys the Contois’ model [8], the growth rate can be expressed as:
µ = µmax
S Ks X + S
(6) where appear the maximum specific growth rate µmax, the substrate concentration S, the bacterial concentration X, and the substrate accessibility constant Ks.
4.3 Implementation The parametric identification for the Contois’ model was performed by fitting the growth experimental data owing to the Fletcher-Reeves optimisation [9] available with the MatlabTM software. Parameters µmax (h-1) and Ks (g.mg-1) have been computed despite the dispersion of the experimental data for biomass concentration. In this way, the yield parameters were fixed by calculating both ferrous iron and protein concentrations between the beginning and the end of bacterial cultures. The state variables Xο, Sο and Pο refer to the beginning of the bacterial culture. Values of both fixed and identified parameters are given in Table 2. Table 2. Biological characteristics of A. ferrooxidans culture (DSM 583) in batch bioreactor (40 L) with the 9K Grenoble medium Fixed values Identified values
Xo (mg.L-1) 1.3 µmax (h-1) 0.131
So (g.L-1) 5.8 KS (g.mg-1) 0.098
Po (g.L-1) 0.8
YXS 2.4
The evolution of protein and Fe2+ concentrations in the culture medium versus time are shown in Figure 2. As it can be seen, the modelled growth curves are in good agreement with experimental results.
340
7
16
6
14 12
5
10 4 8 3 6 2
4
1
2
0
Protein Concentration en (mg L-1)
Fe 2+ Concentration (g L-1)
Bioleaching Applications
0 0
5
10
15 20 Time (h)
25
30
35
Figure 2. Biomass growth curve of A. ferrooxidans (DSM 583), 9K Grenoble medium: ● biomass cell, □ susbtrate, - from the model It is thereby possible to determine the instantaneous biomass production rate (mg protein.L-1.h-1) during the growth phase as well as the bacterial growth cycle during which A. ferrooxidans growth is maximal. Simulated curves resulting from the Contois’ growth model are shown in Figures 3 and 4. The maximal biomass production rate (0.87 mg protein.L-1.h-1), deduced from Fig. 3, is obtained at the end of the exponential growth phase. This optimal value is reached when 65% of the initial substrate has been biologically oxidised (see Fig. 4.).
-1
0,9
-1
Biomass production rate (mg.L .h )
1,0 0,8 0,7 0,6 0,5 0,4 0,3 0,2 0,1 0,0 0
10
time (h)
20
30
Figure 3. Simulated biomass production rate vs. time according to the Contois’ model 341
Bioleaching Applications
-1, -1
Biomass production rate (mg L h )
1 0,9 0,8 0,7 0,6 0,5 0,4 0,3 0,2 0,1 0
0
20
40
3+
60
80
% Fe / Fetotal
100
Figure 4. Simulated biomass production rate vs. Fe3+/ total iron ratio within the culture medium 4.4 Continuous biomass production by electrochemically-assisted cultivation Under these working conditions, the product of bacterial metabolism (Fe3+) was reduced by using an E3P© electrochemical reactor. A schematic reaction diagram for this bio-electrochemical process is proposed in Figure 5.
+
Power Supply
-
O2 + 2 H+
Fe3+
Fe3+
T.f
2 H2O
e-
H2O
e
O2 + 4 H+
-
Electrochemical reactor
Fe2+
Fe2+
Fe2+ Biological reactor
Figure 5. Diagram of the chemistry during electro-cultivation of A. ferrooxidans in the process The substrate concentration (Fe2+) was monitored owing to an oxidoreduction potential PID controller using Labview TM software. The redox potential set point value was selected to obtain the optimal bacterial growth conditions and particularly a 60% 342
Bioleaching Applications
Fe3+/total iron ratio was fixed. Moreover, the pH was kept constant to 1.4 by addition of concentrated H2SO4 solution preventing precipitation of jarosite. Experimental results obtained in such conditions are illustrated in Figure 6. For 17 hours, the biomass oxidises the substrate (Fe2+) and tends to develop in an exponential phase. Thereby, the Fe2+ concentration decreases down to approximately 40% of the total iron concentration present in the cell culture (2.64 g.L-1). At the 18th hours, the electrochemical reactor starts on to continuously regenerate the substrate in order to maintain such a physiological bacteria state for more than 100 hours. Fe2+ concentration is thus kept constant. The cell density increases from 1.3 mg protein.L-1 to more than 110 mg protein.L-1 at the end of the culture.
E3P© electrochemical reactor
7 Off
120
On
100 -1
Protéin concentration (mg L )
2+
-1
Fe Concentration (g L )
6 5
80 4 60 3 40 2 20
1 0 0
20
40
60 Time (h)
80
100
0 120
Figure 6. Growth curves for A. ferrooxidans using electrochemically-assisted cultivation: ● biomass concentration, ⎯ Fe2+ concentration 5.
CONCLUSION A biomass production process has been designed coupling a classical biological rector, a E3P© electrochemical reactor, and static mixers. After a 115-hour cultivation period, a protein concentration up to 110 mg.L-1 was reached. Owing to its experimental performances and operating scale, the process proposed here may be regarded as one of the most efficient by comparison with previous results reported in literature [10-11]. Further experiments are now required to investigate the continuous production of A. ferrooxidans biomass for longer operating duration and to point out the possible limiting factors of the process. NOMENCLATURE Xο = Initial biomass concentration (protein) X = Biomass concentration (protein)
g.L-1 g.L-1 343
Bioleaching Applications
Sο S Pο P YXS YXP µ µmax
= = = = = = = =
Initial substrate concentration (Fe2+) Substrate concentration (Fe2+) Product concentration (Fe3+) Initial product concentration (Fe3+) Biomass/substrate yield Biomass/product yield Biomass specific growth rate Maximum biomass specific growth rate
g.L-1 g.L-1 g.L-1 g.L-1 mg.g-1 mg.g-1 h-1 h-1
REFERENCES 1. C. Morra, N. Gondrexon, J.P. Magnin, P. Ozil, Récents Progrès en Génie des Procédés, (86) 15 (1998) 11. 2. J.P. Magnin, F. Baillet, A. Boyer, R. Zlatev, M. Luca, A. Cheruy, P. Ozil, Canadian. J. Chem. Eng., 76 (1998) 978. 3. J.P. Magnin, P. Ozil, Génie des procédés biotechnologiques, Ed CNRS, 1 (1998), 125. 4. M.P. Silverman and D.G. Lundgren, J. Bacteriol., 77 (1959) 642. 5. P. Roquero, G. Lacoste, P.L. Fabre, P. Duverneuil, A. Ghanem-Lakhal, P. Cognet and J. Berlan, Chem. Eng. Sci., 51, (1996) 1847. 6. O.H. Lowry, N.J. Rosebrough, A. L. Farr et R. J. Randall, J. Biol. Chem., 193 (1951), 267. 7. M.K. Muir and T. N. Anderen, Metal. Trans., B-8b (1977) 517. 8. D.E. Contois, J. Gen. Microbiol., 21 (1959) 40. 9. R. Fletcher, M. Reeves. Computer J., 7 (1964) 149. 10. R.C. Blake II, G.T. Howard, S. McGiness, Appl. Environ. Microbiol., 60 (8) (1994) 2704. 11. N. Matsumoto, H. Yoshinaga, N. Ohmura, A. Ando, H. Saiki, Biohydromal., 9 (1999) 757.
344
15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Removal of dibenzothiophene from fossil fuels with the action of iron(III)-ion generated by Thiobacillus ferrooxidans: Analytical aspects V.P. Beškoskia, V. Matića, S. Spasića and M.M. Vrvića,b∗ a
Department of Chemistry, Institute of Chemistry, Technology and Metallurgy, 11001 Belgrade, Njegoševa 12, P.O.Box 473, Serbia and Montenegro b Faculty of Chemistry, University of Belgrade, 11001 Belgrade, Studentski trg 16, P.O.Box 158, Serbia and Montenegro
Abstract Among various classes and numerous kinds of organic compounds with sulphur, identified in fossil fuels, the most represented is dibenzothiophene (DBT) and its derivatives. Therefore, this compound can be considered to be the model substrate of organically bonded sulphur in fossil fuels. In focus of our interest is bacterial removal of sulphur (desulphurization) from oil (bituminous) shales as fossil fuel and potential/alternative source of “synthetic” hydrocarbons liquid fuels. The application possibility of Thiobacillus ferrooxidans for removal of organic sulphur out of the fossil fuels is in the following idea: ”To convert DBT into the (water soluble) sulphur-free form with the oxidation by bacterially generated-regenerated iron(III) sulphate from pyrite”! The results presented in this paper are just an initial step in realization-checking of the mentioned idea and relate to the development and adaptation of well-known analytical procedures for the quantitative analysis and monitoring of changes on DBT molecule as the model substrate. The analytical methods are also checked in the interaction of DBT with potential-model solutions: sulphuric acid solution (pH 2.5), iron-free medium 9K, medium 9K and medium 9K in which iron(II) is chemically oxidated beforehand. The shake flasks testing technique has been applied in these experiments. For the quantitative analysis, gas chromatography is the choice method, while the UV spectrophotometry is the most convenient for fast detection of changes in the DBT structure. 1H and 13C-NMR and mass spectrometry are the instrumental structural methods by means of which it is possible to monitor DBT transformation pathway. Keywords: analytical aspects, dibenzothiophene, fossil fuels, removal, Thiobacillus ferrooxidans
∗ Author for correspondence. Phone: +381-11-637-237. Fax: +381-11-636-061. E-mail:
[email protected]
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1.
INTRODUCTION Reduction of emission of sulphur dioxide as the combustion product of various sulphur forms from fossil fuels is the imperative in protection of environment in forthcoming years which directly depends on the content of sulphur in them. Thus, solving the problem is brought down to reduction of sulphur quantity in fuels resp. their desulphurization in order to avoid post-combustion refining of gases from sulphur. This global ecological problem is followed also by the recommended an/or adopted strict legislature (in relation to the permitted emission of sulphur dioxide in the air and the total sulphur content in fuels), which varies in some countries and the regions of the world [1,2]. Sulphur in solid fossil fuels (coals and oil shales) is found in three basic forms: sulphate, pyrite and organic sulphur [3-5]. Elementary sulphur is also present in fossil fuels [6]. In crude oil are identified the most diversified classes of organosulphuric compounds: from simple alkylthiol-mercaptan to polycondensed benzophenanthrotiophene, their derivatives and polyaromatic compounds with more sulphur atoms and other heteroatoms [7, 8]. Over 200 organic compounds containing sulphur is identified in petroleum [9]. DBT and its derivatives are found in all fossil fuels and they are the main carriers of organic sulphur [10, 11].
7
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8 7
4b
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S
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Structural formula of DBT and atoms numbering The amount of overall sulphur and individual forms, classes and compounds of sulphur in fossil fuels considerably varies subject to origin-source-deposit. Its distribution is not constant, not even within the same deposit [4-5]. Generally, desulphurization of fossil fuels, herein only bioprocessing is studied, understands separation of the basic sulphur forms. Sulphate sulphur is not a problem since sulphates are well soluble in water. Bacterial depyritization and segregation of elementary sulphur with high efficiency (over 90%) is possible by action of Thiobacillus ferrooxidans (Th. ferrooxidans) and mixed cultures Th. ferrooxidans i Th. thiooxidans [12-14] and by thermophilic Sulfolobus acidocaldarius [15,16]. The biggest problem is separation of organic sulphur. A great number of microorganisms in heterotrophic conditions of cultivation-growth, among which also Sulfolobus acidocaldarius [17] is tested as the “biological agent” for removal of organic sulphur, both from coals and from crude oil and hydrocarbons fuels at which different efficiencies are obtained but considerably less than depyritization [18-20]. DBT is a typical representative of organosulphuric compounds in fossil fuels and therefore it can be considered a model substance resp. a model substrate of organic sulphur. The pathways of microbiological separation of sulphur from DBT are different, but essentially reduced to: (1) removal of sulphur with no cleavage of carboncarbon (C-C) bonds (“4S” pathway) with occurrence of diphenyl and its ortho-hydroxy derivatives [21], and (2) with cleavage of C-C bonds at which DBT is the sole source of carbon, sulphur and energy, while as a result are obtained in the final products of mineralization but also aromatic carbonylic compounds, subject to the applied bacterial species [22,23]. Sulphate is the product of microbiological desulphurization of DBT. 346
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Oil (bituminous) shales (compact sedimentary rocks of homogeneous fine-grained composition) are potentially and alternatively an important source of "synthetic" hydrocarbons liquid fuels [24], due to which they are the subject of geochemical investigations and economic interests. Therefore, this fossil fuel is the subject of our many-year researches. Majority of oil shales (about 80%) are inorganic components viz. carbonates, alumino-silicates and pyrite. Kerogen, as insoluble and of heterogeneous macromolecular crosslinking structure, is a dominant organic substance (approx. 95% out of the total organic matter) while, in the organic solvents, soluble bitumen is present in the quantities of several percentages [25-26]. Fundamental organic-geochemical studies of kerogen require preparation of its concentrates with relatively pure and unaltered kerogen. Removal of carbonates and alumino-silicates is realized by the action of mineral acids (diluted hydrochloric and concentrated hydrofluoric acids). Partial removal of aluminosilicate is also possible by application of Bacillus circulans strains (bacterial desilicification) [27-29]. Bacterial removal of pyrite which is closely associated with kerogen from crude oil shale and its concentrates by Th. ferrooxidans is exceptionally efficient (approx. 97%), at which rich concentrates of kerogen with unchanged organic substance are obtained [30-34]. The structure of bituminous coal with included pyrite crystals [35], shown in Fig. 1 is the most alike the characteristical crosslinked structure of the organic part of bituminous shales with pyrite. H
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H OH HH
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O H
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H FeS2 |
HH
H
Figure 1. Representative structure of bituminous coal with included pyrite crystals. Arrows indicate the atoms of organically bonded sulphur
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1.1 Scientific hypothesis of DBT removal from fossil fuels by the action of iron(III)ions generated by Th. ferrooxidans Microbial desulphurization of DBT is a complex enzyme process. Our scientific hypothesis of DBT removal from fossil fuels by the action of iron(III) ion generated by Th. ferrooxidans is based on the following idea: "To convert DBT into the (water soluble) sulphur-free form with the oxidation by bacterially generated-regenerated iron(III) sulphate from pyrite"! The hypothesis can be schematically shown as in Fig. 2. FeS2 (s) Th. ferrooxidans + H2O (l)
Fe2(SO4)3 (aq) Th. ferrooxidans
S DBT
(s)
FeSO4 (aq)
+ H2SO4 (aq). R'R DS-DBT
(s)
Figure 2. Schematic presentation of hypothesis. DS-DBT-Desulphurized DBT. R and R' = -H and/or -OH Hypothetical molecular chemical equation resulting from the schematic hypothesis diagram would have the following form: R=S + 2 Fe2(SO4)3 + 4 H2O → RH2 + 4 FeSO4 + 3 H2SO4
(1)
Thermodynamic computations [36-39] of DBT desulphurization (R=S) to biphenyl (RH2) by the action of iron(III) sulphate according to the equation (1) indicate that free energy of this process under standard conditions (unit activity and 298 K) and at pH 2.5 has the value ∆Go2.5 = –91 kJ. This means that the process can be spontanous developed, being one of the proofs that theoretically the hypothesis is correctly set. The presence of pyrite as the source of iron(III)-ion (oxydans) together with DBT should favor oxidation of thiophenic sulphur. In case the hypothesis is proved, this form of desulphurization would be bioprocessing by an indirect mechanism.
1.2 Aim and scope of the study This paper represents an initial step in realization of the stated idea-checking of its correctness. Therefore, its essential aim is characterization of DBT as a model substrate and checking and adaptation of analytical procedures for its determination. For that purpose it was necessary to solve the following issues: a) to characterize the commercial DBT by the instrumental analytical methods; b) to adapt the methods of UV spectrophotometry and gas chromatography for fast proving the changes of DBT structure and its determination; and c) to check applicability of analytical procedures and DBT stability in the interaction with model solutions. The literature data [40] indicate that the acceptable value for an average DBT concentration in fossil fuels is approx. 25 mg/kg(L). Therefore, the concentrations of 348
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solutions for analytical purposes as well as in model solutions would be close to this value.
2.
MATERIAL AND METHODS
2.1 Model substrate DBT purity 98% (Aldrich, Catalog No. D3,220-2). Molecular formula C12H8S, relative molecular mass 184.26, CAS No. [132-65-0]. 2.2 Model solutions and model systems Solution of sulphuric acid (0.27 mL concentrated acid in 1 L solution) with pH about 2.5 (designation: pH 2.5), medium 9K free from iron(II) sulphate heptahydrate (designation: 0 K), medium 9 K [41] (designation: 9 K) and medium 9 K in which iron(II) sulphate is chemically oxidated into iron(III) sulphate (designation: 9 K 3). pH of all solutions is adjusted to about 2.5. Model solution 9 K 3 is obtained by oxidation of medium 9 K with 30% (m/m) hydrogen peroxide with heating to complete decomposition of peroxide. Peroxide consumption is about 10 mL/L. The concentrated solution DBT in ethyl acetate (EtOAc) is added in solutions so that the final concentration of DBT should be 25 mg/L. Then, the solvent is separated on the rotary vacuum evaporator. Thus obtained model systems are slightly opalescent and DBT is homogeneously suspended. They are further "cold sterilized" and poured into sterile Erlenmeyer flasks equipped with sterile microbilogical stoppers. 2.3 Chemicals All used chemicals are of pro analysi purity, resp. of the appropriate purity for application in analytical purposes. EtOAc was additionally refined by destilation. Demineralized water was used. 2.4 Shake flasks testing Experiments were conducted in the room termostated at (28±1°C) on the reciprocating shaker (New Brunswick Scientific, model R-82) with 200 strokes/min. Erlenmeyer flasks of the same geometry and the total volume of 5 L with 1 L solution (volume ratio 1:5) were used. They were plugged with identical microbiological stoppers made of cotton and gauze to ensure constant and reproducible oxygen transfer in the same conditions. The tests lasted 15 days.
2.5 Analytical methods The following methods and instruments were used. DBT extraction from model system. pH solution was adjusted prior to DBT extraction by EtOAc. Then the solutions were saturated by the solvent (about 50 mL/L) and multipleextracted in the ultrasonic bath for 30 min. Every time the organic phase was separated in the separatory funnel. The collected fractions were dried overnight by anhydrated sodium sulphate and after filtration the total volume was filled up to 100 mL. pH. pH-meter (Radiometer, type PHM 26) with a combined electrode PHC 2401 of the same manufacturer. UV spectrophotometry. UV-Vis spectrophotometer: Beckman, model DU-50. Quartz cuvettes 1 cm. 349
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FT-IR spectra. KBr pellet (approx. 1:100). Instrument: Perkin Elmer, model PE 1725 X.
NMR spectra. 1H-NMR in acetone-d6 (at 200 MHz). 13C-NMR in DMSO-d6 (at 50 MHz), ambiental temperature. Internal standard TMS. Instrument: Varian, model Gemini 2000. Mass spectrometry: Chemical ionization (CI) with iso-butane. Mass spectrometer: Finnigan-Mat, model 8239. Gas Chromatography (GC). GC analysis was conducted on the gas chromatograph (Varian, model 3400) with flame-ionization detector (FID). Temperature program: 50285°C, 15°C/min. Hydrogen flow rate 1 mL/min. Column (J&W Scientific): length 30 m, internal diameter 0.25 mm, film thickness 0.25 µm and filling DB-5. DBT standards were prepared by weighing the required substance amount and by dissolving it in the appropriate EtOAc volume.
3.
RESULTS AND DISCUSSION
3.1 Spectral DBT characterization UV spectrum of solution of the model substrate, concentration 25.1 mg/L in EtOAc in relation to solvent, is shown in Fig. 3. Characteristics of the spectrum in ethanol on the basis of literature data [42] is shown in table on the figure. Compared with literature data it can be seen that the peak at 237 nm is missing which is the result of application of EtOAc as solvent and the instrument is double beam. Sharp signals in the spectrum correspond to the positions stated in literature but are shifted by 3-6 nm toward the ultraviolet region, which is not unusual, the more so because different solvents are in question. Since the characterization on the basis of UV spectrum is fast and simple, it stands as the optional method for quality analysis of changes in DBT structure in the future work.
Figure 3. UV spectrum of DBT 350
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FT-IR spectrum of the used DBT is shown in Fig. 4. Signals in the range of 1600-400 cm can be considered characteristical for DBT [43] which means that the changes in this spectrum range and appearance of other signals indicate the changes in DBT molecule structure. -1
Figure 4. FT-IR spectrum of DBT The protonic NMR spectrum of the commercial DBT (Fig. 5) compared with the literature one (inserted table) [44] does not show any differences in spite of 98% purity. Signals at about 2.1 and 2.9 ppm are the solvent impurities.
Figure 5. 1H-NMR spectrum of DBT 13
C-NMR spectrum model substrate shown in Fig. 6. Compared with the literature data (inserted table) [42], as also in the case of 1H spectrum, conincides well. Signals at about 40 ppm are the solvent impurities. 351
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Figure 6. 13C-NMR spectrum of DBT On the basis of NMR characteristics, it can be said that in the future work this method could be useful for determination of the structural changes of DBT molecules. DBT mass spectrum is shown in Fig. 7. The signal at m/e 185, which is also the base pick (intensity 100%) corresponds to the molecular ion. All signals on larger masses than this one, and of small intensity, originate from impurities.
Figure 7. Mass spectrum of DBT Peak at m/e 240, related to intensity, point to fragmentation and rearrangement of the basic molecule. These results clearly show that the mass spectrometry in further researches can, together with NMR, be the key structural instrumental method for explanation of DBT transformation – desulphurization mechanisms.
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3.2 Checking and adaptation of the gas chromatographic methods for determination of DBT Gas chromatography is the analytical method used for determination of DBT. Characteristical gas chromatogram for the standard solution model substrate (25 mg/L) is shown in Fig. 7.
Figure 7. Gas chromatogram of standard DBT solution The signal with retention time (RT) at somewhat more than 10 min originates from DBT. Another two outstanding peaks are impurities from the solvent which was confirmed by recording the chromatogram only for the solvent and because of which we used to check GC purity EtOAc prior to each analysis. Linearity of the calibration diagram and detection limit were defined within the concentration range (c) 5 to 50 mg DBT/100 mL (attenuation 32, injected volume 1 µL), which in the described experimental conditions for testing of interaction with model solutions correspond to DBT concentrations in 1 L. By processing the data with linear regression for area counts [AC = f(c)] by Microcal Origin 5.0 program, we obtained the following equation of the standard line in the explicit form: AC = 2726±124 c - 3725±3488
(2)
Statistical criteria (r = 0.9939, SD = ±4976 and p < 0,0001 for n = 8) indicate that the method is accurate, precise and reliable. The method is also sensitive (α ≈ 90°), and the detection limit is 1.4 mg DBT/100 mL(L) for AC = 100. All GC results indicate that DBT can be directly determined (without the internal standards) and that it is applicable for the quantitative DBT analysis in desulphurization experiments.
3.3 Conditions for DBT extraction from the model system Since the model substrate resp. DBT in fossil fuels is in water suspension, it is indispensable to extract it beforehand in order to determine its concentration. Therefore, we have checked for the adopted DBT concentration in fossil fuels and at the initial pH of above 2.5 (model solution pH 2.5), the influence of the number of extractions and pH solution on DBT recovery, for which UV (λmax = 255 nm) are applied as faster and more simplified, resp. GC method.
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For the triple extraction in the range of pH 1-14, the results shown in Fig. 8 are obtained. The highest recovery (91.7%) is obtained at pH 7, which means that it is necessary to adjust pH model system to this value prior to extraction. With the triple extraction at the stated pH, the extraction degree of 91.6 ± 2.8% (n=5) is obtained, and the double extraction gave the recovery of 87.2 ± 4.9% (n=5).
Figure 8. Influence of pH on DBT recovery from the model system with solution pH 2.5 3.4 DBT interaction with model solutions For the purpose of checking the applicability of analytical and instrumental structural methods, as well as DBT stability/changes as the model substrate of the organic sulphur in fossil fuels in the interaction with model solutions which are by their composition very close to the expected ones in the real experiments for separation of the organically-bonded sulphur, with participation of Th. ferrooxidans, and generation-regeneration of iron(III)ions from pyrite, we have conducted four series of testings with various, the aforegoing model solutions. The qualitative changes were identified by recording UV spectra of the extracted DBT after the finished tests. Shifting of the maximum absorption or phenomena of new peaks is not noticeable in any model system. However, for model solutions 9 K and 9 K 3, the peak shape at 255 and 263 nm is somewhat changed, so that inflections are noticeable on the portion toward the lower of the first signal and the portion toward the higher wave lengths of the second peak. The changes are more expressive on the signals for the extracted DBT from the solution 9 K 3. This could indicate certain changes in DBT molecule by the action of iron(III) ion, which occurs by oxidation of oxygen from the air in the solution 9 K, resp. iron(III) found in the solution 9 K 3. The results of the quantitative analysis are shown in Table 1.
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Table 1. Results of the quantitative analysis of DBT in model systems Model solution pH 2.5 0K 9K 9K3
Recovery of DBT, [%] 93.1 88 33.1 4.9
Extraction degree for the first two solutions is expected on the basis of the results with standards, though there is certain possibility that the salts from the solution 0 K affect lower extraction efficiency. Small and extremely low recovery in case of other two solutions does not mean that significant transformations of DBT occurred (neither shown UV spectra), but that the extraction procedure is not efficient enough. Namely, during adjustment of the optimal pH for extraction, the voluminous precipitates of iron(II) and iron(III) hydroxide drop which, due to their sorptive, occlusive and inclusive properties (physical bonding) make aggregations with DBT which is not found in the real solution but only is well suspended. This unables an efficient contact of DBT particles with the solvent resp. reduces the extraction efficiency which is more expressive with iron(III) hydroxide. This is confirmed by our preliminary results [45], after which the usual extraction is carried out in the extraction funnel, with which the quantity of the extracted DBT was below the detection limit improved by the procedure with ultrasound. On the gas chromatogram for DBT from 9 K 3, besides the signal for the solvent contamination there are also two peaks with RT close to DBT position. One is at the place which corresponds to a lower and the second to a higher polarity of DBT. Their surfaces are about 3, resp. about 5% in relation to DBT signal surface. These results too point at possible oxidation-desulphiruzation changes of DBT structure by the action of iron(III) sulphate. In the outstanding chromatograms, neither these nor other signals were identified except the expected ones.
4.
CONCLUSIONS All the results obtained and discussed point at the following conclusions: 1. Thermodynamic computations for hypothetic chemical equation of DBT desulphurization by the action of iron(III)-ion speak in favor of foundation of the idea related to the possibility of removal of the organic sulphur from fossil fuels in the described way; 2. At the same time the presence of pyrite and organically-bonded sulphur in fossil fuels, and in oil shales too, which are the subject of our special interests, may favor removal of DBT by oxidation with iron(III) sulphate which would be generated by Th. ferrooxidans from that pyrite, whereby the overall desulphurization would be realized in the same process; 3. The commercial DBT (purity 98%), as a model substrate, is characterized with the structural instrumental methods, viz. by: UV, FT-IR, 1H and 13C-NMR and mass spectra; 4. UV and FT-IR spectra are suitable for the qualitative analysis of changes in DBT structure at which, due to simplicity and the speed of conducting the advantage is given to UV spectrophotometry;
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5. Other spectral methods are applicable for determination of the product structure and mechanisms of DBT transformations; 6. Gas chromatography is the appropriate quantitative method for the analysis of DBT content; 7. The conditions for DBT extraction from reactive mixture are: pH 7 and minimum three times repeated procedure (recovery about 90 and more percentages); 8. Checking of applicability of analytical procedures and stability/changes of DBT at interaction with the model solutions by shake flasks test technique indicates that there are no changes in the structure of the model substrate when the solution of sulphuric acid with pH 2.5 and medium 9 K without iron(II) sulphate are applied, while in the case of the medium 9 K and the same medium with iron(III)-ion possible desulphurization is noticed, to which the changes in the signal structure in UV spectrum point out, resp. in UV spectrum and gas chromatogram; and 9. Aggregates of the iron hydroxides precipitates in model solutions with iron(II) and iron(III)-ion occurred at pH adjustment prior to extraction are physically bonding DBT, so that the recovery is not acceptable (this does not occur with iron-free solutions), exspecially in the case of iron(III) hydroxide (only about 5%), which means that DBT extraction method from the model system resp. under real experimental conditions, represent a series analytical problem in preparation of DBT specimen for further testings, to which attention should be paid.
ACKNOWLEDGEMENTS The authors wish to express their sincerely gratitude to the Center for Instrumental Analysis-the Department of Chemistry, Institute of Chemistry, Technology and Metallurgy in Belgrade, for recorded spectra and gaschromatographic analysis. Much merit goes to the director of the Department professor Vlatka Vajs who’se interest, skill and experience added to the quality of the work. The financial support of the Ministry for Science, Technologies and Development of Serbia (Grants No. 1740 and 0295) and the Company NRK Engineering Ltd. from Belgrade, has been important to the study. REFERENCES 1. D.J. Monticello, CHEMTECH, July, (1998) 38. 2. B.L. McFarland, D.J. Boron, W. Deever, J.A. Meyer, A.R. Johnson and R.M. Atlas, Critical Rev. Microbiol. 24 (1998) 99. 3. L.-K. Ju, Fuel Sci. Technol. Int. 10 (1992) 1251. 4. A. Attar, Hydrocarb. Proces. 58 (1979) 175. 5. A.M. Khalid, D. Bhattacharyya and M.I.H. Aleem, J. Ind. Microbiol. Suppl. 5 (1990) 115. 6. R.T. Greer, Scan. Elec. Micros. 1 (1979) 477. 7. J.P. Boudou, J. Boulègue, L. Maléchaux, M. Nip, J.W. de Leeuw and J.J. Boon, Fuel, 66 (1987) 1558. 8. K.G. Kropp and P.M. Fedorok, Can. J. Microbiol. 44 (1998) 605. 9. C.J. Thompson, in Organic Sulfur Chemistry, R.K. Freidlina and A.E. Skorova (eds.), Pergamon Press, Oxford 1981, p. 181. 10. C.L. Spiro, J. Wong, F.W. Lytle, R.B.Greegor, D.H. Maylotte and S.H. Lanson, Science, 226 (1984) 48. 356
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42. C.W. Bird and G.W. Cheeseman, in Comprehensive Heterocyclic Chemistry, A.R. Katritzky and C.W. Rees (eds.), Vol. 4, Part 3, Pergamon Press, Oxford, 1984, p. 7. 43. J. Pouchert (ed.), The Aldrich Library of Infrared Spectra, Aldrich Chemical Company, Milwaukee, 1981, p. 1266A. 44. J. Pouchert (ed.), The Aldrich Library of NMR Spectra, Aldrich Chemical Company, Milwaukee, 1983, p. 555C. 45. M.M.Vrvić, V. Matić, S. Spasić, V. Beškoski, Report of Researches on Grants No. 1740 and 0295 with the Ministry for Science, Technology and Development of Serbia in 2002, Not public published, Department of Chemistry, Institute of Chemistry, Technology and Metallurgy, Belgrade, 2003.
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Solids loading in the bioleach slurry reactor: mechanisms through which particulate parameters influence slurry bioreactor performance S.T.L. Harrison, A. Sissing, S. Raja, S.J.A. Pearce, V. Lamaignere and M. Nemati* Bioprocess Engineering Research Unit, Department of Chemical Engineering, University of Cape Town, Rondebosch 7701
Abstract Operation of the slurry bioreactor at maximal solids loading is a key factor in determining the economic performance of processes such as mineral bioleaching in which the solid phase represents the key nutrient or reactant in the system. Through studies of mesophilic and thermophilic minerals bioleaching systems, limits to the solids loading have been proposed. Similarly, limits on solids loading have been illustrated in the model yeast-quartzite slurry reactor system. In the paper presented, the combined influence of solids loading and particulate parameters such as particulate size and mineral quality on process performance are presented across three microbial systems: the mesophilic bacterial leaching of mineral sulfides dominated by Acidithiobacillus ferrooxidans and Leptospirillum ferrooxidans; the thermophilic microbial leaching of pyrite and chalcopyrite concentrates as well as concentrate – quartzite mixtures dominated by Sulfolobus species; and the Saccharomyces cerevisiae quartzite model system used for the study of slurry bioreactors. The influence is discussed in terms of process performance, specific activity of the microbial phase as well as structural variation in the biophase. Mechanisms explaining the altered performance under specified culture conditions are sought through an analysis of microbial cell damage resulting from the hydrodynamic environment. Keywords: bioleaching, microbial cell damage, metabolic activity, solids loading
1.
INTRODUCTION It is well recognised that microorganisms express biological responses to stress incurred in their culture environment. These stresses include osmotic, oxidative, thermal, chemical and hydrodynamic stresses. While less well studied than other stress systems, responses to adverse hydrodynamic conditions are reported to include changes in specific growth rate, nutrient uptake rate, product formation rate and morphology of the micro-
* Current address: Department of Chemical Engineering, University of Saskatchewan, Saskatoon SK, Canada, S7N 5C5.
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organisms (Logan and Dettmer 1990, Toma et al. 1991). The presence of particulates in bioprocesses may further aggravate the hydrodynamic response. Agitation of a microbial phase in the presence of a particulate phase is a well-recognised means of cell disruption for product liberation through the bead mill (Currie et al. 1972, Schutte et al. 1986). Yet it is frequently necessary to grow a microbial phase in the presence of a particulate phase, particularly where this particulate phase forms the source of nutrients or is critical to the microbial energy provision such as occurs in mineral bioleaching. The particulate phase may form through the precipitation of products formed such as the precipitation of metal sulphides formed during biological sulphate reduction or provide the support in immobilised cell systems. In mesophilic mineral bio-oxidation systems, a critical solids loading has been reported above which the process is detrimentally affected (Gormley and Brannion 1989, Torma et al. 1992). Studies of solids loading in mesophilic processes have indicated that loadings of 18-20% (m/v) may be used routinely (Oguz et al 1987). Beyond this, performance is impaired. While this may result from gas liquid mass transfer limitation restricting the oxygen or carbon dioxide available and is affected by the grade of concentrate or ore used (Torma et al. 1992, Bailey and Hansford 1993), the influence of mechanical stress has not been rigorously discounted. Limited study on the influence of the solids phase on thermophilic leaching performance is reported. Initial reports suggest that the thermophiles used for bio-oxidation appear to be sensitive to hydrodynamic conditions (Jordan et al 1993, Clark and Norris, 1996) and the presence of solids (Norris and Bar 1988, Le Roux and Wakerley, 1988, Escobar et al. 1993, Torres et al. 1995, Nemati and Harrison, 2000). These disadvantages may be due to the fact that the Sulfolobus lack a rigid peptidoglycan cell wall (König and Stetter 1986, König 1988, Michel et al. 1980). In addition, an increase in temperature causes the fluidity of tetraether-based cellular membranes to increase (Kelly and Deming, 1988). The effect of agitation intensity on cell damage has been investigated in the mineral bio-oxidation and animal cell-microcarrier systems. In the mineral bio-oxidation system, Hackl et al. (1989) found that an impeller tip speed of 5.3 m.s-1 was detrimental to the acidophilic iron and sulphur oxidising mesophiles. However normal leach rates were observed when reducing the tip speed of the Rushton turbine to 3.3 m.s-1. Investigating the growth of the acidophilic iron and sulphur oxidising mesophiles in the presence of 2% (v/v) pyrite, Pearce (1993) determined that the cell growth was inhibited at a tip speed of 2.6 m.s-1 for a 6-bladed Rushton turbine, while an impeller tip speed of 1.4 m.s-1 was not detrimental to the cells. In this paper, the influence of the presence of the solids phase on microbial cell damage in the slurry bioreactor is presented across the following microbial phases: the model system Saccharomyces cerevisiae in both the exponential and stationary growth phases, the acidophilic iron and sulphur oxidising mesophiles used in mesophilic bioleaching and Sulfolobus metallicus used in thermophilic bioleaching. Parameters of the solids phase investigated include solids loading, particle size, nature of the solids phase used, agitation intensity and thereby energy dissipation rate as well as impeller tip speed. The studies consider both the performance of the bio-phase as well as the biological responses observed at a cellular level, By comparison across a range of microbial systems, generic findings are sought.
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2.
MATERIALS AND METHODS
2.1 Microorganisms and Reactor Systems Used The studies were carried out over the following microbial systems: a mesophilic mineral bioleaching system using a mixed culture of acidophilic iron and sulphur oxidising mesophiles, two thermophilic mineral bioleaching systems using Sulfolobus metallicus BC, a model Saccharomyces system under growth conditions and a similar system under stationary phase conditions. Details of the microbial system, growth conditions, particulate phase and reactor configuration for each system is given in Table 1. 2.2 Analytical Procedures The concentration of cells free in suspension was measured by direct counting using a Petroff-Hauser-type cell counter (haemocytometer) of 0.02 mm depth and 1/400 mm2 area under the light microscope at magnification. Cell counting and cell size determination was also conducted by volume displacement using the Cell Facts cell sizer (Microbial Systems, U.K.). For the yeast cultures, cell viability was determined by methylene blue staining using a modified Ringer salt solution. Disruption of the yeast cells was determined by soluble protein release, using the method of Lowry et al. (1951). Maximum protein release was determined by a dual pass through the French Press. The transmission electron microscopy methodology is detailed in Lamaignere (2002). The pH and redox potential were measured at room temperature. The redox electrode was a combined platinum/reference redox cell. The concentration of ferrous iron in solution was determined by titration against 0.017 M potassium dichromate in the presence of N-phenyl anthranilic acid as an indicator (Vogel, 1989). To determine the concentration of total iron in solution ferric iron was reduced to ferrous iron using stannous chloride as reducing agent, followed by titration against potassium dichromate. The ferric iron concentration was estimated by difference between the ferrous iron and the total iron concentration. Since part of the iron was precipitated during leaching, the iron concentration was determined in the supernatant of both pre- and post acid washing of the suspension. Atomic absorption spectroscopy was used to determine copper ion concentrations. A Varian Spectra AA-200 Atomic absorption spectrophotometer incorporating Spectra AA 100/200 (version 1.1) software was used. The operational parameters were as follows: slit width of 0.2 nm; air/acetylene flame; lamp current of 4 mA; and wavelength of 217.9 nm. The concentration of iron and, where appropriate, copper in solution was determined using the supernatant of a centrifuged sample. Table 1. Microbial systems and reactor conditions used across the five studies conducted System Microorganism
System 1 At. Ferrooxidans At. Thiooxidans L.ferrooxidans
Growth phase
Growing system
Growth medium
9K medium
System 2 System 3 Sulfolobus Sulfolobus metallicus metallicus BC BC Growing Growing system system . -3 0.4 kg m (NH4)2SO4 0.5 kg.m-3 MgSO4.7H2O 0.2 kg.m-3 KH2PO4
System 4 S.cerevisiae (Baker’s yeast) Growing system Minimal glucose medium
System 5 S.cerevisiae (Baker’s yeast) Stationary phase Phosphate buffered saline
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System 1
Inoculum Particulate phase Std particle size
Quartzite Mean: 53 µm
System 2 System 3 0.1 kg.m-3 KCl 20% by volume, resulting in a cell concentration of 2-4 x 108 cells.ml-1 Pyrite & Chalcopyrite quartzite & quartzite
37-75 µm
Reactor configuration
2 hr stress period, 1 dm3 STR, followed by growth in a shakeflask in the absence of solid phase
Operating temperature
30°C
Agitation rate (rpm)
Std: 772 rpm R: 400-090 rpm
Standard: 560 rpm Range: 560-760 rpm
Impeller tip speed
Std: 2.2 m.s-1 R: 1.1-3.1 m.s-1
Reference
Pearce 1993
Standard: 1.7 m.s-1 Range: 1.7-2.3 m.s-1 Sissing & Harrison et al. Harrison 2003 2003
3.
1 dm3 baffled, aerated STR, with working volume of 0.7 dm3. Pitched blade impeller.
68 – 70°C
68 – 70°C
System 4
System 5
10% by volume
Dry biomass conc. of 53 kg.m-3
Quartzite 600-850 µm Baffled, flat bottom, aerated 2 dm3 STR. Working volume of 1.5 dm3
Baffled, flat bottom, 3 dm3 STR. Working volume of 2.45 dm3
30°C
<20°C
Std: 565 rpm R: 460-850 rpm Standard of 2.3 m.s-1 Lamaignere 2002
200-1000 rpm 0.77-3.87 m.s-1 ScholtzBrown 1997
RESULTS
3.1 The influence of solids loading in the slurry reactor on performance The authors have studied the influence of solids loading on microbial cell damage in slurry bioreactors across the following microbial phases: the model system Saccharomyces cerevisiae in both the exponential and stationary growth phases, the thiobacilli used in mesophilic bioleaching and Sulfolobus metallicus used in thermophilic bioleaching. Scholtz et al. (1997) have shown the increase in cell disruption of stationary phase Saccharomyces cerevisiae with increasing solids loading across the range of quartzite loading of 5 to 40% (v/v). Further they have illustrated that disruption in excess of 90% is found independent of solids loading across the range studied. The disruption is first order with respect to the concentration of intact micro-organisms present. The first order disruption rate constant, k, was a function of both the power input per unit volume (P/V) and the volume fraction of solids present (Φ): k = 7.11x10-5(P/V)0.56Φ1.64 (1) This study has been extended to study Saccharomyces cerevisiae in a growing system (Lamaignere 2002). The growth of S.cerevisiae in the presence of increasing solids loading of quartzite over a 28 hour time period is illustrated in Figure 1. The reduction in both the rate and extent of growth with increased solids loading is clearly illustrated by a decrease in the stationary phase population (Figure 1), a decrease in specific growth rate µmax and the decrease in biomass yield, YX/S (Table 2). A critical solids loading exists 362
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below which the solids loading exhibits an insignificant effect on cell growth. At a 1.5 and 2% solids loading, a decrease in the performance was observed with time. At 5% (v/v) solids loading, growth ceased and the cell number decreased. These observations are consistent with the hypothesis of increasing cell death with increasing volume fraction of inert particles.
Cells (million/mL)
70 60
0%
50
0.5%
40
1%
30
1.5%
20
2%
10
5%
0 0,0
10,0
20,0
30,0
Time (h)
Figure 1. Effect of solid loading (quartzite) across the range 0 to 5% on a volume basis on total Saccharomyces cerevisiae cell concentration in the agitated, aerated slurry bioreactor under growth conditions (system 4) Table 2. Effect of solid loading across the range 0 to 5% on a volume basis on growth parameters of Saccharomyces cerevisiae (system 4) Solid loading
tlag (h)
0.0% 0.5% 1.0% 1.5% 2.0% 5.0%
3.0 (±1.0) 4.0 3.5 (±0.5) 5.0 6.5 0.0
µmax exp (h-1)
0.254 (±0.026) 0.158 0.135 (±0.018) 0.134 0.112 0.000
Yx/s (109 cells/g)
7.53 (±1.08) 5.87 7.15 (±0.96) 5.90 5.50 0.00
% of viable cells at the end of the growth 98.2 (± 1.3) 94.6 88.5 (± 5.6) 93.0 85.5 2.6
The influence of the solids loading on the acidophilic iron and sulphur oxidising mesophiles was studied by exposure of the microbes to agitation in the presence of varying volume fractions of quartzite (nominal diameter 53 µm) over a 2-hour period. The growth and rate of ferrous iron oxidation in the absence of a solid phase was monitored following this exposure to identify any effect on microbial performance. The results, shown in terms of change in Eh due to a change in the ratio of ferrous to ferric iron, are given in Figure 2. Here it is clearly seen that microbial damage or reduction in physiological condition of the bacteria induced by exposure to agitation in the presence of the solids phase caused an increased lag phase or period of adaptation prior to active metabolism of the bacteria. This lag phase was extended with increasing volume fraction of solids. The acidophilic thermophilic microorganisms exhibiting iron and sulphur oxidation potential are archae, possessing a different cell envelope structure to both the bacteria and yeasts. Owing to their perceived reduced structural resilience, comprehensive studies have been undertaken to quantify their performance under varied solids loading conditions. Nemati and Harrison (2000) reported the growth and performance of Sulfolobus metallicus BC under conditions of increasing pyrite loading across the range 3 to 18% (w/v) using a 363
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pyrite size fraction of 53 to 75 µm in a one litre stirred tank reactor at an agitation rate of 500 to 550 rpm. The results reported, summarised in Figure 3 in terms of iron release, showed little effect of solids loading in the range 3 to 9% loading, a two phase leaching rate in the presence of 12 and 15% solids in which the leach rate was further impaired in stationary phase over exponential growth. At 18% loading, the system failed when an agitation rate of 550 rpm was applied and rapid cell death was observed. While microbial growth and physicochemical conditions were monitored, the changing physicochemical conditions did not allow the contribution of the solids loading to the reduced leaching performance to be established clearly. 750 Control Solids fraction: 0.5 Solids fraction: 0.10 Solids fraction: 0.20
700
Eh (mV)
650
600
550
500
450
0
20
40
60
80
100
120
Time (hours)
Figure 2. The rate of ferrous iron oxidation, given in terms of Eh, of the acidophilic iron and sulphur oxidising mesophiles subsequent to their exposure to and agitation in the presence of increasing volumetric loadings of quartzite of nominal diameter of 53 µm and an agitation rate of 772 rpm (system 1)
Figure 3. The effect of mineral pulp density on the bioleaching of pyrite through Sulfolobus metallicus in a 1-litre laboratory stirred tank reactor In order to establish the effect of solids loading while restricting the changing of physicochemical conditions, two further studies of the performance of Sulfolobus metallicus BC were conducted. In these, a constant mineral concentration was maintained of 3% (w/v) pyrite and 3% (w/v) chalcopyrite respectively. In each study, the solids loading was varied by the addition of inert quartzite of the same particle size distribution (38 to 53 µm nominal particle diameter) in 3% (w/v) increments across the range 0 to 24% and 0 to 18% for the pyrite and chalcopyrite studies respectively. The resultant bioleach 364
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0,12
100
0,1
80
0,08
extent of Fe solubilisation (%) .
Leach rate (kg Fe/m3/h) .
rates for these studies are presented in Figures 4 and 5 respectively. The leach rate of pyrite decreased from 0.113 kg.m-3.h-1 (r2 = 1.00) in the absence of quartzite through 0.095 kg.m-3.h-1 (r2 = 0.98) in the presence of 6 to 15% quartzite to 0.057 and 0.035 kg.m-3.h-1 (r2 = 0.98 and 0.99) in the presence of 21 and 24% quartzite respectively. Extent of leaching achieved varied from 91% in the absence of quartzite through 86% in the presence of 6 to 15% quartzite to 62% and 35% in the presence of 21 and 24% quartzite respectively. Similarly, on the leaching of 3% chalcopyrite in the presence of increasing loadings of inert quartzite across the range 0 to 18%, a similar decrease in the rate of solubilisation of both iron and copper was displayed (Figure 5). The rate of copper solubilisation decreased from 0.032 kg.m-3.h-1 in the absence of quartzite to 0.029 kg.m-3.h-1 in the presence of 9% quartzite. Thereafter the rate decreased linearly to 0.022 kg.m-3.h-1 in the presence of 18% quartzite. The rate of iron solubilisation decreased from 0.054 kg.m-3.h-1 in the absence of quartzite to 0.039 kg.m-3.h-1 in the presence of 9% quartzite. Thereafter the rate decreased linearly to 0.022 kg.m-3.h-1 in the presence of 18% quartzite.
60
0,06
40
0,04
20
0,02 0
0 3
9
15
18
21
24
27
solids loading (%) (w/v) leach rate
extent
Figure 4. Bioleach rate and extent of iron solubilisation as a function of total solids loading (comprised of 3% pyrite, the remainder quartzite) on leaching of pyrite in the presence of S.metallicus B.
0,035 0,03 0,025 0,02 0
5
10
15
20
Solids loading (%) (w/v)
25
Leach rate (kgFe/m3/h)
3
Leach rate (kgCu/m /h)
A. 0,055 0,045 0,035 0,025 0,015 0
5
10
15
20
Solids loading (%) (w/v)
Figure 5. Effect of solids loading on the rate of (A.) copper leaching and (B.) iron leaching of chalcopyrite in the presence of Sulfolobus metallicus on exposure to 3% chalcopyrite and 0 to 18% inert quartzite Microbial cell growth was determined in terms of the planktonic cell concentration. Both Nemati and Harrison (2000) and Sissing and Harrison (2003) have shown that planktonic cells account for the dominant active microbial population under the tank leaching conditions employed. In Figures 6 and 7, the specific growth rate determined during the exponential growth phase (typically in the range 20-70 hours), is shown to decrease as a function of increasing solids loading in the presence of 3% (w/v) pyrite and chalcopyrite respectively. In the presence of pyrite, the decreasing growth rate is reported 365
25
Bioleaching Applications
at solids loadings in excess of 9% (> 6% quartzite) while in the presence of chalcopyrite this decrease is seen at a loading of 9% and greater. At both a 27% solids loading in the presence of 3% pyrite and a solids loading of 15% or greater in the presence of 3% chalcopyrite, negative specific growth rates reported illustrate cell death under extreme hydrodynamic stress. Similarly this was reported at a solids loading of 18% by Nemati and Harrison (2000). 0,025 growth rate (h-1)
0,020
1,5
0,015
1
0,010
0,5
0,005 0,000
biomass yield (X/Fe) (1e14 cells/kg Fe)
2
0 3
9
18 24 solids loading (%)
growth rate
27
yield (X/Fe)
Figure 6. Microbial growth rate and biomass yield (YX/Fe) in terms of microbial cells produced per kg iron oxidized as a function of total solids loading in system 2 (comprised of 3% pyrite, the remainder quartzite)
-1
Growth rate (h )
0,025 0,02 0,015 0,01 0,005 0 -0,005 -0,01 0
5
10
15
20
25
Solids loading (%)
Figure 7. Microbial growth rate as a function of total solids loading in system 3 (comprised of 3% chalcopyrite, the remainder quartzite) 3.2 The influence of particulate size on performance The particle size of the particulates present influences the nature of the interaction between the microbial particle and the non-biological particle in terms of frequency of collision, momentum of collision and path of the particle with respect to fluid flow in the reactor. Further the size distribution of the particulate phase may influence the physicochemical properties of the suspension. In our initial studies of the yeast system, the influence of particle size on the disruption of stationary phase yeast was studied. This data is presented in terms of the first order rate constant for disruption (k) and the extent of disruption achieved during a 2-hour exposure in Figure 8. Below a particle size of 300 µm, the disruption rate constant is less than 20% of that achieved at greater particle diameters while complete cell disruption is not achieved in a 2-hour period. Increase of the particle size beyond 700 µm showed no 366
Bioleaching Applications
0,0007
90 80 70 60 50 40 30 20 10 0 1500
0,0006 k (1/s)
0,0005 0,0004 0,0003 0,0002 0,0001 0 0
500
1000
Ri/Rm (%)
further effect on the disruption rate constant. These findings were consistent with the bead sizes recommended for optimum cell disruption in the bead mill (Schutte et al. 1986). Harrison et al. (2003) have illustrated that this dependence of disruption rate constant on particle size can be correlated in terms of particle momentum.
Particle diameter (micrometers)
k (1/s) L (k (1/ ))
Ri/Rm (%) L (Ri/R (%))
Figure 8. The influence of particulate size (given as geometric mean diameter) on the disruption of stationary phase Saccharomyces cerevisiae (system 5) on agitation at an impeller tip speed of 2.2 m.s-1 (772 rpm) and a solids volume fraction of 0.20 is given in terms of the first order disruption rate constant (k) and the extent of disruption expressed as the fraction of protein released over that available for release (Ri/Rm) The effect of particle size across the range 53 to 255 µm on the hydrodynamic stress response of acidophilic iron and sulphur oxidising mesophiles on agitation in the presence of a 0.05 volume fraction of particulates is presented in Figure 9. Samples were taken before and after a 2-hour agitation period in the stirred tank reactor and their performance monitored by ferrous iron oxidation in shake flask culture in the absence of a solid phase over 5 days. In all cases, a lag phase was induced by the exposure of the culture to agitation in the presence of particulates. These data are consistent with Figure 2. A similar lag period was induced across mean particle sizes of 53 to 161 µm. 750 Control 53 micron 114 micron 161 micron 255 micron
700
Eh (mV)
650
600
550
500
450 0
20
40
60
80
100
120
140
Time (hours)
Figure 9. The rate of ferrous iron oxidation of the acidophilic iron and sulphur oxidising mesophiles subsequent to their exposure to and agitation in the presence of quartzite of varying size, given as nominal diameter, at a volume fraction of 0.05 and an agitation rate of 772 rpm (system 1) 367
Bioleaching Applications
Studies on the effect of particle size on the leaching of pyrite in the presence of Sulfolobus metallicus on both the microbial growth and leaching performance have been conducted. Nemati et al. (2000) report the increasing rate of leaching with decreasing particle size across a particle size range of 37 to 150 µm nominal diameter. This is in accordance with predictions based on enhanced particle surface area and thereby area for reaction to occur. Their studies were extended to consider smaller particle size distributions. At a nominal mean particle diameter less than 25 µm, it has been found that both pyrite leaching performance as well as specific growth rate of Sulfolobus metallicus is reduced with decreasing particle size (Figure 10). These studies suggest that the particle size distribution may also affect the physicochemical properties of the slurry to provide a limiting environment for bioleaching. For example, it is well known that slurry viscosity increases inversely to particle size (Thomas 1965). Further study is required to provide a mechanistic understanding of these findings. 6,0E-02
1,2E-01 5,0E-02
-1
GROWTH RATE (hr )
FE OXIDATION RATE (g/l/hr)
1,0E-01 8,0E-02 6,0E-02 4,0E-02
4,0E-02
3,0E-02
2,0E-02
2,0E-02 1,0E-02
0,0E+00 0
20
40
60
80
100
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140
160
180
200
0,0E+00 0
-2,0E-02
MEAN PARTICLE VOLUME DIAMETER (um)
20
40
60
80
100
120
140
160
180
200
MEAN PARTICLE VOLUME DIAMETER (um)
Figure 10. The influence of pyrite particulate size, given as nominal particle diameter, on the iron leaching rate and specific growth rate of Sulfolobus metallicus (system 2) 3.3 The influence of energy dissipated through agitation on performance The influence of agitation rate in the slurry bioreactor was investigated in system 1 (acidothiobacilli) over the range 400 to 1100 rpm, corresponding to impeller tip speeds in the range 1.1 to 3.1 m.s-1. As shown on investigation of other solids parameters, agitation in the presence of the particulate phase resulted in an extended lag phase in subsequent culture of some 40 hours (Figure 11). While little difference is seen between the performance at agitation rates of 400 and 770 rpm (impeller tip speeds of 1.1 and 2.2 m.s1 ), the lag phase was further extended on agitation at 1090 rpm, corresponding to an impeller tip speed of 3.3 m.s-1. Further, the extent of reaction was also reduced. Previously Hackl et al. (1989) reported impairment of bioleaching performance at impeller tip speeds of 5.3 m s-1 relative to that at 3.3 m.s-1. The impairment at lower impeller tip speeds in our experiment is expected owing to the use of a Rushton impeller. The effect of agitation rate on the bioleach performance in the presence of Sulfolobus metallicus was investigated by increasing the agitation rate in system 2 in the range above the critical impeller speed used to ensure fully suspended solids. The agitation rates investigated were 560, 660 and 760 rpm, corresponding to tip speeds of 1.67, 1.97 and 2.27 m.s-1 respectively, with a solids loading of 3% pyrite and 15% quartzite. These results are presented in Figure 12. The system failed at an agitation rate of 760 rpm (tip speed 2.27 m.s-1), where the rate of damage to the cells exceeded their growth rate. The rate of iron release and the extent of pyrite solubilisation was higher at 660 rpm than at 560 rpm due to the higher initial microbial cell concentration at 660 rpm. However, the specific growth rate of the microorganisms was slightly lower at the higher agitation rate in 368
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accordance with the increased hydrodynamic stress. The specific activity of the microorganisms at 560 rpm was higher than that at 660 rpm, suggesting an adverse effect of higher agitation rate on the microorganisms. This may be due to increased damage on increased energy dissipation at the higher agitation rate. A specific cell death rate of 0.026 h-1 was observed at an agitation rate of 760 rpm. 750
700
Eh (mV)
650
600
550
Control 772 rpm 1090 rpm
500
450
0
20
40
60
80
100
Time (hours)
Figure 11. The rate of ferrous iron oxidation of the acidophilic iron and sulphur oxidising mesophiles subsequent to their exposure to and agitation in the presence of quartzite at varying agitation rate and a volume fraction of 0.05 (system 1)
100
0,1
80
0,08
60
0,06
40
0,04
20
0,02 0
0 560
660
760
agitation rate (rpm) rate
16 14 12 10 8 6 4 2 0
0,04 specific growth rate (h-1)
0,12
0,02 0,00 -0,02 -0,04 -0,06 560
660
activity (1e-17 kg Fe/cells.h)
B extent of solubilisation (%)
rate (kg Fe/m3.h)
A
760
agitation rate (rpm)
extent
growth rate
activity @ t=50h
Figure 12. Performance of Sulfolobus metallicus BC as a function of agitation rate under conditions of complete suspension and a solids loading of 3% pyrite and 15% quartzite (w/v): A. Rate of iron release and extent of pyrite dissolution B. Specific growth rate and microbial activity Energy dissipation rate was investigated with respect to the model Saccharomyces cerevisiae slurry system through varying the agitation rate across the range 460 to 850 rpm, equivalent to impeller tip speeds in the range 1.9 to 3.5 m.s-1, using system 4. As illustrated in Table 3, an optimum agitation rate was found to be in the impeller tip speed range of 2.3 m.s-1. While cell viability reduced with increasing impeller tip speed, it is apparent that a minimum energy dissipation rate was required to provide the mass transfer and mixing necessary for maximum specific growth rate.
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Table 3. Effect of impeller speed on growth parameters of Saccharomyces cerevisiae (system 4) Agitation rate (rpm)
Impeller tip speed (m s-1)
Lag time (h)
460
1.9
565
2.3
600 850
2.5 3.5
4.0 3.5 (± 0.5) 3.5 4.0
Max. specific growth rate (h-1) 0.077 0.135 (± 0.018) 0.125 0.114
Stationary phase cell concentration (106 cells ml-1) 38
YX/S (109 cells g-1)
5.14 7.15 (± 0.96) 6.17 4.92
55 50 43
% of viable cells at the end of the growth 94.7 88.5 (± 5.6) 86.2 77.4
3.4 Response of the biophase to solids-induced stress in the slurry bioreactor In order to ensure that the response to hydrodynamic stress may be predicted, and where possible negative responses may be overcome, it is necessary to generate an understanding of these at the cellular level. Studies have been initiated to investigate the influence of hydrodynamic stress generated in the slurry reactor on metabolic activity, cell morphology and metabolic efficiency. During the exponential growth phase of Sulfolobus metallicus on pyrite in the presence of quartzite, Figure 13 illustrates that a higher specific activity (rate of iron released per microbial cell) was observed at lower solids loading, i.e. 0-15% quartzite exhibited higher activity than 21% quartzite (24% total loading), and 24% quartzite (27% total loading) exhibited the lowest activity. The similarity in activity across the solids loading range during the stationary phase suggested that the efficiency of the microorganisms in converting ferrous iron present to ferric iron at the various solids loading was similar in the absence of microbial growth. However, the microbial cell concentration decreased with solids loading. Hence the decrease in pyrite oxidation with solids loading in the stationary phase corresponded to a lower microbial cell concentration as opposed to lower microorganism activity. In the exponential growth phase, both reduced specific activity and a decreased biomass concentration contributed to the decreased leaching performance. activity (1e-17 kg Fe / cells.h)
30 25 20 15 10 5 0 0
3
6
9
12
15
18
21
24
27
30
total solids loading (%) (w/v)
20 h
50 h
100 h
Figure 13. Microbial cell activity in terms of specific pyrite oxidation rate as a function of solids loading and duration of experiment Further, changes in morphology of Sulfolobus in response to hydrodynamic stress were reported qualitatively by Nemati and Harrison (2000). The change in morphology of 370
Bioleaching Applications
Sulfolobus metallicus has been observed quantitatively in terms of cell size in response to growth on chalcopyrite in the presence of varying quartzite loadings. As illustrated in Figure 14, Sulfolobus metallicus is typically found to have a cell diameter of approximately 1.0 to 1.1 µm under conditions of good physiological status. Exposure to increased hydrodynamic stress, associated with reduced leaching performance (Figures 4 and 5) and reduced specific growth rates (Figures 6 and 7), were accompanied with a decrease in the cell size shown in Figure 14. The decrease of some 25% on a diameter basis corresponds to a decrease of some 40% on a volume basis. This is consistent with the reduced cell size and change in phase brightness reported on growth of Sulfolobus metallicus in the presence of pyrite only while investigating both solids loading and size effects (Nemati and Harrison 2000, Nemati et al. 2000). 1.1 1.05 Xo (um)
1 0.95 0.9 0.85 0.8 0.75 0.7 0
5
10 15 20 Solids loading (%) (w/v)
25
Figure 14. Effect of solids loading on diameter of Sulfolobus metallicus as determined on exposure to 3% chalcopyrite and 0 to 18% inert quartzite, using the Cell Facts cell counter Further the yield of Sulfolobus metallicus biomass based on iron solubilised (YX/Fe) decreased with increased solids loading in the presence of 3% pyrite (Figure 6). From the results of similar studies conducted with Saccharomyces cerevisiae in the slurry bioreactor under growth conditions (Table 1), a decreasing biomass yield (YX/glucose) was also reported with increasing solids loading across the range 0 to 5%. Concomitantly, a decrease in cell viability was found. As the solubilisation of iron and glucose, respectively, are required for energy generation in the two systems investigated, the decreased yield indicated a decrease in the fraction of energy generated that was used for cell synthesis, suggesting a greater energy requirement for cell maintenance. Using transmission electron microscopy, analysis was conducted to detect change in cell shape and size following growth of Saccharomyces cerevisiae under hydrodynamic stress in the slurry bioreactor. Yeast cell shape was modelled as an ellipsoid. No significant difference could be detected in mean length and mean width on comparison of growth at 0% and 1% solid loading (Table 4). Dimensions recorded were in good agreement with the literature (Srinorakutara 1998, Smith et al. 2000). However the appearance of the yeast differed with the level of stress applied. Figure 15 compares S.cerevisiae grown in the absence of a solid phase (at t=28h) and at 1% solids loading (at t=28h). Cells grown with 1% solid are visibly damaged compared to those grown in the absence of solids. Cell walls are not clearly defined and cell shape is less regular.
371
Bioleaching Applications
Table 4. Size and shape of S.cerevisiae as a function of solids loading during growth (system 4) 0% solid 1% solid Srinorakutara (1998) Smith et al. (2000)
Mean length (nm) 3678 ± 266 3614 ± 710 3330 ± 70 3420 ± 620
Mean width (nm) 2738 ± 400 2797 ± 415
Figure 15. Comparison between yeast grown in (a) the absence of a solid phase and (b) at 1% solid loading at t=28h (system 4) Cell wall thickness was measured across a range of samples taken under three levels of hydrodynamic stress: agitation at 560 rpm in the absence of a solids phase, agitation at 560 rpm in the presence of 1% quartzite loading, and agitation at 850 rpm in the presence of 1% quartzite. The cell wall thickness of 71 ± 11 nm determined in the absence of hydrodynamic stress compared well with literature values (Moor and Muhlehaler 1963, Srinorakutara et al. 1998). On increasing the hydrodynamic stress, an increase in cell wall thickness was observed (Figure 16). In quantifying the cell wall thickness, it was observed that a certain percentage of cells exhibit a thin wall even for 1% solid and these increased with time. As dead cells could not be differentiated from living cells with TEM analysis, it was postulated that the thin walled cells were dead cells, while cells adapting to the hydrodynamic stress required thickened cell walls.
Cell wall thickness (nm)
160 140 120 100 80 60 40 20 0 #2 (3h)
#5 (14h)
#7 (28h)
Time
Figure 16. Influence of time and hydrodynamic stress on cell wall thickness (g 0% solid loading, g 1% solid loading) 372
Bioleaching Applications
4.
CONCLUSIONS Biohydrometallurgy is an important example of the use of microorganisms in slurry bioreactor systems. In such systems, efficient microbial performance is essential while exposed to the increased hydrodynamic stress of the slurry environment. In this study, we have investigated the influence of this particulate phase on three microbial systems under differing operating environments in order to determine both the influence of the slurry system on active minerals bioleaching, as well as to seek a generic understanding of responses found. In all cases, a critical solids loading was found in the laboratory scale reactor above which the process performance is impaired. The critical solids loading determined varied as a function of microorganism used, its growth phase as well as the mineral phase used. In all cases, decreased performance was associated with a decreased microbial phase. Further, reduced microbial activity was reported in the growing systems. Reduced particulate size distribution resulted in improved process performance in the mineral system across the range 37 to 150 µm through the provision of an increased surface area. Further the yeast study suggested that a minimum particle size to cell ratio of approximately 45-75 is required for disruption of the microorganisms. This is in accordance with bead mills studies. Poor performance of the bioleaching system at very low particle size distribution (<15 µm in diameter) was unexpected. It is postulated to result from altered physicochemical properties of the suspension resulting in process limitation. Yeast cell disruption can be modelled as a power law function of energy input per unit volume (Scholtz-Brown et al. 1997). Similarly Lamaignere has proposed that the cell death constant in the growing S.cerevisiae system is a function of power input per unit volume. In the mineral bio-oxidation studies, an optimum agitation rate is proposed for maximum performance. It is postulated that lower agitation rates lead to process limitation while increased rates result in cell damage or death. Correlation of damage in terms of energy dissipation rate and impeller tip speed remains to be investigated. At the microbial level, biological response to hydrodynamic stress has been shown to influence cellular structure and morphology through reduced cell size (S. metallicus), irregular appearance (S. metallicus and S. cerevisiae) and a thickened cell wall (S. cerevisiae). Further, an increasing first order death rate constant is proposed, thereby decreasing the apparent specific growth rate. In all cases, reduced performance is associated with a reduced microbial phase. Under specific conditions, the activity of this phase is also reduced. Further, it is apparent that the biomass yield coefficient is reduced on exposure to hydrodynamic stress. From this observation, it is postulated that an increased maintenance energy results to overcome the stress to which the cells are exposed. ACKNOWLEDGEMENTS The researchers gratefully acknowledge the sponsorship of BHP-Billiton, the Department of Trade and Industry, South Africa through the THRIP programme, and the National Research Foundation of South Africa.
373
Bioleaching Applications
REFERENCES 1. Bailey A.D. and G.S. Hansford (1993). Factors affecting biooxidation of sulphide minerals at high concentrations of solids: a review. Biotechnology and Bioengineering, 42(10), 1164-1174 2. Clark D.A. and P.R. Norris (1996). Oxidation of mineral sulphides by thermophilic microorganisms. Minerals Engineering, 9(11), 1119-1125 3. Escobar B, J.M. Cassas, J. Mamani and R.B. Ohlbaum (1993). Bioleaching of a copper concentrate with Sulfolobus BC. In Biohydrometallurgical Technologies (A.E. Torma, J.E. Wey and V.L. Lakshmanan (ed)), The Minerals, Metals and Materials Society, pp 195-204. 4. Gormley, L.S. & Branion R.M.R., Engineering design of microbial leaching reactors. In Proceedings of the International Symposium on Biohydrometallurgy, pp. 499-515 (1989). 5. Hackl R.P., Wright F.R. and Gormley L.S. (1989). Bioleaching of refractory gold ores – out of the lab and into the plant. Proceedings of the International Symposium on Biohydrometallurgy, Eds. Salley, McReady, Wichlacz. Wyoming USA. Pp533-549. 6. Harrison S.T.L., Scholtz N.J. and Pearce S.J.A. (2003). The effect of inert particulate parameters on microbial cell disruption in a slurry bioreactor. J. Chem Technol. Biotechnol. (submitted). 7. Harrison S.T.L., Sissing A., Raja S., van der Merwe S. and Nemati M. (2003). Identifying and quantifying biological responses of Sulfolobus to high pulp densities in the slurry bioreactor. International Minerals Processing Congress, September 2003. Cape Town. 8. Jordan M.A., D.W. Barr and C.V. Phillips (1993). Iron and sulphur speciation and surface hydrophobicity during bacterial oxidation of a complex copper concentrate. Minerals Engineering, 6 (8-10), 1001-1011 9. Kelly R.M. and J.W. Deming (1988). Extremely thermophilic archaebacteria: Biological and engineering considerations. Biotechnology Progress, 4(2), 47-62 10. König H. and K.O. Stetter (1986). Studies on archaebacterial S-layers. System. Appl. Microbiol, 7, 300-309 11. König H. (1988). Archaebacterial cell envelopes. Canadian Journal of Microbiology, 34, 395-406 12. Lamaignere V. (2002). Effects of hydrodynamic stress on growing Saccharomyces cerevisiae in a slurry bioreactor. MSc dissertation, Department of Chemical Engineering, University of Cape Town. 13. Le Roux N., W. and D.S. Wakerley (1988). Leaching of Chalcopyrite (CuFeS2) at 70°C using Sulfolobus. Proceedings of Biohydrometallurgy ’87, Eds. P.R. Norris and D.P. Kelly, Science and Technology Letters, Surrey, United Kingdom, 305-317 14. Logan B.E and Dettmer J.W. (1990). Increased mass transfer to micro-organisms with fluid motion. Biotechnol. Bioeng., 35, 1135-1144. 15. Lowry, O.H., Roseborough, N.J., Farr, A.L. & Randall, R.J., Protein measurement with the Folin reagent. J. Biol. Chem. 193: 265-275 (1951). 16. Michel H., D.-Ch. Neugebauer and D. Oesterhelt (1980). The 2-D crystalline cell wall of Sulfolobus acidocaldarius: structure, solubilisation and reassembly. In Electron Microscopy at Molecular Dimensions (W. Baumeister, W. Vogell (ed)), SpringerVerlag, Berlin, Heidelberg, pp27-35
374
Bioleaching Applications
17. Nemati M. and S.T.L. Harrison (2000). Effect of solid loading on thermophilic bioleaching of sulphide minerals. Journal of Chemical Technology and Biotechnology 75, 526-532 18. Nemati M., Lowenadler J. and S.T.L. Harrison (2000). Particle effects in bioleaching of pyrite by acidophilic thermophile Sulfolobus metallicus. Appl. Microbiol. Biotechnol., 53, 173-179. 19. Norris P.R. and D.W. Barr (1988). “Bacterial oxidation of pyrite in high temperature reactors. Proceedings of Biohydrometallurgy ‘87, Eds. P.R., Norris and D.P. Kelly, Science and Technology Letters, Surrey, United Kingdom, 532-536 20. Oguz H., A. Brehm and W.D. Deckwer (1987). Gas/liquid mass transfer in sparged agitated slurries. Chemical Engineering Science, 42, (7), 1815-1822 21. Pearce, S.J.A., Disruption of microorganisms due to agitation in slurries of fine particles. M.Sc dissertation, University of Cape Town, (1993). 22. Scholtz N J, Pandit A B and Harrison S T L (1997). Effect of solids suspension on microbial cell disruption. In Bioreactor & Bioprocess Fluid Dynamics, (ed. A W Nienow), pp 199 - 215. 23. Sissing A and Harrison S T L (2003). Influence of solids loading and agitation intensity on thermophilic mineral bioleaching performance: studies on a pyritequartzite system. J. Chem. Technol. Biotechnol. (submitted) 24. Smith, A.E., Moxham, K.E., and Middelberg, A.P.J (2000a), Wall material properties of yeast cells. Part II. Analysis, Chemical Engineering Science, 55, 2043 – 2053. 25. Srinorakutara, T. (1998), Determination of Yeast Cell Wall Thickness and Cell Diameter Using New Methods, Journal of Fermentation and Bioengineering, Vol.86, 3, 253 – 260. 26. Toma, A.E., Ruklisha, M.P., Vanags, J.J., Zeltina, M.O., Leite, M.P., Galinina, N.I., Viesturs, U.E. & Tengerdy, R.P., Inhibition of microbial growth and metabolism by excess turbulence. Biotechnol. Bioeng. 38: 552-556 (1991). 27. Torma, A.E., Walden, C.C., Duncan D.W. & Branion R.M.R., The effect of carbon dioxide and particle surface area on the microbiological leaching of a zinc sulphide concentrate. Biotechnol. Bioeng. 14: 777 (1972). 28. Torres F., M.L. Bláquez, F. González, A. Ballester and J.L. Mier (1995). The bioleaching of different sulphide minerals using thermophilic bacteria. Metall. and Materials Trans. B, 26B (June), 455-465 29. Vogel, A.I. (1989), Vogel’s Textbook of Quantitative Chemical Analysis, 5th ed., Longman Group Ltd., London, UK, 287-310
375
15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
The development of a hybrid biological leaching-pressure oxidation process for auriferous arsenopyrite/pyrite feedstocks I. Dymov , C.J. Ferron and W. Phillips*
SGS Lakefield Research Limited 185 Concession St., POB 4300, Lakefield Ontario K0L 2H0 Canada *Kinross Gold Corporation 40 King St. West, Toronto, Ontario M5H 3Y2 Canada (former Consultant to TVX)
Abstract A process has been evaluated, on a continuous pilot plant scale, to treat refractory gold arsenopyrite/pyrite concentrates. The process consisted of biological leaching to partially oxidize sulphides, followed by pressure leaching for complete sulphide oxidation and arsenic precipitation, followed by liquor neutralization and cyanidation of the autoclave residue for gold recovery. This paper provides a brief description of the treatment flowsheet and discusses some of the main parameters and results of the process. Keywords: refractory gold, arsenopyrite/pyrite, piloting, biological leaching, pressure oxidation, gold recovery, arsenic precipitation
1.
INTRODUCTION For several years, TVX Hellas Company has been developing the Olympias property in Greece. The deposit consists of a pyrite concentrate stockpile, a zinc flotation tailings stockpile and in-situ ore reserves. Run-of-mine ore was proposed to be milled and floated in a two-stage flotation circuit to produce lead and zinc concentrates, followed by flotation of the remaining sulphides in the zinc tailings to produce a gold-containing arsenopyrite/pyrite concentrate. Previous studies indicated that pressure oxidation of the Olympias pyrite concentrate, as compared to bacterial oxidation, resulted in similar or slightly higher gold recovery at significantly lower cyanide consumption. Capex estimate of the pressure oxidation process was higher than bacterial oxidation, but a significant portion of the arsenic was precipitated in the autoclave as an environmentally stable compound. A feasibility study was conducted at SGS Lakefield Research facilities to investigate the possibility of treating the refractory pyrite concentrates using a combination of bacterial oxidation (BIOX®) and pressure oxidation (POX) technologies. By combining the two technologies, the majority of the sulphide sulphur present in the concentrate (approximately 70%) could be oxidized with air, by applying the BIOX® process, leaving a relatively small portion of the sulphides to be oxidized with oxygen in an autoclave, to complete the oxidation and precipitate the arsenic. 377
Bioleaching Applications
The BIOX® section of the dual process could be designed to produce a partially oxidized product containing sufficient residual sulphur to ensure auto-thermal operation of the autoclave. Based on extensive bench scale testwork and previous pilot plant campaigns, the recommended process flowsheet is illustrated in Figure 1. ROM Ore
Milling
Lead Flotation Zinc Flotation
Reclaim Tails
Lead Concentrate Zinc Concentrate
Pyrite Flotation Pyrite Flotation
Stockpile Conc
Flotation Tailings
Flotation Concentrate Wash Circuit Tailings Dam
Regrind Circuit
Neutralization BIOX® BIOX® Stock Tank
BIOX® Thickener Overflow
BIOX® Product
CN Destruction
Vent
POX Stock Tank Quench Water
POX Thickener Overflow
Autoclave Oxygen
CCD
Gold Recovery Carbon in Leach
Figure 1. Suggested flowsheet for the treatment process of the Olympias concentrate As a part of the bankable feasibility study, TVX Hellas conducted a series of fully integrated pilot plant campaigns at SGS Lakefield Research in 1999 to 2000 to simulate the flowsheet illustrated in Figure 1. Each unit process of the pilot plant will be discussed in sequence. The base metal flotation section, and concentrate wash circuit are not discussed in this paper. TVX Hellas constructed a BIOX® pilot plant, according to the design and equipment specifications provided by GoldFields Limited to treat 100 kilograms concentrate per day. The partially oxidized BIOX® product was used as feedstock for the continuous autoclave testing. The oxidation pilot plant circuit consisted of the following unit operations: 1. BIOX® circuit 2. Thickening (by decantation) of the BIOX® discharge to remove ~10% BIOX® liquor volume (this portion of the liquor by-passed the autoclave and was forwarded directly to the neutralization circuit). 378
Bioleaching Applications
3. Pressure oxidation of the thickened, partially oxidized BIOX® pulp (plus a small portion of untreated pyrite concentrate) in a continuous pilot plant autoclave operation. 4. Thickening/washing of the POX discharge to produce PLS for neutralization. 5. Continuous neutralization of the POX thickener overflow and a portion of the BIOX® liquor. 6. Thickening of the neutralized product in a ‘high density sludge’ mode. 7. Cyanidation/CIL of the washed autoclave discharge samples and compartment samples. These tests were conducted in a batch mode. The overall objectives of the pilot plant campaign were to: evaluate and optimize the BIOX® circuit; evaluate the behaviour of the partially oxidized BIOX® slurry in the autoclave; evaluate arsenic dissolution/precipitation in the autoclave; evaluate gold recovery; neutralize POX and BIOX® liquor to produce environmentally acceptable effluent and stable waste product.
2.
DESCRIPTION OF FEED SAMPLES It was proposed that the oxidation plant would treat pyrite concentrates from three sources: Stockpile and reclaimed tailings pyrite concentrate, run-of-mine (ROM) pyrite concentrate, and so-called post-East concentrate (PEC). The average chemical analyses of the three concentrate samples treated are shown in Table 1.
Table 1. Chemical analyses of the concentrate samples Concentrate
Au, g/t
Ag, g/t
S=, %
Fe, %
As, %
Stockpile ROM Blend PEC Blend
23 23 25
22 32 41
39 29 32
38 30 32
8.6 11.2 10.8
Fe:As molar ratio 6.0 3.6 3.9
Continuous flotation pilot plant campaigns were conducted at SGS Lakefield Research on the ROM 2001-2003 and PEC ores. The flowsheet involved the use of cyanide as pyrite depressant during base metal flotation. It was observed that the cyanide reacted with sulphides to form thiocyanate, which "adsorbed" onto the pyrite concentrate surfaces. Thiocyanate is highly toxic to the BIOX® bacteria, even at a fairly low concentration in solution. Therefore, washing of the concentrate prior to the BIOX® circuit was found to be extremely important.
3.
PILOT PLANT DEMONSTRATIONS
3.1 BIOX® circuit The BIOX® program and circuit operation were directed by TVX Hellas and GoldFields Mining Services Ltd. representatives. The circuit was operated on a 24-hour basis for 160 days by TVX and SGS Lakefield personnel. The main objective of the continuous BIOX® circuit was to produce semi-oxidized BIOX product as feed to the continuous POX circuit. These pilot plant campaigns presented the first opportunity to treat the fresh concentrate blends in a continuous mode. All previous piloting had been conducted on the Stockpile concentrate only. The BIOX process uses a mixed culture of thiobacillus ferrooxidans, thiobacillus thiooxidans and leptospirillum ferrooxidans to break down the sulphide mineral matrix. The active 379
Bioleaching Applications
inoculum produced during small pilot plant operations in Greece was shipped to Lakefield for reactivation and inoculation in the pilot plant reactors. The BIOX® plant consisted of two identical trains of one primary reactor and two secondary reactors fabricated from stainless steel. Figure 2 is a photograph showing the BIOX® pilot plant, and Figure 3 - the continuous autoclave. The primary reactors had an unaerated volume of 622 liters each and the secondary reactors, 208 liters each. The plant was designed to treat 100 kilograms of concentrate per day, at a retention time of 5 days, 3 days in the primary stage and 1 day per each secondary stage. This arrangement allowed a reduction in retention time by taking the secondary reactors off line, should this prove to be necessary. The pilot pant was controlled at the following operating conditions: Liquid/Solid Ratio 4/1 and 5.7/1 Slurry Temperature 40 to 45°C Slurry pH 1.2 to 1.8 Dissolved Oxygen Concentration 2 to 6 mg/L The pilot plant was fed on a continuous basis at a predetermined rate to give a specific retention time. The required nutrient salts, consisting of ammonium sulphate, ammonium phosphate and potassium sulphate, were added to the feed make-up tank.
Figure 2. Continuous BIOX® plant
Figure 3. Continuous agitator assemblies
autoclave
The slurry pH was controlled by manual addition of limestone slurry. Limestone was also added to the feed make-up tank to give a total carbonate content of 3% in the feed to the plant. The oxygen uptake rate, which is an instant measurement of the rate of oxygen depletion in an active inoculum and probably the most important criterion in assessing bacterial activity, was measured routinely. Ferric and ferrous iron concentrations were also measured routinely, as well as redox potential. The following samples were regularly taken from the circuit: feed samples, profile samples from each reactor, and final products from each train. Typical results from the continuous operation are illustrated in Figure 4 and summarized in Table 2. The bacterial activity was very high during the pilot plant operation, with an average sulphide sulphur removal over 90% after 5 days; however, the sulphide oxidation to sulphate was considerably lower due to the formation of elemental sulphur or poly sulphide species. 380
Bioleaching Applications
The elemental sulphur content in the BIOX® discharge was initially in the order of 1011%, but later stabilized between 7-9%. The reasons for the presence of too much elemental sulphur were not clearly understood. The analytical data suggested that there was no clear correlation between the operating pH, iron and arsenic dissolution and the formation of elemental sulphur. It is likely that the formation of elemental sulphur in the BIOX® discharge is due to the slower kinetics of the last step of conversion of elemental sulphur to sulphate. More elemental sulphur is produced as the overall sulphide removal increases, but only a fraction of the elemental sulphur further reacts to form sulphate. The iron and arsenic dissolution averaged 60-75% and 80-90% respectively, with the BIOX® liquor containing up to 85 g/L Fe (mostly as ferric) and 40 g/L As.
Table 2. BIOX® pilot plant results Blend
1
2
3
Retention time days
Tank
Solids Analyses =
o
Solution Analyses
S %
S %
As %
Fe %
As g/L
Fe g/L
SO4 g/L
Feed
-
39
-
9.0
39
-
-
-
Primary
3
17
6
2.1
20
19
69
44
Secondary R1
1
5
9
1.3
13
21
78
48
Secondary R2
1
2
9
1.0
11
22
80
49
Feed
-
29
-
11.3
31
-
-
-
Primary
3
16
6
3.4
20
30
67
41
Secondary R1
1
5
8
1.4
11
35
82
48
Feed
-
32
-
11.0
33
Primary
3
14
7
5.3
21
28
63
36
Secondary R1
1
8
8
2.8
15
41
86
46
100
%S Oxidation
90 80 70 60 50 40 30 S= Removal
20 10
S= Oxidation to SO4
0 0
2 4 BIOX® Reactors Retention Time, days
6
Figure 4. Sulphide oxidation as a function of retention time The BIOX® circuit discharge was forwarded to a pressure oxidation stage to complete the oxidation of residual sulphides and other sulphur species. A percentage of the BIOX® solution was decanted from the slurry and bled directly to the neutralization circuit. The design target of the bleed was 10% volume, with the actual volume varying between 10 and 25%, depending on the BIOX® discharge slurry density. The BIOX® product must 381
Bioleaching Applications
contain enough residual sulphides and elemental sulphur to maintain auto-thermal operation of the autoclave. Therefore, it was established that the BIOX® plant only required a primary and one secondary stage to achieve the required level of sulphide oxidation. The residual sulphide grade in the BIOX® discharge was controlled by varying the number of reactors on-line and the reactor configuration. Based on the pilot plant results, the following plant operating parameters were recommended by GoldFields Mining Services: Feed Slurry Density 20% solids Operating pH in primary and secondary reactors 1.2 to 1.7 Operating Temperature 40 to 45°C Retention Time
Primary reactors Secondary reactors
3 days 1 day
3.2 POX circuit The main objective of the POX circuit was to oxidize all sulphides and sulphur present and to expose gold for extraction by cyanide leaching. Additional objectives were to precipitate out arsenic as a stable precipitate and to produce final effluents meeting industrial standards. The test program was designed in consultation with SNC-Lavalin and TVX Hellas representatives. SGS Lakefield’s continuous horizontal autoclave, constructed of Grade 12 titanium, is 172.7 cm in length with an inside diameter of 25.0 cm, and is divided into six compartments by means of weir plates. Oxygen gas is sparged at controllable flow rates into all compartments. The oxygen is normally distributed with greater than 80% of the total flow directed into the first two compartments. Total oxygen flow is typically in the order of 28-40 liters per minute. Three continuous integrated pilot plant campaigns were conducted on each pyrite concentrate blend, partially oxidized by the BIOX® process. Autoclave feed comprised principally of BIOX® product, with small additions of untreated pyrite concentrate in order to achieve the target sulphide plus sulphur grade of ~15%. This value was determined by the MetSim model as the minimum amount of sulphide sulphur for autothermal autoclave operation. The target solid content varied between 15 and 20% solids. The autoclave feed contained approximately 10% elemental sulphur. Quebracho or Lignosol were added at a rate of 5kg/t and 2.5 kg/t respectively, as an elemental sulphur dispersant, to prevent occlusions by sulphur of unreacted sulphide particles. Autoclave target operating conditions were 225°C, 100 psig oxygen overpressure and 30 to 70 minutes nominal residence time. Filtered and washed autoclave discharge and compartment samples were submitted for neutralization-cyanidation/CIL for the recovery of gold and silver. The tests were conducted in a batch mode. The samples were neutralized at 30% solids with hydrated lime to pH 11, for at least 12 hours in order to reach a stable pH prior to cyanidation. Activated carbon was added at 10 g/L solution and CIL was carried out at 0.5 g/L NaCN for 24 hours. A summary of the autoclave campaign results is presented in Table 3. The results indicated that sulphide sulphur oxidation was typically greater than 95% after 30 minutes in the autoclave. Elemental sulphur conversion to sulphate was greater than 98% after 10 minutes in the autoclave.
382
Bioleaching Applications
Table 3. Pressure oxidation pilot plant results Campaign 1 Parameters
POX Feed
Campaign 2
Campaign 3
POX POX POX POX Discharge Discharge Feed Feed
POX Discharge
A/C Conditions Temperature, °C
-
225
-
225 225 225
-
225
225
225
Time, min
-
70
-
70
40
30
-
65
50
36
As, g/L
20
13
15
5
8
9
14
7
7
10
Fe, g/L
65
60
48
22
24
34
48
34
35
45
SO4, g/L
140
300
80
190 190 200
93
190
185
230
FA, g/L
<10
47
<10
62
56
44
<10
41
46
37
Fe/As Molar Ratio
4.4
6.2
4.3
5.9
3.9
5.2
4.6
6.5
6.7
6.0
As, %
3
6
2
5
5
5
2
Fe, %
0.19
0.19
Solution Analyses
Solids Analyses 20
22
16
22
21
22
18
=
15
0.2
6-11
0.1
0.1
0.7
11
0.23
o
10
<0.5
7-10 <0.5 <0.5 <0.5
9
<0.5 <0.5 <0.5
S ,% S,%
Results Au Extraction, % =
98
98
98
97
96
97
98
S Oxidation, %
-
99
-
99
98
85
-
96
98
98
As precipitation, %
-
50
-
85
79
72
-
86
85
75
The fast kinetics of oxidation of sulphide and elemental sulphur indicated that the risk of sulphide occlusion by elemental sulphur was minimized by the addition of Lignosol. In association with the sulphide and sulphur oxidation, iron conversion from ferrous to ferric was greater than 98%. Similarly, arsenic oxidation to As(V) was greater than 98%; the residual As(III) solution concentration was about 120 mg/L. Kinetic profiles of sulphide oxidation versus gold recovery and arsenic dissolution/precipitation are also illustrated in Figures 5 and 6. The results showed that gold recoveries closely followed the sulphide sulphur oxidation profile. Gold extraction from the autoclave discharge was excellent and averaged 96-98%, leaving a residue assaying 0.3-0.9 g Au/t cyanidation residue. Sodium cyanide and lime consumptions for the autoclave discharge composites were in the range of 1 kg/t and 80-90 kg/t of cyanidation feed, respectively. Arsenic precipitation efficiency was evaluated based on arsenic distribution between the solid and solution phases. The distribution of arsenic to the solids increased from 20 to 40% in the feed to 60-80% in the discharge. The profile of arsenic precipitation across the autoclave suggested that arsenic precipitated as an unstable compound in the first compartment, followed by redissolution in the second compartment and re-precipitation as a stable compound in the last two compartments. The re-dissolution of the precipitate in the second compartment appeared to coincide with an increase in free acid formation through sulphur oxidation. 383
Bioleaching Applications 100
% Suphide Oxidation % Au Extraction
90 80 70 60 50 40
Au Extraction
30
S= Oxidation
20 10 0 0
20
40
60
80
Autoclave Nominal Retention Time,min
Figure 5. S= Oxidation and gold extraction as a function of autoclave retention time 100 As Distribution to Solids, %
90 80 70 60 50 40 30 20 10 0 0
20 40 60 Autoclave Nominal Retention Time, min
80
Figure 6. As Distribution to solids as a function of autoclave retention time In simple arsenic-acid systems, it is well known that the solubility of arsenic increases with increasing acidity, with a minimum solubility in the pH range 3 to 5. The high acid level in the autoclave discharge, 40-60 g/L, must be one of the reasons for the high residual arsenic in solution. The results obtained during the autoclave pilot plant campaigns indicated that the residual arsenic and iron in the autoclave discharge depended to a certain extent on the initial arsenic and iron concentration in the POX feed/BIOX® discharge, with relatively high soluble iron and arsenic in the feed to POX (campaign 1) leading to relatively high arsenic in solution in the POX discharge. Arsenic precipitation in the autoclave was only ~50% in campaign 1, versus ~85% in campaigns 2 and 3. The iron to arsenic ratio is probably also important, since work by Monhemius and Swash1 has shown that the precipitation of scorodite or other ferric-arsenate components is inhibited at higher Fe/As ratios, in the presence of sulphuric acid. Mineralogical analyses of selected autoclave discharge samples suggested that the major phase present was a basic iron sulphate and anhydrite. Arsenic was present as a low-level constituent of the iron sulphate. 384
Bioleaching Applications
3.3 Solid/liquid separation and liquor treatment The autoclave discharge slurry and the small bleed of the BIOX® liquor that bypassed the autoclave were combined and treated in a multi-stage continuous CCD/neutralization circuit. The main objectives of this stage of the pilot program were to: demonstrate, at a pilot plant scale, that As(III) present in solution can be effectively oxidized to As(V) using SO2 and air as an oxidant; precipitate arsenic as a stable product suitable for disposal; and produce final effluent that meets industrial regulatory limits. As(III) oxidation was carried out in a series of three cascading 45L tanks, with additions of sodium metabisulphite (Na2S2O5) and air into each oxidation tank. Retention time within the oxidation circuit was 4.2 hours. The results indicated that the autoclave discharge could be thickened to 55-60% solids. The SO2/air oxidation was an effective process for As(III) oxidation. On average, the oxidation feed As(III) concentration was decreased from 125 mg/L to below 20 mg/L. Average consumption of SO2 during the oxidation stage was 2.5 g/L of CCD overflow solution. The SO2 consumption was not optimized during the pilot plant campaign. Overall alkali consumption, during the neutralization stage, including limestone and lime as CaO equivalent, was in the range 500-600 kg/t of autoclave discharge solids. The final neutralized solution met industrial effluent standards for the required elements. In order to determine the stability of the final waste products, samples of the neutralized sludge, the CCD underflow and the CIL residue and were submitted for TCLP 1311 leachate testing. The results showed that the concentrations of all the elements in the leachates for samples tested were below regulatory limits. However, elevated concentrations of manganese (Mn) and zinc (Zn) were observed in the leachates of the neutralization sludge samples. 4.
CONCLUSIONS From intensive and integrated pilot plant campaigns, it was confirmed that: • Acceptable sulphide oxidation could be achieved by two-stage oxidation with BIOX® (4 days) followed by POX (40 minutes at 220°C). • Gold extractions from the POX discharge were excellent, ranging from 96-98%, with only 1 kg/t NaCN consumed. • Arsenic precipitation in the autoclave varied between 50-80%. • Neutralization of the plant liquors (BIOX® and POX) was successful, resulting in effluents meeting effluent standards. • The results of TCLP 1311 testing of the final waste products yielded leachates that were below regulatory limits, confirming that arsenic had been stabilized. The Olympias process could be simplified, and capital and operating costs significantly lowered, if all of the BIOX® liquor was allowed to bypass the autoclave directly to the neutralization circuit. In this regard, the following factors are pertinent: • The precipitate produced during neutralization under atmospheric conditions is as stable as the ferric arsenate compound produced in the autoclave, as indicated by the most recent research findings (Monhemius and Swash). • The total residence time in the autoclave was determined by the rate of ferric arsenate precipitation (50-70 minutes) rather than the rate of sulphide and sulphur oxidation (30-40 minutes). The autoclave size could therefore be significantly reduced, if the autoclave design was based solely on the requirement for efficient gold recovery. 385
Bioleaching Applications
• Bypassing the BIOX® liquor around the autoclave will reduce the iron, arsenic and sulphate concentration in the feed to POX, which will in turn increase the efficiency of precipitation of any remaining arsenic in the autoclave. • It will also allow the feed to the autoclave to be adjusted to the required density and sulphide plus sulphur concentrations for auto-thermal autoclave operation. Therefore, it will not be necessary to blend untreated pyrite concentrate with BIOX® discharge in the POX feed to achieve the required heat balance. This will further reduce the size of the autoclave and lower capital costs.
ACKNOWLEDGEMENTS The authors wish to acknowledge the work of all investigators and pilot plant operators who have been involved in the technical evaluation of the Olympias project. The permission of TVX Hellas in publishing this paper is gratefully acknowledged. REFERENCES 1. A.J. Monhemius and P.M. Swash, The Removal and Stabilisation of Arsenic from Copper Refining Circuits by Hydrothermal Processing, J. of Minerals, Metals and Materials Society, 1999, Vol. 51, No. 9, pg 30. 2. J.A. van Niekerk, Continuous BIOX® Pilot Plant Treatment of Pyrite Concentrate from the Olympias Mine at Lakefield Research Canada, Internal and Confidential Report, No. PR 00/006, Dec. 14, 2000. 3. K.G. Thomas, Research Engineering Design and Operation of a Pressure Hydrometallurgy Facility for Gold Extraction, Ph. D. Thesis, Technische Universiteit Delft, 1994. 4. D.C. Van Aswegen, Commissioning and Operation of Bio-Oxidation Plants for the Treatment of Refractory Gold Ores, Hydrometallurgy Fundamentals, Technology and Innovations, J.B. Hiskey and G.W. Warren, eds., Society for Mining, Metallurgy and Exploration, Inc. of the AIME, Littleton, Colorado, USA, pg 709-725, 1993.
386
15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
The development of the first commercial GEOCOAT® heap leach for refractory gold at the Agnes mine, Barberton South Africa Todd J. Harvey and Murray Bath GeoBiotics, LLC, 12211 W. Alameda Pkwy, Suite 101, Lakewood, CO, 80228, USA
Abstract Facing lower margins, gold mine operators throughout the world are seeking ways to reduce costs for treating refractory sulphide ores. Fortunately, oxidation technologies for refractory ores need not be complex or costly. GeoBiotics, LLC. has developed the GEOCOAT® heap biooxidation technology for the treatment of refractory gold ores and concentrates. GEOCOAT® uses simple low cost unit operations common to heap leaching of gold and copper ores. The cost effectiveness and flexibility of GEOCOAT® heap biooxidation allows mining companies to reduce cut-off grades and increase mine life. In addition, GEOCOAT® can be utilized to supplement or replace existing expensive or environmentally unfriendly oxidation processes such as roasting, autoclaving, or agitated tank biooxidation. The GEOCOAT® technology involves coating sulphide concentrates onto sized support rock, stacking in a heap environment, irrigating with acidic nutrient solutions and supplying low-pressure air to the heap base. After biooxidation the concentrate is stripped from the support rock, neutralized and leached in conventional cyanidation facilities. African Pioneer Mining (APM) has adopted the GEOCOAT® technology for use at the redeveloped Agnes Mine in Barberton, South Africa. This paper details the development of the GEOCOAT® process including laboratory testing and construction and commissioning of the first commercial GEOCOAT® plant. The Agnes Mine is expected to initially produce approximately 25,000 ounces of gold per year with the potential for expansion.
Keywords: biooxidation, refractory ore, sulphide, pyrite, GEOCOAT®
1.
INTRODUCTION
The GEOCOAT® process incorporates elements of two successful and commercially proven technologies: heap leaching and biooxidation. Gold-bearing sulphide minerals are concentrated by flotation and thickened. The resulting slurry is thinly coated onto crushed, screened support rock, stacked on a lined pad, and allowed to biooxidize. Coating is accomplished by spraying the concentrate slurry onto the support rock as it discharges from the end of a stacking conveyor onto the biooxidation heap as shown in Figure 1. The
387
Bioleaching Applications
coating solids density is highly dependent on the slurry viscosity and densities of 50-65% have been successfully coated at scale. The hydrophobic nature of the concentrate assists in the formation of a coating on the support rock. No binding agents are required. The concentrate naturally adheres to the support rock and does not wash out of the heap during solution application or during heavy rainstorms.
Figure 1. Coating Operation The support rock is uniformly sized, in the range of 6 to 25 millimeters in diameter and the concentrate coating is relatively thin, less than one millimeter in thickness. The weight ratio of support rock to concentrate is in the range of 5:1 to 10:1. Figure 2 illustrates the concentrate coated support rock from a pilot test.
Figure 2. Concentrate Coated Support Rock Depending on the desired temperature of operation, the heap is inoculated with naturally occurring sulphide-oxidizing bacteria. Nutrients are added to the heap via recirculating solutions. As biooxidation progresses, the sulphides in the concentrate are oxidized and the solubilized iron, arsenic and sulphate are carried from the heap by the recirculating solution. A portion of the solution stream is bled from the circuit for neutralization to maintain a maximum iron or arsenic level. The relatively uniform size of the support rock leads to large interstitial spaces within the heap and subsequently a low resistance to air and liquid flows. Sufficient air for biooxidation and heat removal is supplied to the heap by low-pressure blowers through a system of perforated pipes laid in the drain rock below the base of the heap.
388
Bioleaching Applications
After biooxidation the coated rock is unloaded from the pad and the oxidized concentrate removed by trommeling or wet screening. The concentrate residue is then neutralized and subjected to conventional gold recovery methods. The support can be recycled or, in the case of low-grade sulphide ore, a portion can be bled out for gold recovery and replaced with fresh gold bearing ore. Figure 3 presents a simplified schematic representation of the process(1,2). Gold Sulfide Ore
Concentrate Storage
Air
Initial Inoculum Generation
Tailings
Flotation
Thickener
Concentrate
Concentrate
Crushed Support Rock Barren or Low Grade
Spray Coating
-25 +6 mm
Biooxidized Concentrate
GEOCOAT ® Heap Crushed Support Stockpile
Air
Radial Stacker
Suport Recycle
Blower PLS Pond
Recycle Solution
Trommel Thickener
Recirculating Solution
PLS Bleed
CIL Lime/Limestone
Tailings
Figure 3. GEOCOAT® flowsheet
Fe/As Removal
Tailings
Gold Recovery
Neutralization
The Agnes mine, located in Barberton, Mpumalanga, South Africa, decided to employ the GEOCOAT® technology as it presented a lower cost alternative refractory gold treatment scenario. Additionally, the process is simple and easy to control. The mine, owned by African Pioneer Mining, is scheduled to produce approximately 25,000 ounces of gold per year from an ore throughput of 500 tonnes per day from the Galaxy deposit. The GEOCOAT® plant is designed to treat approximately 50 tones per day of sulphide concentrates containing 50 g/t Au, 15% Fe, 1% As and 15% sulphide sulphur. The GEOCOAT® plant is operating at full capacity with respect to concentrate delivery and the first biooxidized concentrates are expected to be treated in the CIL circuit by the end of May 2003. The plant was commissioned using temporary materials handling equipment but the balance of the plant including stacking and reclaim conveyors, trommel, CIL, neutralization and cyanide destruction are scheduled to be commissioned at the same time as the first concentrates are completing their biooxidation cycle. This paper details the testwork employed to develop the plant and the ongoing construction and commissioning.
2.
LABORATORY TEST PROGRAM A series of biooxidation tests have been performed at Lakefield Research Africa to define the performance of the GEOCOAT® process on the Galaxy and Princeton concentrates. These tests include amenability testing and column tests. The amenability tests are conducted as batch stirred tank biooxidation and the column tests are conducted in 150mm diameter by 6m high columns. Several process variables have been investigated 389
Bioleaching Applications
including coating ratio, heap height, temperature regime, and grind. Additionally, mineralogical investigations were undertaken before any testwork was initiated. The amenability tests are conducted at approximately 5% solids w/v using a heated and aerated stirred tank reactor. Periodic solution and solids samples are removed to determine the extent of biooxidation and the gold extraction. Both amenability and column tests employ adapted bacteria that is maintained using the concentrate under investigation. Approximately 2 months of adaptation is required prior to the commencement of any biooxidation testing. The columns tests are conducted by batch coating the concentrate onto representative substrate at a known ratio and loading into the column. The columns are equipped with zone heating to ensure uniform temperatures. Low pressure humidified air is applied to the base of the column at rates in excess of stoichiometric. Acid solutions are applied to the top of the column via a peristaltic pump, the effluent solutions are collected separately at the column base. Effluent solutions are generally recycled. However, solution is removed on a periodic basis to maintain the desired PLS profile. The column is fitted with sampling ports that allow for both solid and liquid sample removal. At the termination of the biooxidation cycle the column is acid/water rinsed, allowed to drain and then emptied. The concentrate residue is removed from the support rock by simple wet screening. The biooxidized concentrate is then subjected to CIL testing to determine gold extraction and reagent consumptions.
2.1 Mineralogy The Galaxy and Princeton deposits are composed of a very similar mineral makeup. The primary gold carrier is pyrite (FeS2) with minor amounts of arsenopyrite (FeAsS). The primary gangue material is quartz (SiO2) with appreciable quantities of siderite (FeCO3) and ankerite (CaCO3*(Mg, Fe, Mn)CO3). Both ores are refractory due to the small gold grains encapsulated within the sulphide matrix. The majority of the gold is less than 5um. The Princeton ore is a higher sulphur grade and is more refractory, providing a lower baseline gold extraction with direct cyanidation. Table 1 shows the average composition of the Princeton and Galaxy flotation concentrates. As shown from Table 1, the Princeton concentrate has a much higher sulphide content but at a lower gold grade. The Agnes orebody is also comprised of a variety of other reefs but most are similar to either the Galaxy or Princeton veins. Table 2 shows the complete testwork program for the Agnes orebody. Table 1. Princeton and Galaxy Concentrate Mineralogy Mineral Au FeS2 FeAsS S= CO32-
390
Composition Princeton Galaxy 20.1 g/t 50.0 g/t 75.5% 28.2% 1.8% 1.4% 40.3% 15.1% 1.1% 7.4%
Bioleaching Applications
Table 2. Testwork Summary Concentrate
Test
1
2
3 4
5
6
7
8
9
10
11
12
Concentrate chemical analysis
•
•
•
•
•
•
•
•
•
•
•
•
Concentrate particle size analysis
•
•
•
•
•
•
•
•
•
•
•
•
Cyanidation of un-oxidised concentrate
•
•
•
•
•
•
•
•
•
•
•
•
Batch biooxidation amenability test (BAT)
•
•
•
•
•
•
•
•
•
•
•
•
Cyanidation of BAT oxidised residue
•
•
•
•
•
•
•
•
•
•
•
•
•
•
•
2m GEOCOAT® column test
•
®
•
6m GEOCOAT column test Cyanidation of oxidised column residue
•
•
•
•
Cyanide optimisation of oxidised column residue
•
•
•
•
Thickening of column oxidised residue
•
•
Column solution neutralisation
•
•
TCLP stability test on neutralisation precipitate
•
•
Key:
1 - Svengali 1 2 - Svengali 2 3 - Giles 1
4 - Giles 2 5 - Woodbine 1 6 - Woodbine 2
7 - Galaxy 1 8 - Galaxy 2 9 - Princeton 1
• •
•
10 - Princeton 2 11 - Princeton 3 12 - Princeton 4
2.2 Batch amenability biooxidation testing Batch biooxidation amenability tests were carried out at Lakefield on nine samples of concentrates from five different ore zones. Initial samples (Svengali 1, Woodbine 1, Giles 1, and Princeton 1) were relatively low-grade, a result of very high mass yields in the flotation tests. Subsequently, the batch amenability tests were repeated on higher grade concentrates (Svengali 2, Woodbine 2 and Giles 2). Only one concentrate was produced from the Galaxy ore for amenability. The results of this work are summarized in Table 3. As shown by Table 3, all concentrates were extremely amenable to biooxidation producing sulphide oxidations ranging from 74.3% to 100% with gold extractions ranging from 92.5% to 96.1%. Figure 4 shows these results graphically. Figure 4 shows that the concentrates exhibit a wide-ranging level of refractoriness as shown by the baseline CIL extractions. Gold extractions range from 47.9% to 76.0% for the unoxidized concentrates. It is also apparent that in order to achieve high overall extractions, near complete oxidation is required. The reagent consumption of these BAT tests was extremely dependant on the concentrate employed and the subsequent elemental sulphur formed during the biooxidation process. Generally, the biooxidation process becomes more efficient at elemental sulphur removal as the bacterial adaptation period is lengthened(3).
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Table 3. Batch Amenability Tests Conc
Svengali 1
Biooxidation Conc. Head time Grade Au days* g/t
Sulphur Oxidation %
Gold Extraction %
0 44.0 0.0 58.6 11 93.8 95.4 19 95.3 92.5 47.9 74.6 0.0 Svengali 2 0 59.2 0.1 4 68.5 24.5 6 79.2 47.3 10 92.9 74.3 15 Woodbine 1 0 14.7 0.0 56.0 11 94.2 90.2 19 96.6 92.8 54.2 44.0 0.0 Woodbine 2 0 63.9 0.0 4 76.2 37.6 6 84.1 57.1 10 92.8 87.5 15 Giles 1 0 9.9 0.0 71.6 11 94.4 89.9 19 98.3 94.2 Giles 2 0 41.0 0.0 73.8 3 8.4 77.6 6 70.1 83.1 10 77.5 90.2 15 99.5 90.1 Princeton 1 0 34.0 0.0 54.5 10 79.8 85.7 19 93.4 92.7 Princeton 2 0 27.7 0.0 64.9 5 30.0 78.7 10 90.5 88.9 15 97.0 93.1 20 99.5 90.6 76.0 50.6 0.0 Galaxy 1 0 93.5 82.1 5 95.5 98.6 8 95.7 99.3 10 96.1 100.0 12 * 0 bio-oxidation time represents the concentrate sample as-received ** reagent consumptions - kg/t unoxidized concentrate
392
Reagent Consumption** NaCN CaO kg/t kg/t 4.3 1.1 9.0 2.6 13.2 3.4 1.0 6.0 6.4 18.2 9.6 21.0 15.3 34.8 17.0 42.8 4.5 0.8 9.8 2.8 11.4 3.3 1.0 6.0 6.7 21.2 10.1 21.8 17.1 32.2 14.7 36.5 4.0 0.9 10.7 3.5 10.1 2.8 6.3 1.0 18.2 8.3 19.0 10.1 27.4 11.8 25.2 14.4 8.6 0.7 41.7 37.1 43.9 23.3 6.8 3.9 28.9 26.9 23.2 19.1 39.2 20.0 41.4 21.7 0.5 9.7 8.9 21.9 8.7 22.3 7.6 18.1 7.0 17.0
Bioleaching Applications
100
Gold Extraction (%)
90
80 Svengali 1
70
Svengali 2 Woodbine 1 Woodbine 2
60
Giles 1 Giles 2
50
Princeton 1 Princeton 2 Galaxy 1
40 0
10
20
30
40
50
60
70
80
90
100
Sulphide Oxidation (%)
Figure 4. Batch Amenability Test Results 2.3 Column Biooxidation Testing The batch amenability tests were followed by a series of GEOCOAT® column tests. Column tests are designed to simulate conditions in a GEOCOAT® heap. Support rock type, coating ratio, solution application rate, and solution management scheme are all selected to duplicate the operation of the full-scale heap. Column diameter is 150mm and the height is typically 2m. However, a final test was carried out in a 6m tall column, designed to simulate fully a GEOCOAT® heap stacked to the design height of 6m. This test will provided additional information on solution chemistry and confirmed the solution management strategy proposed for the Agnes project. The concentrates for each of the first three column tests were produced from Princeton ore, while a fourth column test was carried out on Galaxy concentrate produced from the Agnes plant. Separate concentrates were produced for each of the column tests. The first test used a barren support rock, while the other tests were run using the Alpine waste rock that it is proposed to use in the full-scale heap. Intermediate solids samples were extracted from each of the columns twice during the course of the tests. At the termination of the test, the columns were emptied and the contents wet screened to separate the support rock from the oxidised concentrate. As with the intermediate samples, the oxidised concentrate slurry was filtered, washed and dried, and the solids analysed and subjected to cyanidation testing. Table 4 shows the results of the biooxidation column testing. As shown, the column tests produced similar gold extraction results to those of the BAT tests with lower cyanide consumptions. The use of a 6m column had no impact on the biooxidation process and excellent results were achieved.
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Bioleaching Applications
Table 4. GEOCOAT® Column Test Results Cyanidation Conditions Au Ext Column % No. NaCN Other** ppm Prin 1 0 0 2000 47.5 (Conc. 2) 20 23.1 2000 74.7 46 53.0 2000 85.6 80 (final) 77.8 500 81.2 80 (final) 77.8 2000 92.4 53.1 2000 0 Prin 2 0 93.9 2000 68.4 (Conc. 3) 21 89.0 2000 81.0 35 82.4 No carbon 2000 91.5 53 (final) 88.8 2000 91.5 53 (final) Prin 3 0 0 2000 66.8 (Conc 4) 24 39.1 2000 76.7 36 68.8 2000 81.8 50 86.7 2000 86.6 74 95.1 500 55.6 74 95.1 1000 85.5 74 95.1 1500 87.5 74 95.1 2000 92.5 Galaxy 2 0 0 2000 77.6 (6m) 24 63.9 2000 91.9 38 89.6 2000 89.4 59 93.0 500 90.1 59 93.0 1000 97.0 59 93.0 2000 96. 7 * based on bioleach concentrate feed, ** 24hr CIL tests if not stated Bioox. Time days
3.
Sulphur Ox %
Reagent Consumption* NaCN CaO kg/t kg/t 6.8 3.9 16.7 6.9 14.2 3.7 6.7 4.4 24.3 4.2 1.3 7.0 6.8 10.7 7.4 14.3 4.1 11.7 4.1 11.5 7.0 1.3 24.8 8.6 18.0 6.5 21.3 7.7 6.7 5.9 12. 8 5.9 18.4 5.7 23.6 5.9 3.7 3.2 21.9 8.6 15.5 6.3 7.4 8.4 13.1 7.9 16.7 7.4
GEOCOAT® PLANT CONSTRUCTION
The Agnes mine began construction of the GEOCOAT® plant in October of 2002 and the first concentrate was stacked on the pad at the beginning of February 2003. This despite construction being delayed for a month by the Christmas season. It is expected that the recovery of the first biooxidized concentrate will commence by the middle of May 2003 coinciding with plant completion. The biooxidation cycle is expected to require approximately 60 days to complete. The first heap will require a slightly longer period as the support and base rock need to be pH stabilized due to their carbonate content. Additionally, the speed of the process will improve as a large population of adapted mesophilic bacteria is formed. Currently bacteria is grown on site for heap inoculation in 3-10m3 fermentors. The heap has been biooxidizing at an average rate of 1.7% sulphide oxidation per day, above the design rate of 1.5% and the heap is maintaining its desired operating temperature between 35-45°C. The Agnes GEOCOAT® plant was designed and built to provide African Pioneer Mining with flexibility for future expansion. The pad base was enlarged as was the air supply to facilitate the treatment of other concentrates such a those to be derived from the Princeton deposit. The Table 5. provides the design statistics of the Agnes plant.
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Bioleaching Applications
Table 5. GEOCOAT® Plant Statistics Design (for Galaxy only)
Actual/Proven
Stacking Rate
34.5 tph
100 tph
Concentrate Rate
4.6 tph
8.0 tph
Biooxidation Time
60 days
N/A
10-30 L/m2/hr (max 80m3/hr) Using Wobblers®
10-30 L/m2/hr (max 120m3/hr) Using Wobblers®
Centrifugal Fan 360 m3/min at 2.5 kPa
Centrifugal Fan 3 x 360 m3/min at 2.5 kPa
Irrigation Rate Aeration Aeration Method Pad Size Heap Size Pond Size Stacking Method Concentrate Recovery
Perforated pipes in drain rock base 50 x 75 m
50 x 150m
6 x 45 x 60m
6 x 45 x 130m
3
1 x 7000 m3
1 x 2500 m
Slewing radial stacker with automated material handling FEL, Trommel, Thickener
Gold Recovery
24 hr CIL - 6 x 11 m3 Tanks
24 hr CIL - 6 x 20 m3 Tanks
Effluent Disposal
Heap bleed solutions are neutralized by mixing with carbonate float tails. CIL cyanide is destroyed using excess acid bleed.
Performance Monitoring
Solution analysis, solid sampling, and temperature monitoring.
The following pictures document the construction of the plant.
Figure 5. Site – Sept 2002
Figure 6. Site showing pad and pond – Oct 2002
395
Bioleaching Applications
Figure 7. Liner System and Protective Sand Layer – Oct 2002
Figure 8. HDPE Aeration Pipes – Jan 2003
Figure 9. Stacking Heap – Feb 2003
Figure 10. Heap – April 2003
Figure 11. Steam Rising from Heap – April 2003 4.
CONCLUSIONS
The first commercial GEOCOAT® plant for the treatment of refractory gold sulphides is currently being commissioned in Barberton, Mpumalanga, South Africa at the Agnes Gold Mine. African Pioneer Mining has selected this technology based on its cost 396
Bioleaching Applications
advantages, its inherent simplicity and safety. Laboratory testwork has shown that the process should yield sulphide biooxidations well above 90% in under 60 days with gold extractions also above 90%. The commercial heap has yet to complete its first biooxidation cycle but indications are that it will yield the desired results. The heap is now operating within the desired temperature range and solution and solids assays indicate an average sulphide biooxidation rate of 1.7% per day, which is in excess of design.
ACKNOWLEDGEMENTS The authors would like to thank African Pioneer Mining for the permission to write this paper and also all of those people who assisted in making this project a success. REFERENCES 1. Harvey, T.J. et al, 1998. Heap Biooxidation For The Treatment Of Low Grade Refractory Ores At Ashanti Goldfields' Obuasi Operations. Randol Gold And Silver Forum'98, Denver, Colorado. 2. Kohr, W.J., 1998. U.S. Method Of Biotreatment For Solid Materials In A NonStirred Surface Reactor. GeoBiotics, Inc., US, US Patent Number 5,766,930. 3. Nicholson, H., 1993. Selection of a refractory gold treatment process for the Sansu Project. BioMine'93. Adelaide, South Australia, Australian Mineral Foundation, p. 20.1-20.11.
397
15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
The electrochemistry of chalcopyrite bioleaching using bacteria modified powder micro-electrode* Li Hongxua, Wang Dianzuoa, Hu Yuehuab, Ruan Renmana a
b
General Research Institute of Nonferrous Metals, Beijing 100088, China College of Resources Processing and Biological Engineering, Central South University, Hunan Changsha, 410083, China
Abstract The anode behavior of chalcopyrite in the presence and absence of bacteria in 9K mediums was studied using a combination of standard electrochemical techniques by bacteria modified powder microelectrode at 30°C. It was found that the anode oxidation process of chalcopyrite includes many intermediate transient reactions. During the dissolution, it is exist the intermediate production of chalcocite and covellite. At lower scanning potential, the iron of chalcopyrite is extracted by ferrous form, but at the relative high potential, it is the ferric. When in the presence of Thiobacillus ferrooxidans, the peak current and reversibility of oxidation reaction increased, and the oxidation potential negatively moved. It demonstrated that apart from enhancing the metallic ion extracting oxidation reaction, the Thiobacillus ferrooxidans has also contribution to the oxidation of element sulfur formed on the surface of chalcopyrite during the intermediate process. The effect of ferric and pH on the oxidation of chalcopyrite was also investigated. Keywords: chalcopyrite, Thiobacillus ferrooxidans, electrochemistry, bioleaching, oxidation, bacteria modified, powder microelectrode, mechanism
1.
INTRODUCTION The bioleaching research has been a great success for metallurgy industry. As a result of this work, a significant number of commercial applications have emerged and are able compete with conventional processing, especially the application for the copper recovery. Further more, bioleaching treatments have the great advantage of being environment friendly [1, 2]. However Bioleaching applications for copper extraction are mainly concentrated on the treatment of secondary copper minerals. chalcopyrite is the most abound ore of the sulfide minerals of copper [3], but it gives very slow kinetics and limited recovery, so it need to elucidate the oxidation mechanism. We know that the microorganism can catalyze the copper dissolved. It has demonstrated that chalcopyrite and most other metallic sulfide are dissolved by the electrochemistry mechanism, and the most of sulfides are semiconductors, using the electrochemistry methods to study the oxidation mechanism of sulfides is affective [4-10]. * supported by national science foundation of China [50204001]
399
Bioleaching Applications
Furthermore, there are large number of studies on the anode dissolved mechanism of chalcopyrite in different media including culture media or other experimental condition [11, 12], however there are fewer information on the experiments when microorganism in presence. One of the most important reasons is it is difficult to guarantee the affective attachment of microorganism on the chalcopyrite surface when the electrochemical quick scan carrying on, and the convention polished nature massive specimen used as work electrode and the microorganism added in solution media [13]. For this reason, we used a new method of the bacteria modified powder microelectrode as the work electrode to over the above said difficulties. Based on the characters of the microelectrodes, apart from the affective attachment of leaching bacteria on the surface of chalcopyrite powder, the more information about transient intermediate reaction and other useful information during electrochemistry scanning can be obtained [14]. Although it is reported that some microorganisms like Sulfulobus are more effective on the chalcopyrite resolution, but in ordinary temperature, thiobacillus ferrooxidans is still the main microorganism for bioleaching [15], so in this work, we use thiobacillus ferrooxidans as bioleaching microorganism to do some preliminary investigations.
2.
MATERIALS AND METHODS
2.1 Ore The nature chalcopyrite was get from Hunan museum, with high quality and showing no foreign inclusions under the microscope. The massive specimen was crushed and milled under N2 ambience in order to prevent the surface oxidation. The size of particle was under 50um. 2.2 Bacterial culture The mixed cultures of acidophilic bacteria were obtained from Guangdong Dabaoshan copper mining. The thiobacillus ferrooxidans was isolated from mixed cultured in laboratory, The standard composition of the nutritive media was 9K media, g/L: (NH)4SO4, 3.0; KCl, 0.1; K2HPO4, 0.5; MgSO4.7H2O, 0.5; Ca(NO3)2, 0.01. The water used in the experiment was ferrous ion free. 2.3 Electrode Prepared an micro plate electrode using platinum wire, and the diameter is 100 um, corroded the micro electrode surface to a small 100um deep pitch; washed the chalcopyrite powder with acetone prior to use, added the bacteria to chalcopyrite powder, blended evenly and keep for few hours in order to guarantee the bacteria attached and adhered to the powder surface absolutely, so the bacteria modified sulfide powder was prepared, and compressed it into the electrode pitch using a glass plate to make the powder electrode surface flawless. The detail method has described in other papers [14, 16]. The powder microelectrode is composed of a micro plate electrode and a thin layer electrode. The structure of the electrode is shown in Fig. 1.
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Where (b) is the tip of (a); (a) is platinum wire; (b) is glass jacket; (c) is chalcopyrite powder
2.4 Electrochemical experiment The electrochemical measurements were performed in a typical cell (500mL) with three electrodes: the working electrode (bacteria modified chalcopyrite powder microelectrode), the counter electrode (Pt plate), and the reference electrode (KClsaturated calomel electrode). The cell was keep at constant temperature by connecting it to a circulating thermostatically controlled water loop. The electrolyte used in the experiment was standard 9K media without ferrous ion. The used water was ion removed. The electrochemical experiments were carried out using Solartron 1287. In the paper, all the potentials value is Vs SCE. 3.
RESULTS AND DISCUSSION
3.1 Influence of bacteria Fig. 1 and Fig. 2 are the cyclic voltammograns of the chalcopyrite powder microelectrode when in absence and presence of bacteria. The initial sweep potential is the rest potential. Comparing different sweep cyclic we can see that the oxidation of chalcopyrite include multi intermediate steps. From Fig.1 and Fig.2 we know that from initial potential the shape of the anode prewave of first cyclic is different from the second and third cyclic, comparing the anode direct oxidation reaction and cathode direct reduction reaction, it demonstrates that during the oxidation there is a thin layer product covered on the chalcopyrite surface, and the anode process is controlled by diffuse steps. This is similar to the results described by Biegler et al using massive chalcopyrite electrode in acid medium. [17-20]. It is predicted that in the anode oxidation process, there are element sulfur and other intermediate phase will be created on the chalcopyrite surface. In Fig. 1, A represents the "prewave", in which chalcopyrite is transformed to CuS, through an intermediated non-stochiometric phase (Cu1-xFe1-yS2-z), producing So and Fe2+, the oxidation of chalcopyrite through the following reaction: CuFeS2 → CuS + Fe 2+ + So + 2e
(1) Eo=0.055 In the potential zone from –0.2V to 0.5V (the initial potential of peak A has a small movement to negative during second and third cyclic sweep in anode direction. It is possible to conclude that because of the deform of the crystal lattice of chalcopyrite by 401
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polarized and the intermediate production formed by cathode reduction reaction, the oxidation reaction of Cu2S may occur by: Cu 2 S → 2CuS 2+
2+
CuFeS2 → Cu + Fe + 2S + 4e o
Eo=-0.021
(2)
Eo=0.231
(3)
Figure 1. The cyclic voltammograns of the chalcopyrite in absence of bacteria (ro=5×105 m, 5 mVs-1, T=25°C, pH=2, initial sweep direction: anode) 1-first cyclic, 2-second cyclic, 3-third cyclic
Figure 2. The cyclic voltammograns of the chalcopyrite in presence of bacteria (ro=5×10-5 m, 5 mVs-1, T=25°C, pH=2, initial sweep direction: anode) 1-first cyclic, 2-second cyclic, 3-third cyclic
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At the potentials more positive than 0.7V, the Fe2+ and So would be oxidized to Fe3+ and SO42- for further steps in the C range. The overall dissolution of chalcopyrite takes place through the following reactions: CuFeS2 → Cu 2+ + Fe3+ + 2So + 5e CuFeS 2 + 8H 2 O → Cu
2+
+ Fe
3+
Eo=0.274 2-
(4)
+
+ 2SO 4 + 16H + 17e
(5) It is consistent to the Macillan et al’s results using massive electrode, but the peak of the oxidation of intermediate process is more clearly. Potential range from about 0.3V to 0.7V (peak B) represents the chalcopyrite oxidation reaction: In the reverse scan, the range of D and E may represent the reduction of Fe3+ and Cu2+ produced in anode reaction respectively, the peak F may represents the reverse reaction of (3), by the express of Biegler et al [20-23]. We know that during the cathode scanning there would be forming a reduction layer on the chalcopyrite particles surface. At the peak G and H, there would occur a series of reaction as following: 2CuS + 2H + + 2e → Cu 2 S + H 2 S
(6)
+
S + 2H + 2e → H 2S o
(7) The peak may represent the oxidation of element sulfur formed on the surface of chalcopyrite during the anode and cathode process. In this range the reverse reaction of (7) would occur. Comparing Fig. 1 and Fig. 2, we know that thiobacillus ferrooxidans can enhance the oxidation of chalcopyrite, because in Fig. 2, the shapes of the oxidation reactions represented by prewaves of A, B and C are more apparently, especially the peak B and C. The anode sweep results of currents and potentials of the reaction can be shown in Table 1. From the results we know that when in the presence of thiobacillus ferrooxidans, the peak current density and reversibility of the oxidation reaction increased, and the oxidation potential negatively moved. As described above, the reaction of (4), and the conversion of ferrous to ferric and the element sulfur to sulfate on the ore surface have elevated. In cathode sweep the peak D in Fig.2 is more clearly than in Fig. 1, it demonstrate that there are more mount of ferric have been produced on the surface layer during anode process when the thiobacillus ferrooxidans in presence. The peak I in Fig. 2 is more flat than that is in Fig1, it is show that there are less amount of element sulfur formed on the surface when in the presence of thiobacillus Ferrooxidans during the anode oxidation and cathode reduction process. It demonstrated that the attachment of Thiobacillus ferrooxidans on the surface can accelerate the oxidation reaction of chalcopyrite, especially the oxidation of ferrous, apart from enhancing the metallic ion extracting, it also has contribution to the oxidation of element sulfur formed on the surface.
Table 1. The anode sweep results of currents and potentials by cyclic voltammogran (first cyclic)
Peak currents density (e6A/cm2) Initial potentials of prewave (V)
Peak ranges
A
B
C
I
Reactions Inoculated Uninoculated Inoculated Uninoculated
(1) 6.59 3.17 -0.23 -0.20
(3) 22.33 7.66 0.26 0.32
(4) 18.70 6.27 0.68 0.75
(7) 1.51 2.76 -0.68 -0.53 403
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3.2 Influence of pH Fig. 2 is the cyclic voltammograns of the chalcopyrite in the presence of bacteria under different acid condition.
Figure 2. The cyclic voltammograns of the chalcopyrite in the presence of bacteria (ro=5×10-5 m, 5 mVs-1, T=25°C, pH=2, initial sweep direction: anode) 1—pH=2 2—pH=1.5 From the results we know that when the pH decrease from 2 to 1.5, the anode oxidation potentials of chalcopyrite (especially prewave B and C) become more electropositive, the current density decrease, and the anode oxidation reaction is inhabited slightly. This result is similar to the description by C. Comze et al [11, 20, 21]. At the cathode sweep, the shifts of at least 0.15V in the initial potential of F and G reactions towards the more electropositive zone because of the decrease of pH from 2.0 to 1.5 have been observed. It is means that there is a clear influence of pH to the cathode scan by the following reactions; 2CuS + 2H + + 2e → Cu 2 S + H 2 S So + 2H + + 2e → H 2S In which we know the increase of pH would enhance the reactions towards right hand direction, and the E-pH dependence can be expressed as: E=E′-0.059·pH.
3.2 Influence of Fe3+ Fig. 3 is the cyclic voltammograns of the chalcopyrite when the Fe3+ free and added in the electrolyte. At the anode scan, the current density of prwave of A, B, C and I increased and the initial potentials of prewave negatively moved. The addition of Fe3+ makes the reaction rates of the anode oxidation of chalocopyrite increased greatly, and we know that the Fe3+ can oxidize the intermediate CuS or Cu2S film formed on the chalcopyrite. At the same time, from the information given by prewave of I, we know that by the increase of oxidation rates there are more element sulfur formed on the chalcopyrite surface, although the diminish effect by thiobacillus ferrooxidation.
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Figure 2. The cyclic voltammograns of the chalcopyrite in presence of bacteria (ro=5×10-5 m, 5 mVs-1, T=25°C, pH=2, initial sweep direction: anode) 1—[Fe3+] = 0.000 mol.dm-3 2—[Fe3+] = 0.179 mol.dm-3 In the reverse scan, the peak current density of the relative reduction reactions corresponding to the oxidation reactions increased and the initial potentials electro positively moved. Comparing with the cyclic without added ferric and attachment of bacteria on chalcopyrite, the prwave of the Fe3+ reduction reaction: Fe 3+ + e → Fe 2+ is more clearly when added ferric shown by peak D. On the freshly prepared surface of chalcopyrite there is almost no reduction reaction occurred in Fig.1, due to the slow kinetics and the irreversibility of the Fe2+/Fe3+ couple on this sulfide. It demonstrate that the thiobacillus ferrooxidans and ferric have positively affected on the dissolution of chalcopyrite. Because of the more production produced during the anode oxidation process, the current density of the other reduction reactions in cathode scan increase relatively.
4.
CONCLUSIONS The anode oxidation process of chalcopyrite includes many intermediate transient reactions. During the dissolution, it is exist the intermediate production of chalcocite and covellite. At lower scanning potentials the iron of chalcopyrite is extracted by ferrous form, but at the relative high potential it is by the ferric. When in the presence of Thiobacillus ferrooxidans, the peak current density and reversibility of the oxidation reaction increased, and the oxidation potential negatively moved. It demonstrated that apart from enhancing the metallic ion extracting reaction, the Thiobacillus ferrooxidans has also contribution to the oxidation of element sulfur formed on the chalcopyrite surface during the intermediate process. The decrease of pH from 2.0 to 1.5 gives slightly inhibition to the dissolution of chalcopyrite. The Fe3+ added in the medium can enhance the anode oxidation of chalcopyrite. ACKNOWLEDGEMENTS This study was funded by national science foundation of China. 405
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REFERENCES 1. Ehrlich, H.L., (1999) Past, present and future of biohydrometallurgy. In: Internaltional Biohydrometallurgy Symposium (Ballester, A. and Amils, R., Eds.), Vol. 1, pp. 3-12. Elsevier, Amsterdam. 2. Brierley, J.A. and Brierley, C.L., (1999) Present and future commercial applications of biohydrometallurgy. In: International Bio-hydrometallurgy Symposium (Ballester, A. and Amils, R., Eds.), Vol 1, pp.3-12. Elsevier, Amsterdam. 3. P. Romano, M.L. Blazquez, F.J. Alguacil, et al., (2001) Comparative study on the selective chalcopyrite bioleaching of a molybdenite concentrate with mesophilic and thermophilic bacteria, FEMS Microbiology Letters, 196, 71-75. 4. G.S. Hansford, T. Vargas, (2001) Chemical and electrochemical basis of bioleaching processes, Hydrometallurgy, 59, 135-145. 5. Wolfgang Sand, Tilman Gehrke, Peter-Georg Jozsa, Axel Schippers, (2001) Biochemistry of bacterial leaching - direct vs. indirect bioleaching, Hydrometallurgy, 59, 159-175. 6. Michael J. Nicol, Isabel Lazaro, (2002) The role of EH measurements in the interpretation of the kinetics and mechanisms of the oxidation and leaching of sulphide minerals, Hydrometallurgy, 63, 15-22. 7. Adibah Yahya, D. Barrie Johnson, (2002) Bioleaching of pyrite at low pH and low redox potentials by novel mesophilic Gram-positive bacteria, Hydrometallurgy, 63, 181-188. 8. Abha Kumari, K.A. Natarajan, (2001) Electrobioleaching of polymetallic ocean nodules, Hydrometallurgy, 62, 125-134. 9. B.F. Giannetti, S.H. Bonilla, C.F. Zinola, T. Raboczkay, (2001) A study of the main oxidation products of natural pyrite by voltammetric and photoelectrochemical responses, Hydrometallurgy, 60, 41-53. 10. Mishra, K.K., Osseo-Asure, (1988) Aspect of the interface electrochemistry of semiconductor, J. Electrochem. Soc., 135, 2502-2508. 11. C. Cómez, M. Figueroa, J. Muñoz, M.L. Blázquez, (1996) Electrochemistry of chalcopyrite, Hydrometallurgy, 43, 331-344. 12. Larazo, I., Martinez-Medina, N., Rodirguez, I. et al., (1995) The use of carbon paste electrode with non conducting binder for the study of minerals-chalcopyrite, Hydrometallurgy, 38, 277-287. 13. Palencia, I., Wan, R.Y., Miller, I.D., (1991) The electrochemical behavior of semiconductor natural pyrite in the presence of bacteria, Metal. Mater. Trans. 22B, 765-773. 14. C.S. Cha, C.M. Li. (1994) Powder microelectrodes. J. Electroanalytical. Chem. 368, 47-54. 15. Jose Pizarro, Eugenia Jedlicki, Omar Orellana and Jaime Romero, (1996) Bacterial populations in samples of bioleached copper ore as revealed by analysis of DNA obtained before and after cultivation, Appl. Enviro. Microbio., 62, 1323-1328. 16. Li Hongxu, (2001) Studies on the electrochemical mechanism and technology of sulfide bioleaching, PhD thesis, (In Chinese) Hunan, Changsha, China, 17-20. 17. Parker, A.J., Paul, R.L., Power, G.P., (1981) Electrochemistry of the oxidative leaching of copper from chalcopyrite, J. Electroanal. Chem., 118, 305-316. 18. Warren, G.W., Wadsworth, M.E., El-Raghy, S.H., (1982) Passive and transpassive anodic behavior of chalcopyrite in acid solutions, Metall. Trans. B, 13B, 571-579. 19. Stankovic, Z.D., (1986) The anodic dissolution reaction of chalcopyrite. Erzmetall, 39, 623-628. 406
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20. Biegler, T., Swift, D.A., (1979) Anodic electrochemistry of chalcopyrite, J. Appl. Electrochem., 9, 545-554. 21. Biegler, T., Home, M.D., (1985) The electrochemistry or surface oxidation of chalcopyrite, J. Electrochem. Soc., 132, 1363-1369. 22. Holliday, R.I., Richmond. W.R., (1990) An electrochemical study of the oxidation of chalcopyrite in acidic solution, J. Electroanal. Chem., 288, 83-98. 23. Warren, G.W., Sohn, H.J., (1985) The effect of electrolyte composition on the cathodic reduction of CuFeS2, Hydrometallurgy, 15, 133-149.
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
The influence of crystal orientation on the bacterial dissolution of pyrite S. Ndlovu and A.J. Monhemius Department of Earth Science and Engineering, Royal School of Mines, Imperial College of Science Technology and Medicine, London SW7 2BP, United Kingdom Abstract In order to understand the influence of crystallographic orientation on the mechanism of pyrite bioleaching, single crystals cut to expose plane orientations of 100, 111 and 110 were used for the study. Experiments were carried out both in the presence and absence of Thiobacillus ferrooxidans. Experiments to compare the extent of dissolution of the pyrite surfaces in bacterial and sterile solutions under similar solution conditions were also undertaken by matching the conditions in sterile solutions to those in bacterial leaching using an electrolysis cell. Differences in the reaction rates of the pyrite surface planes in both sterile and bacterial solutions have been observed. Furthermore, the results for the comparison between the bacterial and sterile leaching of pyrite samples under similar conditions indicate higher dissolution rates in the presence of bacteria. The microbial corrosion patterns generated on the surfaces were further used to study the leaching process. Microbial leaching of pyrite was observed to create surface corrosion patterns distinct from those of sterile leached samples. In addition the morphology of corrosion patterns arising from microbial leaching were found to slightly differ from one crystal plane to another while those in sterile leaching generally reflected the symmetrical arrangement of the crystallographic planes in the lattice on which they formed. The variation of corrosion patterns observed on the surfaces of bioleached samples seems to indicate a variation in cell-surface interaction from one crystal plane to the other. The results show that the surface properties of mineral sulphides may control the evolution of corrosion patterns and the initial oxidation kinetics in acid bacterial leaching. The overall analysis seems to indicate an influence of the primary cell-mineral interaction during the early leaching stage.
Keywords: bacterial attachment, pyrite, crystal orientation, corrosion patterns 1.
INTRODUCTION Pyrite is the most common metal sulphide in the mineral processing industry and is normally found in three main crystal forms, the most common being the cubic form where (100) surfaces predominate. This surface is close to ideal, and has a bulk termination of Fe and S species with a five-fold coordination of Fe sites and three-fold coordination of S sites, these being the respective bonding environments for the uppermost surface of Fe and S sites on a flat terrace. It is also found as octahedral and pyritohedral crystals, terminated by the (111) and (210) surfaces respectively. According to Guevremont et al (1998), the 409
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111 surfaces can either be sulphur or iron terminated with the surface Fe atoms and S2 groups being three coordinated and with each bonded to three species (S2 or Fe atoms respectively) in the layer below. Rarely, pyrite is found in the form of dodecahedral crystals terminated by the (110) surfaces. This surface structure has a four-fold coordination and lies perpendicular to the 110 direction, which according to Edwards et al (1998) is the direction characterized by a high density of disulphides in the pyrite lattice. Due to its ubiquity an understanding of the reactivity of pyrite is important especially for such applications as froth flotation and leaching, as well as for geochemical processes like the production of acid mine waters. All of these processes involve reactions at pyrite surfaces and, as a result, it is essential to understand the nature of the reactions occurring at these surfaces. Although some research has been done on the effect of crystallographic orientation on the dissolution process of pyrite, there are relatively few studies in the field of bioleaching that explain the aspects involved, especially the association if any, between the surface structure and the leaching process in the presence of surface attached bacteria. The main objectives of the work described in this paper are to investigate the influence of crystallographic orientation on the bacterial leaching of pyrite crystals by undertaking an examination of the dissolution rates and the surface corrosion patterns associated with the dissolution process and establishing the correlation, if any, between the surface attached bacteria, the observed dissolution pits, and the surface structure of the planes.
2.
MATERIAL AND METHODS
2.1 Bacteria culturing The strain of Thiobacillus ferrooxidans used in this experimental study was originally obtained from University College, Cardiff and was propagated in 9K medium at pH 1.8 (Silverman and Lundgren, 1959). The same medium was modified accordingly and used for experimental processes. The number of bacterial cells was estimated by direct counting using the improved Helber counting chamber. The final cell concentration used for experimental inoculation was approximately 1x108cells/ml. 2.2 Pyrite sample preparation Single crystal cubes (approximately 1cm3 and weighing 4.5-8g) of natural pyrite were characterised by the Laue X-ray diffraction method (Phillips Analytical X’Pert Data Collector). The 100 planes were found to be the principally occurring planes on the faces of the samples. The samples were cut with a diamond saw and polished from initial symmetric cubes parallel to the plane orientations of 100, 110 and 111. The samples were prepared with each surface of the required face to limit the leaching to about 0.5-1cm2 exposed area and covering the remaining surface with araldite epoxy resin. These were then washed with ethanol and dilute HCl before each experiment to remove any soluble material on the surface. The samples were made in pairs so that matching faces could be used for duplicate experiments. 2.3 Matched leaching experiments The aim of the experiments was to compare the extent of dissolution of the different crystallographic pyrite surfaces in bacterial and sterile solutions under similar solution conditions, using a method previously adopted by Driessens et al (1999) in the study of sphalerite. In the present study, experiments were done by simulating conditions experienced in bacterial leaching by controlling the redox potential in sterile leaching through ferrous to ferric iron electrolytic oxidation. The apparatus used was a two410
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compartment electrolysis cell in which the two compartments were separated by an ion exchange membrane (Figure 1). The sterile leaching reactions were done in the working compartment of the electrochemical cell, whilst the bacterial leaching experiments were done in a separate reactor. Both the cathode and anode were made of platinum foil. The redox potential in the working compartment was matched to those measured in parallel bacterial leaching experiments, by manipulating the current to the platinum cathode. Thus, in sterile leaching a current was applied to oxidise the ferrous to generate the same solution redox potentials as measured in the bacterial experiments. The redox potentials were measured using a platinum electrode with a silver/silver chloride reference electrode. Redox potential electrode
Stirrer
Potentiostat
Working cell
Pyrite sample enclosed in epoxy resin
Counter cell
Membrane separator
Figure 1. Schematic diagrams of the apparatus for re-oxidation of ferrous iron in chemical leaching experiments The reaction for pyrite mineral dissolution is
FeS2 + 8H 2O → Fe2 + + 2SO4
2−
+ 16 H + + 14e −
While reoxidation at the platinum electrode is defined mostly by,
Fe 2+ → Fe 3+ + e − 2.3.1 Leaching media The leaching experiments were carried out in 9K basal medium (3.00 g (NH4)2SO4, 0.50g K2HPO4, 0.50g MgSO4.7H2O, 0.01g Ca(NO3)2 and 0.10 g KCL dissolved in 1000ml distilled water) at a pH of 1.8 adjusted with sulphuric acid. A cell suspension (10% v/v.) with an initial population of 1x108cells/ml was added to the bioleaching solution. Initially, a ferrous iron solution giving a total Fe (II) concentration of 0.05M was added to each solution to provide a source of energy for the bacteria in the bioleaching experiments and to gradually generate ferric ions by electrolytic oxidation in the sterile leaching experiments. The leaching volume used for each experiment was100ml. 2.4 Analytical methods A scanning electron microscope, JEOL JSM T220, was used to monitor bacterial adhesion and surface changes occurring on the samples during leaching. The samples were removed from solution, washed with acidified water, then acetone, dried, coated with a very thin layer of gold and observed under SEM. The samples were observed at weekly 411
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intervals. The total iron concentration in the bio-oxidation and chemical leaching solutions was measured at 2-day intervals using a Perkin-Elmer 1100B atomic absorption spectrophotometer with an air /acetylene flame.
3.
RESULTS AND DISCUSSION
3.1 Leaching rates The results for the leaching experiments indicated that the surfaces of pyrite react at different rates (Figure 2a and 2b). This is most likely due to the differences in the surface atomic geometry and chemical surface states that predominate on pyrite surfaces. However, since these states are only present on the pyrite surfaces rather than in the bulk crystal structure, it is possible they will play a major role only during the early stage of the reaction process. Released Iron:Bacterial Leaching
600
400
400
200
200
Redox potential/ mv
0
0 0
2
4
6
8 10 12 Tim e/days
110 111B redox potential
111A 100
14
800
800
600
600
400
400
200
200
0
redox potential/ mv
600
Released iron/ mg/L/cm2
800
Released iron/ mg/L/cm2
Released Iron: Sterile Leaching
800
0 0
2
4
6 8 10 12 14 Tim e/days 110 111A 111B 100 redox potential
(a)
(b) Figure 2. Leaching trends of the planes in (a)-bioleaching and (b)- sterile solutions under matched conditions. Leaching conditions: bacterial leaching: 9K medium, 10% (vol./vol.) bacteria inoculumn, 0.05M Fe2+ solution concentration, pH 1.8. Sterile leaching: 9K medium, 0.05M Fe2+ solution concentration, pH 1.8 The experiments were carried out in duplicate, corresponding to the matching faces of the cut crystal surfaces, A and B. While the behaviour of the matching faces of the 100 and 110 planes was not different, those of the 111 surfaces varied (Figure 2a and 2b). From surface analysis by SEM, a lower pit density was observed on one 111 surface layer, whilst the other adjacent surface had a higher pit density. In addition, the surface with a high pit density had a lower dissolution rate compared to the other over the first seven days of leaching. Since for pits to develop, it is necessary that dissolution proceed faster in the direction of the pit than on the surface surrounding it, it can be suggested that the plane with a lower pit density had the most reactive atoms on its surface. Conversely, the one with a higher pit density had the least reactive surface atoms. Furthermore, since the dissolution rate was measured in terms of released iron, it can then be assumed that the plane with a lower pit density is the iron-terminated surface (thereafter referred to as A), and the other one with the higher pit density is taken as the sulphide terminated surface (B plane). The overall reaction rate trend observed at the end of the 14-day leaching period for both sterile chemical and bacterial leaching solutions was: 111A>111B>110>100 412
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3.1.1 Bacterial leaching The 111A surface showed the highest overall iron dissolution after 8-10 days of leaching. However, this surface generally showed a lower dissolution rate compared to the 100 and 110 planes during the early stages of leaching when the redox potential was low. The iron released from the 111A plane only increased above that of the other planes after a sharp increase in solution potential, Figure 2(a). In addition, although the 111A plane showed an overall higher release of iron in the presence of bacteria compared to that achieved in sterile solutions, the degree of leaching enhancement (ratio of iron released under bacterial leaching to that released in sterile leaching) did not vary much, remaining generally between 1.15 and 1.20. The 111B surface, which is assumed to be the sulphurterminated plane, showed the lowest dissolution during the early leaching stages, but this gradually increased to give an overall iron release greater than that of the 110 and 100 planes, but slightly less than that of the 111A plane. The surface geometry and lower surface atom coordination leads to the 110 plane being more reactive and producing a higher amount of dissolved iron compared to the 100 plane. In addition the results indicated a higher amount of dissolved iron for the 100 and 110 surfaces in the presence of bacteria under similar conditions during the early stages of leaching. This seems to indicate that during this initial stage of leaching, the presence of bacteria at low concentrations of ferric iron in solution enhances the initial oxidation process. This is further discussed in Section 3.2. 3.1.2 Sterile matched leaching In general simulation of conditions experienced in bacterial leaching through control of the solution potential greatly increased the dissolution of pyrite under sterile leaching conditions. Sasaki et al (1998), have reported that chemical dissolution of Fe is more rapid from pyrite than oxidation of S species in the lattice, leading to the formation of elemental sulphur on pyrite. As a result, passivation due to the inaccessibility of Fe sites to the leach media occurs. In acid solutions, the overall oxidation reactions involve the formation of sulphur and/or a metal-deficient sulphide at low overpotentials and sulphate at high overpotentials. It is possible that by maintaining a high redox potential, bacteria catalyse the sulphate-forming reactions, preventing the accumulation of a sulphur layer. Thus, if conditions in sterile leaching are matched to those of bacterial leaching, high redox potentials are maintained and sulphate, instead of sulphur, formation is promoted. Significantly, pyrite dissolution is enhanced and it becomes possible to obtain characteristic bioleaching rates in an abiotic system provided that the redox potential is kept high. A further comparison of the results for matched leaching experiments and for experiments undertaken in sterile solutions whose redox potentials were not matched to the bacterial conditions indicated that the amount of iron released for non-matched sterile conditions is on average about 20-25% of that released under matched conditions. This shows the significant influence of the solution potential and hence ferric iron in the overall dissolution process. 3.2 Pitting morphology 3.2.1 Bacterial leaching One of the objectives of this study was to establish the role if any, of cell-mineral interactions on the evolution of corrosion pits in bacterial leaching during the early stages of leaching in the absence of significant amounts of ferric ion. This was to be done by determining whether pits were establised at locations on the mineral surface where there were attached bacteria. However, although bacteria cells were observed on the 111B, 110 413
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and 100 surfaces, their distribution was random and there was no close proximity to the corrosion pits as observed by Bennet and Tributsch (1978), Rodriguez-Levia and Tributsch, (1988). No significant bacterial colonisation was observed on the 111A surfaces. On the other hand, the observed corrosion patterns, although irregularly distributed over the surface, showed a significant orientation. It was observed that the surface planes with high ratios of sulphur/iron atoms e.g. the 111B, 100 and 110 generally generated elongated pits similar in shape to bacteria, but much larger than the bacteria cell dimensions (Figure 3a and 3b). [110] [100]
(b) (a) Figure 3. (a)-Corrosion patterns on (a)-110 plane; note the circular pits in close proximity to well-developed elliptical pits and (b)-corrosion pits on 111B planes One other interesting aspect was the observation of surface films on these 110, 111B and 100 surfaces (Figure 4a). These were observed both in the early stages and after about a week of leaching and have previously been observed by other researchers on pyrite particles (Rodriguez-Levia and Tributsch, 1988, Edwards et al, 2001). Unlike in the work of these authors however, no bacteria were generally observed in close proximity to these films.
(a) (b) Figure 4. (a)-Surface films observed on 110, 100 and 111B surfaces. Inset shows enlarged view of the film possibly generated by bacteria. (b)-pitting on the 111A plane (bar size 10µm). 414
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The film layers exceeded the amount of surface covered by the cells indicating that the film spreads beyond the area of bacterial contact (inset Figure 4a). The fact that surface films were observed mostly on those planes developing elongated dissolution pits seems to indicate that the extension of the film beyond the area of direct bacteria contact extends the zone/compartment of cell interaction with the surface, subsequently generating larger pits. The fact that not all the planes developed this film (111A did not) and on those where it did, only certain areas of the surface were covered, suggests that film formation depends on bacteria recognizing certain sites on the surface for colonisation, with the subsequent formation of a film. Furthermore, since the surface films were observed even after two weeks of leaching without any bacteria being observed in close proximity to the films or the pits, this may suggest that the film contains some constituent that enhances leaching after the bacterial cells have moved from these specific sites. Calculations of pit depths and pyrite lattice layers consumed for all the planes further support this hypothesis. Thus, it seems that it is the initial recognition of active sites on crystal surfaces and subsequent attachment of bacteria on these specific sites that controls the initial attack on the surface and subsurface layers (Ndlovu and Monhemius, 2003). It has been mentioned in the literature that bacteria seem to have a preferential attachment to specific sites such as sulphur enriched zones (Mustin et al 1993, Edwards et al 1998) and they preferentially oxidize sulphur during the lag phase, while the released iron, despite being preferentially leached into solution, remains in the ferrous state (Sasaki et al 1995, 1998, Jae-Young et al 2001). The crystal structure of pyrite indicates that the directions of high sulphide density are the <110> and the <100>. If sulphur removal occurs in a crystallographically controlled manner, then the attachment of bacteria should be concentrated along these sulphur-enriched zones. This is supported by the observed high density of the corrosion pits on the 110 and the sulphide-terminated 111 surfaces (Figure 3a and 3b). The leaching patterns observed on the 110 planes were found to lie parallel to the 110 directions. On the other hand, the 111 plane is bounded by 110 directions and, significantly, corrosion patterns on the 111B plane were observed to occur parallel to the edges of the surface. Both surfaces were covered with elliptical/rod-like pores and quasi-circular pits. In general the elliptical pits developed in pairs and circular pits were found mostly in close proximity with the developed elliptical ones (Figure 3a). This suggests that bacteria possibly grow and divide in-situ with the daughter cell developing a pit near to the mother cell creating pairs of pits as illustrated in Figure 3a and 3b. The importance of the surface structure on the leaching process of the pyrite crystal can also be understood by considering the leaching trend of the 111B plane (Section 3.1.1). This crystal plane is dominated by an immediate sulphide sulphur layer, which initially hinders the direct exposure of iron to the solution. Therefore, initially iron release is slow and sulphur release is favoured until, due to the action of attached bacteria, the surface becomes relatively enriched in iron and iron release becomes favoured. Thus in the early stage of leaching, the rate of erosion of the sulphide surface is limited by the rate that sulphur can be removed from the surface by either the bacteria or the chemical reaction (in sterile leaching). Consequently, as the leaching period progresses, the rate of iron dissolution increases for this plane. By assuming that pit formation is the result of enhanced bacterial attack at the attachment sites during the initial leaching stage, correlations between substrate-based interactions as defined by pit density measurements and solution analysis further indicated that the initial leaching on pyrite surfaces by bacteria is dominated by the reactions occurring at the cell-mineral interfaces (Ndlovu and Monhemius, 2003). 415
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On the 100 plane the leaching patterns appeared as both circular pores that tended to penetrate into the crystal face and elliptical pores that penetrated into and grew along the plane, although not to the same extent in axial length as those observed on the 110 plane (<20µm on 100, compared to up to 50µm for 110 planes). If the elliptical pits arise as a result of the persistence of disulphide atoms along a crystallographic direction, by considering the checkerboard structure of alternate disulphide and Fe atoms existing on the 100 planes, their development will be bounded/restricted by the Fe atoms. As a result they will not develop to a large extent along the 100 directions, resulting in a tendency to circular pit formation. The 111A surface did not reveal any corrosion pits during the first days of leaching. With the gradual production of ferric iron however, circular pits appeared (Figure 4b). This behaviour can be explained from both surface and solution analysis (Section 3.1). The absence of a significant bacterial colonisation and elliptical pits on this surface compared to other surfaces in the early stages of leaching, followed by the presence of mostly quasi-circular etch pits at later stages, suggests that cell-mineral interaction is minimal. Therefore, the dissolution on these surfaces occurs mainly due to the conditions created by the bacteria in bulk solution, that is, ferrous to ferric oxidation. Furthermore, the leaching trend of the 111A plane as observed in Section 3.1 indicated an increase in the dissolution rate with a sharp increase in solution potential and hence ferric iron. Since the dissolution of pyrite by ferric iron is taken as an indirect mechanism, this further suggests that this mechanism contributes significantly in the dissolution process of this pyrite surface. In addition, a comparison of the active surface area as calculated from the pit morphology analysis and solution chemistry confirmed a low cell-surface interaction with a non-localised surface-wide phenomenon governing the dissolution process for this surface (Ndlovu and Monhemius, 2003).
3.2.2 Sterile Leaching Sterile chemical leaching was characterised by leaching patterns distinct from those observed on bioleached samples. There were slight differences between the pitting morphologies observed from one crystal plane to the other, with the corrosion patterns having symmetries related to the crystallographic surface orientations. Figure 5(a) shows typical corrosion pits observed on the 110 and 100 surfaces, while Figure 5(b) shows the pits observed on the 111 surfaces.
Figure 5. Corrosion pits observed under matched leaching conditions. 5(a) -110 and 100 surfaces, rhombic pits; label 1-pit faceting and 2-merging of individual rhombic pits. Figure 5(b) shows triangular pits on the 111 surfaces (see enlargement) 416
Bioleaching Applications
The 100 and 110 surfaces were characterised by rhombic pits, whilst the 111 surfaces were characterised by triangular pits. The pit density and pit wall dimensions changed gradually as leaching proceeded, with some individual pits merging to form dissolution channels (Figure 5a). The pits further showed defined faceting indicating that the influence of crystalline orientation is clearly dominant in the relatively slower reactions that occur in the matched leaching experiments, where the concentration of the ferric iron gradually increases. The formation of faceted pits on the crystal surfaces suggests a contribution of an orientation-controlled type of dissolution mechanism in the leaching process. This suggests a relationship between the etch figures, crystal orientation and the overall dissolution processes of the surface.
4.
CONCLUSION The work carried out has shown differences in the initial reaction rates of the pyrite surface planes in both sterile and bacterial solutions. This has been explained as most likely being due to the influence of the differences in the surface atomic geometry and chemical surface states that dominate on pyrite surfaces. In addition, the correlation between the attached bacteria, the appearance of the surface organic films and the generation of elongated pits supports the influence of cell-mineral interaction in the initial oxidation process and the initiation of the leaching patterns (as observed by SEM analysis) on the crystal surfaces at low concentrations of ferric iron in solution. Most importantly, the absence of bacterial colonisation and elongated pits on the iron terminated 111 surface in the early stages of leaching seems to indicate that the most signficant step in the early leaching process is probably the recognition and subsequent attachment of the bacteria to the active (sulphur) sites on the initial surfaces. This initial process subsequently controls the leaching progression and defines the type and evolution of pits. ACKNOWLEDGEMENTS The financial support by the Institution of Mining and Metallurgy through the Stanely Elmore Fellowship Fund, the Minerals Industry Educational Trust (MIET) and the Arthur Bensusan Memorial Fund (Zimbabwe) is gratefully acknowledged. Thanks are due also to Professor F.D. Pooley of University College, Cardiff for donating the bacterial culture used in this work. REFERENCES 1. Bennet, J.C. and Tributsch, H. (1978) Bacterial Leaching Patterns on Pyrite Crystal Surface, Journal of Bacteriology, 134 (1), 310. 2. Driessens, Y.P.M, Fowler, T.A. and Crundwell, F.K. (1999) A Comparison of the Bacterial and Chemical Leaching of Sphalerite at the Same Leaching Conditions. In: Biohydrometallurgy and the Environment Toward the Mining of the 21st Century, Vol. A, IBS '99, Amilis, R. and Ballester, A. (Edts.), Elsevier Science, Amsterdam, 201. 3. Edwards, K. J. (1998) Microbial Oxidation of Pyrite: Experiments Using Microorganisms From an Extreme Acidic Environment, American Mineralogist, 83, 1444. 4. Edwards, K.J., Hu, B., Hamers, R.J. and Banfield, J.F. (2001) A New Look at Microbial Leaching Patterns on Sulphide Minerals, FEMS Microbiology Ecology, 34, 197.
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5. Guevremont, J.M., Elsetinow, A.R., Strongin, D.R., Bebie, J. and Schoonen, M.A. (1998) Structure and Sensitivity of Pyrite Oxidation: Comparison of the (100) and (111) Planes, American Mineralogist, 83, 1353. 6. Jae-Young, Y., McGenity, T.J. and Coleman, M.L. (2001) Solution Chemistry During the Lag Phase and Exponential Phase of Pyrite Oxidation by Thiobacillus Ferrooxidan, Chemical Geology, 175, 307. 7. Mustin, C., Berthelin, J., Marion, P. and Donato, P. (1993) Surface Sulphur as Promoting Agent of Pyrite Leaching by Thiobacillus ferrooxidans, FEMS Microbiology Reviews, 11, 71. 8. Ndlovu, S. and Monhemius, A.J. (2003) Correlations Between Reaction Rates and the Evolution Of Corrosion Pits in the Bacterial Leaching of Pyrite. Submitted to Trans IMMM. 9. Rodriguez-Levia and Tributsch, H. (1988) Morphology of Bacterial Leaching Patterns of Thiobacillus ferrooxidans on Synthetic Pyrite, Archives of Microbiology, 149, 401. 10. Sasaki, K., Tsunekawa, M., Ohtsuka, T. and Konno, H. (1995) Confirmation of a Sulphur-rich Layer on Pyrite after Oxidative Dissolution of Fe (III) Ions Around pH 2, Geochim. Cosmochim. Acta, 59, 3155. 11. Sasaki, K., Tsunekawa, M., Ohtsuka, T. and Konno, H. (1998) The Role of Sulphuroxidising Bacteria, Thiobacillus thiooxidans in Pyrite Weathering, Colloids and Surfaces A, 133, 269. 12. Silverman, M.P. and Lundgren, D.G (1959) Studies on the Chemoautotrophic Iron Bacterium Thiobacillus ferrooxidans, I: An Improved Medium and Harvesting Procedure for Securing High Cellular Yields. Journal of Bacteriology, 77, 642.
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
The influence of temperature and pH on the bioleaching of copper from a flotation concentrate of chalcopyrite Medrano-Roldán, H.a, Salazar, M.F.M.a, Pereyra-Alférez, B.b, Solís-Soto, A.a, Ramírez-Rodríguez, D.G.a, Alvarez-Rosales, E.c and L.J. Galán-Wongb a
Instituto Tecnológico de Durango, Unidad de Alimentos y Biotecnología Industrial, 34080 Durango, Dgo. México b Facultad de Ciencias Biológicas, Universidad Autónoma de Nuevo León, San Nicolás de los Garza, N.L. México c Cía. Minera Mexicana de Avino S.A. de C.V. Durango, Dgo. México
Abstract The process of bioleaching of copper has been improved through the use of several chemical compounds and environmental and operational conditions in shake flask and tank leaching. The main objective of this project was to increase the knowledge on the effects of temperature and pH values on the bioleaching of copper from a chalcopyrite flotation concentrate, by using a native Thiobacillus thiooxidans strain. The effect of temperature on the copper extraction showed that the asymmetry of the curve is typical. Usually the curve is steeper at supraoptimal temperatures than in the suboptimal region. The effect of pH is similar to that of temperature in that an optimal pH value exists. Nevertheless, the shape of the curve in not quite the same and a broader plateau is observed.
Keywords: chalcopyrite, copper bioleaching, temperature, pH 1.
INTRODUCTION In the biological leaching of copper from chalcopyrite concentrates, the process has been improved through the use of several chemical compounds and environmental and operational conditions in shake flask and tank leaching where technical and economical parameters should be determined (1). Biohydrometallurgical extraction of copper from chalcopyrite can be described by the following electrochemical reactions (3): Anodic: 2 CuFeS2 + 16 H 2 O + H 2 SO 4
→
2 Cu 2 + + 2 Fe3+ + 5 SO 42 − + 34 H + + 34 e −
Cathodic: 34 H + + 34 e − + 8
1 O2 2
→
17 H 2 O
Sum: 1 2 CuFeS2 + 8 O 2 + H 2 SO 4 2
bacteria
2 CuSO 4 + Fe 2 (SO 4 )3 + H 2 O
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With this information in mind, Minera Mexicana de Avino S.A. de C.V. a mine company in Durango State, Mexico, has been interested in applying microbial leaching techniques by use of baffled Erlenmeyer flasks in tests in order to improve oxygen transfer conditions at this level and several temperature and pH values. Experiments were carried out in order to remove copper from a chalcopyrite flotation concentrate.
2.
MATERIALS AND METHODS
2.1 Microorganism An adapted and native culture identified and characterized as Thiobacillus thiooxidans used in this study was originally isolated from samples of acid mine waters obtained from Minera Mexicana de Avino S.A. de C.V. Company. The culture was maintained on a modified Silverman and Lundgren 9K medium (4), in which chalcopyrite concentrate was used as the source of energy. 2.2 Substrate The chemical assay of the chalcopyrite flotation concentrate was as follows: Pb, 5%; Cu, 29.6%; Fe, 15.6%; S, 25.88%; insolubles, 16.4%. 2.3 Shake flask experiments These tests were carried out in 500 ml baffled Erlenmeyer flasks on a gyratory incubator shaker NBS Model G-25 at 120 rpm, 100 ml of iron-free 9K medium, at pH (1.5-3.0), temperature (30-70°C), 20% (w/v) of pulp density, and 20% (v/v) of inoculum. 2.4 Analytical techniques The oxidation of ferrous sulphate was monitored by determining its residual concentration in the medium following the 1,10-phenantroline method (5) and from the redox potential measurements. In order to measure the concentration of total iron in solution the ferric iron was reduced to ferrous, after filtration of the medium, using hydroxylamine as reducing agent and determining this concentration by the previously mentioned method. Subsequently, the concentration of ferric iron in solution was determined by difference between the ferrous and total iron concentrations. Redoxpotential was measured by using a redox electrode with a combination platinum / reference (Ag / AgCl). Soluble copper was determined by the iodometric method (5). 3.
RESULTS AND DISCUSSION
3.1 Effects of temperature and pH The effect of temperature on the copper extraction given in Figure 1 shows that the curve is asymmetry. Usually the curve is steeper at supraoptimal temperatures than in the suboptimal region. Although an Arrhenius equation of the type µ = exp(-E/RT) is suitable for describing the temperature effect below the optimal temperature, this model fails to predict the behavior at optimal and supraoptimal temperatures. Few mathematical models have been published that are adequate for describing the complete curve. Moreover, they do not refer, specifically, to bioleaching system (2).
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Figure 1. Relationship between copper release and temperature values This work has demonstrated that the culture cited above can be adapted to the leaching of copper from a chalcopyrite concentrate with high efficiency. The high extraction (92%) obtained shows that the problem of incomplete chalcopyrite leaching typically associated with the use of T. ferrooxidans as a leaching organism does not ocurr when Thiobacillus thiooxidans is used. Our native culture suggests that Thiobacillus thiooxidans could offer a more economically attractive route for the bioleaching of chalcopyrite than the process using T. ferrooxidans. This aspect was observed in our laboratory. One point that could be of interest to explore is related to the operation at supraoptimal temperatures, where an increase in the leaching activity is to be expected due to a higher ratio of energy metabolism to biomass formation as a function of the copper extraction in the chalcopyrite concentrate.
Figure 2. Relationship between copper release and pH values The effect of pH shown in Figure 2 is similar to that of temperature in that an optimal pH value exists. Nevertheless, the shape of the curve is not quite the same as that of temperature and a broader plateau is observed in this case.
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It has been shown that the growth of T. ferrooxidans in a defined medium with Fe2+ as energy source produces a rise in pH due to proton consumption. The situation is similar in the leaching of actual ores. Acid must be added to keep the pH at the desired value, especially in the earlier stages of leaching. The acid consumption is increased when using tailings, which have been obtained in our laboratory (2).
REFERENCES 1. M.F. Salazar. Tesis de Maestría. Instituto Tecnológico de Durango. México (1999). 2. H. Medrano-Roldán, T.E. Flores, A.H. Pérez, A. M. Rentería, J.L. Galán-Wong, B.E. Orrantia and M. Monroy-Fernández. UBAMARI. 39 (1996) 77. 3. M. Valayapetre and A.E. Torma, Metallurgy, 32 (1978) 1120. 4. M.P. Silverman and D.G. Lundgren, J. Bacteriol., 77 (1959) 642. 5. A.I. Vogel, Vogel’s Textbook of Quantitative Chemical Analysis, 5th Ed., Longman Group Ltd. London (1989).
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
The role of chemolitotrophic bacteria in the oxide copper ore heap leaching operation at Sarcheshmeh Copper Mine Seyed Baghery S.A.*a, Shahverdi A.R.b, Oliazadeh M.c a
b
R & D Division, National Iranian Copper Industries Company, Rafsanjan, Iran Biotechnology Dept., Faculty of Farmacy, Tehran Medical Sciences university, Tehran, Iran c Faculty of Mining Engineering, Tehran University, Tehran, Iran
Abstract The role of chemolitotrophic (inc. mesophilic) bacteria in the oxidation of copper sulfide minerals in oxide copper ore heap leaching operation of Sarcheshmeh Copper Complex has been investigated. In the present study, it was determined that about 34% of the heap copper content, existed in the form of sulfide minerals such as chalcopyrite and chalcocite, the remaining were oxide copper ores. Hence, the heap consisted of a mixed sulfide and oxide copper ore. Sulfide minerals are insoluble or partially soluble in the chemical heap leaching conditions. However, studies showed that about 30% of the heap sulfide copper content has been leached naturally during the pad’s irrigation. Previous samplings from the heap showed there were a lot of native chemolititrophic bacteria belonging to the genera Acidithiobacillus, Leptospirillum and Sulfobacillus. To investigate a probable relation between the existence of bacteria and the leached sulfide ores, several samples were taken from different depths of a newly leached pad and analyzed for their bacterial number, pH, Eh, total soluble iron (TSI) and pyretic iron. It was found that bacteria had an important effect on the value of pH, Eh, and TSI of the heap. As these are very active oxidizing bacteria, they created a suitable condition for copper sulfide ore oxidation. Copper extraction from sulfide ores in the heap can then be attributed to the bacterial activity. It was claimed that the remained sulfide copper ore in the heap can be leached biologically after finishing chemical leaching.
Keywords: copper heap, acidithiobacillus, sulfide ores 1.
INTRODUCTION Natural bioleaching of sulfide minerals has been taking place as long as the history of the world, but it is only in the last few decades that we have realized that bioleaching is responsible for liberating some metals. The application of the bioleaching reactions for copper has been exploited and used to develop suitable methods to recover copper from copper bearing solutions [1]. The heap leaching of copper has been practiced for several decades, mostly with oxide ores. Probably bacteria aided some of these oxide operations, but this was without * corresponding author:
[email protected]
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serious study. This is the subject of the current paper for determining the role of chemolitotrophic bacteria in the chemical (acid) heap leach operation at Sarcheshmeh Copper Mine. It should be noted here that the copper ore used for the heap leaching operation at this mine can be considered as a mixed oxide/sulfide copper ore which is composed of 66% copper in the form of oxide minerals and 34% as sulfide minerals mainly in the form of chalcocite and chalcopyrite. There are some industrial experiences that for such mixed copper ores a process of chemical and then biological operation has been practiced [2, 3]. No considerations were taken for the biological operation part while Sarcheshmeh copper heap was constructing. Copper sulfide minerals are insoluble or partially soluble at the relatively short time (about 90 days) considered for dissolution of copper oxide ores, but studies in this project showed that about 30% of total copper in the form of sulfide minerals has been leached naturally. Several samplings from different parts of the heap operation showed there were a lot of bacteria (more than 105/ ml) from the genera, Acidithiobacillus, Leptospirillum and Sulfobacillus [4]. Hence, sulfide copper dissolution can be attributed to the bacterial activity and therefore it may be possible to recover the copper (which amounts to about 28000 tones) from the buried sulfide minerals. So the first aim of this project was to determine the role of the above bacteria on some of the heap factors and then, propose some processes that may be useful for recovering the left copper in the heap.
2.
MATERIALS AND METHODS In order to find out the role of chemolitotrophic bacteria in the oxidation of sulfide ore fraction of heap leaching operation of Sarcheshmeh Copper Mine, a newly finished irrigation pad was selected in September. The temperature in this month at Sarcheshmeh Copper Mine is around 26.3ºC. Three sampling area, each with a distance of 20 meters were marked and then from the surface of the pad down to the depth 2.75 m, several samples of the leached ore were taken each at intervals of 0.25 m and then a total of 36 samples were transferred to the laboratory. In order to measure the pH and the Eh values, bacterial count and total soluble iron concentration in each solid sample, a one kg sample was added to one liter of acidic (pH=1.9) distilled water in a 5 liter volume beaker and mixed thoroughly for 20 minutes. After settling, a sample of the clear supernatant was used for determining the above parameters. The bacteria were counted microscopically using a slide counting chamber. pH and Eh values were measured by a WTW pH/Eh meter model 323. The total iron was analyzed by the AAS method. The solid residue of each sample was washed, dried, pulverized and analyzed for iron by the AAS method. For each specific depth, an average of the above three sampling sites was recorded. 3.
RESULTS AND DISCUSSION Regarding the environment temperature (around 26.3ºC) at the time of sampling, the dominant bacteria belonged to the genera Acidithiobacillus and leptospirillum. In warmer months, moderately thermophiles were also isolated from this oxide heap [4]. Table 1 shows the temperature at Sarcheshmeh Copper Mine in different months.
Table 1. The average of maximum temperature at Sarcheshmeh Mine from 19731997 Month Tem. (ºC) 424
Jan.
Feb.
Mar.
Apr.
May
Jun.
Jul.
Aug.
Sep.
Oct.
Nov.
Dec.
6.2
7.1
10.5
16.5
22.1
27.4
28.4
27.2
26.3
18.1
13.6
7.8
Bioleaching Applications
A theory of bacterial growth in a heap holds that the major area of bacterial growth is in the top 1.5 meters of a leach pile [1]. In this paper it was decided to take some samples from the top to a deeper depth of 1.5 meters and then compare the results. Figure 1 shows the number of bacteria and the Eh changes at different depth of the pad. There is a close direct relation between these two factors. At the top of the heap, there are few bacteria and a low Eh value as well. This may be because of the very rapid environment changes at the top of the heap such as different day and night temperatures, sunshine and the high evaporation of introducing solutions to the heap. Beneath the surface where the harsh environment conditions disappeared, the number of bacteria and the Eh increased considerably. After that, down to the depth 0.75 m, the above factors decreased; there was no special reason here. There might not be sufficient sulfide minerals to support the bacterial growth. From the depth 0.75 down to 1.25 m, again the bacterial number and the Eh value increased. The lower depths showed a relatively constant number of bacteria. These are nearly non-growing cells washed from the higher levels. According to Schnell, the oxygen level at lower depths of 1.5m in a nonaerated heap (like the present case) drops to below 5% [1]. It can be concluded here that bacteria may have a determining role in the value of oxidation-reduction potential of the mixed copper ore heaps.
Figure 1. Bacterial count and Eh value changes at different depths of the pad A better understanding of the bacteria and their role in the heaps has been shown in figure 2, which demonstrates the bacterial number and the pH value changes at different depths of the pad. At the top of the pad, the pH of the leached ore was high together with few bacteria. Beneath the surface, down to the depth 1.25 m, along with the more active bacteria, the pH was lowered. From this point down to the lower depths, the pH value started to increase gradually. The pH value of the leached solution determines the solubility of ferric ion, which is a key factor for sulfide mineral oxidation. Figure 3 shows the pH and the TSI variations at different depth of the pad. There is an antithetic relation between these two factors. The low concentration of TSI was due to the long time irrigation of the pad. During this time, the total iron in the acid soluble ore form has been leached and what had remained, was the iron in the form of pyrite. So the TSI changes could be attributed to the bacterial activity. The results of bacterial growth and activity on the oxidation of pyrite have been shown in figure 4. Supposing the homogeneous distribution of pyrite and the complete oxidation of acid soluble iron ores, one can observe following bacterial activity at the upper depths, ferric ions were produced and while going down the heap, oxidized the existence pyrite down to the 1.25 m. Around this point bacterial activity was going to 425
Bioleaching Applications
cease and hence, down to the depth 2.75 m, pyretic iron started to increase gradually as there was not enough ferric ions and bacterial growth for continuing pyrite oxidation.
Figure 2. Bacterial count and pH value changes at different depths of the pad
Figure 3. TSI concentration and pH value changes at different depths of the pad
Figure 4. Bacterial count and iron (in the form of pyrite) changes at different depths of the pad
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Bioleaching Applications
4.
CONCLUSIONS AND SUGGESTIONS - According to the results, native bacteria have a determining role in the values of pH, Eh, TSI and sulfide ores at the upper surfaces of the copper heap operation. They create a suitable condition for sulfide mineral leaching. - It should be noted that the results in this research came from a chemical heap leach operation in which no preparations were made for bacterial activity. - Regarding the mixed nature of the ore used for chemical (acid) heap leaching at Sarcheshmeh Copper Mine, it would be useful if some preparations were made for copper sulfide ore oxidation, while heap construction. Now, the leaching of the heap operation has been completely terminated. So, attentions should be paid to the huge amounts of copper in the form of sulfide minerals especially chalcopyrite that has been buried in the heap. - In order to leach the copper sulfide ores left in the heap, it may be possible to bioleach the upper two meters of the heap first, then discard it and run a same process for the next two meters and going down to the bottom of the heap. - Another possible process may be applying a ferric sulfate leaching system in which the produced ferrous ion will biologically be oxidized to ferric ion.
REFERENCES 1. Henry A. Schnell, (1997), Bioleaching of copper, in: D. E. Rawlings (ed.) Biomining: Theory, Microbes and Industrial Processes, 21-43. 2. F. Acevedo and J. C. Jentina, (1993), Bioleaching of minerals – a valid alternative for developing countries, Journal of biotechnology, 31, 115-123. 3. Asok Sen, R. C. Gupta, M. S. Prasad, N. Ramesh, S. K. Ray And K. Pal, (1993) Biohydrometallurgical Operations, in: A. E. Torma, J. E. Way and V. I. Lakshmanan (eds.), Biohydrometallurgical Technologies, Vol. 1, 185-194. 4. S. A. Seyed Baghery, (2001), Isolation and preliminary identification of some ironand sulphur-oxidizing bacteria from the Sarcheshmeh Copper Complex, in: S. T. Ciminelli and O. Garcia Jr. (eds.), Biohydrometallurgy: Fundamentals, Technology and Sustainable Development, Part A, 393-396.
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Three-stage revolving drum biohydrometallurgical reactor for continuous operation G. Loi, P. Trois and G. Rossi (*) Dipartimento di Geoingegneria e Tecnologie Ambientali Università di Cagliari – Piazza d’Armi, 19 – 09123 Cagliari, Italy Abstract Reactor bioleaching of metal sulphide minerals, although much more environmentfriendly than pyrometallurgical processing, has until now only replaced the latter in the industrial practice of pretreatment of refractory gold-bearing complex metal sulphide concentrates. One of the main reasons for this situation are the investment and operating costs deriving from some limitations intrinsic of the bioreactors currently employed in biohydrometallurgy. In view of overcoming this drawback, a prototype of revolving drum bioreactor for batch operation was designed and developed by the Biohydrometallurgy Laboratory of the Geoengineering and Environmental Technologies Department of the University of Cagliari in the last decade of past century. The encouraging performance of this machine justified a programme aimed at the development of a multi-stage continuously operating machine. This paper reports on its construction and operation details and on its performance in the continuous bioleaching of a gold-bearing arsenopyrite/pyrite flotation concentrate [specific gravity about 5 g.cm-3] that has been bioleached in the conventional stirred tank reactors of a commercial plant during the past 20 years. This concentrate was selected in order to make a comparison of the performance of the revolving drum bioreactor with that of the stirred tank reactors employed in the commercial plant, where, however, only a partial leaching is required. The machine can completely bioleach as much as 4 grams of concentrate per cubic decimeter per hour out of a 40% solids pulp hence its performance is better than that of the STR’s where the highest acceptable solids concentration of the pulp is 20%. The power requirement for mixing, and keeping homogeneous with the required atmospheric oxygen transfer, a 40% solids suspension of the above-mentioned concentrate in the Biorotor is considerably lower than that required in the STR’s and the microflora is subjected to practically no shear stresses. The operation of the machine is very simple, considerably insensitive to throughput fluctuations and can be easily adapted to changing feed conditions.
Keywords: bioleaching, revolving drum bioreactor, complex sulfides concentrates, Acidithiobacillus ferrooxidans, Leptospirillum ferrooxidans
( )
* Corresponding author
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Bioleaching Applications
1.
INTRODUCTION Bioleaching of metal sulphides flotation concentrates has not, up to now, met with much success as an alternative to pyrometallurgical metal extraction processes except in some specific cases, like the pre-treatment of complex gold-bearing metal sulphides such as pyrite and arsenic-bearing sulphides. Among the main reasons for this are the relatively slow biooxidation and bioleaching kinetics and the limitations of the bioreactors where the process is carried out. The development of bioreactors that fully exploit the potential of biohydrometallurgical processing is one research field that deserves attention. This paper reports on the latest developments in research pursued over the last decade at the Biohydrometallurgy Laboratory of the Geoengineering and Environmental Technologies Department of the University of Cagliari (DIGITALB).
2.
REACTORS USED IN CURRENT BIOLEACHING PRACTICE The relative density of mineral sulphides, ranging from 4 to 6 g/cc, seriously impairs the performance of the two types of bioreactors currently adopted for mineral flotation concentrates bioleaching, i.e. the Stirred Tank Reactor (STR) and the Air Lift Reactor (ALR) represented by the so-called Pachuca Tank [Rossi, 1990]. Due to the hydrodynamics of these reactors energetic agitation is required in order to ensure the homogeneous suspension and mixing of such relatively high-density particulate solids that is generated by the stirrer in the STR and by the air flow and suspension circulation in the Pachuca tank. In addition, the oxidation process is catalyzed by aerobic chemolithoautotrophic microor-ganisms, implying adequate aeration of the mineral suspension. In the STR’s, aeration is achieved by injecting an air flow through spargers located beneath the stirrer. When the air flow rate exceeds a certain limit, depending on tank and stirrer geometry as well as stirrer rotation speed, part of the injected air is no longer dissolved in the suspension and simply rises as a bubble column escaping through the top. This phenomenon, called "flooding", [Rushton, and Bimbinet, 1960; Warmoeskerken and Smith, 1985] sets a limit on the flow rate of air that can be dissolved in the suspension. A similar phenomenon occurs in the Pachuca’s [Chisti, 1989]. One parameter that typically characterizes the ability of any reactor to dissolve atmospheric oxygen into the water filling the tank is the kLa (oxygen mass transfer coefficient). The higher the value of this parameter the greater the amount of oxygen transferred from the atmospheric air to the water. This parameter may be affected by the physico-chemical conditions of the fluid contained in the reactor tank. The influence of the ions of several elements, particulate solids, dissolved chemical compounds, some of which lower its value significantly, is well documented [Liu et al., 1989; Lee et al., 1982; Ogut and Hatch, 1988; Cieszkovski and Dylag, 1988]. In reactor bioleaching mediated by aerobic microorganisms the kLa plays a particularly important role, oxidation and leaching process kinetics depending directly on the size of the microbial population which, in turn, is a function of its growth kinetics. Among the factors that, for a given throughput, significantly affect the profitability of a commercial biohydrometallurgical operation the following should be taken into account: (i) the number and size of reactors; (ii) power requirements for mixing the solids suspension; (iii) power requirements for aeration. 430
Bioleaching Applications
The number of reactors is related to the residence time in each one. The size of the reactors depends on the solids concentration in the suspension and the power requirements are related to the particle size and density of the particulate solids as well as the solids concentration and atmospheric oxygen requirements. For the STR’s and the Pachuca tanks the interrelationships between the above factors have been exhaustively investigated for a variety of practical operating conditions. Several mathematical models, developed by different workers are utilized for tank and plant design [Bailey and Ollis, 1986; Chisti, 1989; Rossi, 2001]. However, the STR’s and Pachuca tanks, that biohydrometallurgy has borrowed from chemical engineering and hydrometallurgy, probably do not represent the best option for the specificity of bioleaching processes, where the systems consist of three phase suspensions and the microbial population plays a fundamental role. One of the limitations of the STR’s, well documented by the reports on bioleaching plant practice, is the solids concentration of the suspension, defined as the percent ratio of the mass of solids contained in a given suspension volume to the mass of the latter. All the reports available in the literature indicate that this solids concentration never exceeds 20% as shown by Table 1. Evidence has been provided that solids concentration imposes a limit on the kLa [Liu et al., 1989; Lee et al., 1982; Ogut and Hatch, 1988; Cieszkovski and Dylag, 1988; Chisti, 1989], and this fact in itself raises some doubts as to the suitability of STR’s. However, a likely more significant drawback of the STR concerns the effect of the energetic agitation required for mixing and aeration and associated interparticle abrasion [Nienow and Conti, 1978] on the microbial population. Already in the early nineties Ragusa [1990] had demonstrated that strong agitation could have a detrimental effect on the microorganisms insofar as they appear to loose their bioleaching ability. The results of recent investigations [Arredondo, Garcia and Jerez, 1994; Crundwell, 1996; Escobar, Huerta and Rubio, 1997; Fowler, Holmes and Crundwell, 1999; Holmes, Fowler and Crundwell, 1999; Crundwell, Holmes and Harvey, 1996; Kinzler, et al., 2001; Crundwell, 2001] seem to provide a rational explanation for these findings. In effect, if, due to forces acting upon the microorganism - like shear stress or the abrasion caused by solids particles - the cell envelope is partially or completely torn away, it appears that the microorganism can no longer support the electron transfer required by the oxidation process, as it is unable to adhere to the mineral surface. In STR’s and in Pachuca’s the agitation required for mixing and keeping the solids suspended very likely produces such high shear stresses or strong abrasion within the suspension as to either directly damage some of the microbial cells or traumatically detach them from the capsule if they are adhering to the solid surface.
3.
DEVELOPMENT OF A NEW BIOREACTOR The above considerations, together with the experience gained during batch bioleaching tests on several minerals carried out in the laboratory, justified investigating the features of a reactor more suited to biohydrometallurgical processing. The device should (i) ensure thorough mixing of the solids suspension, irrespective of the specific gravity of the solids, minimizing shear stresses; (ii) ensure the complete and homogeneous suspension of the particulate solids; (iii) provide an adequate and readily adjustable kLa; (iv) be supplied by an easily adjustable atmospheric air flow rate. Several prototypes of a device complying with these requirements were designed and built in the 1990’s and finally a rotating drum batch bioreactor, called "Biorotor", was tested with encouraging results [Loi, Trois and Rossi, 1995]. However, batch testing did 431
Table 1. Operating parameters of some bioleaching operations Reactor
% Solids concentration
Total useful bioreactor volume, m3
Daily throughput per bioreactor unit useful volume, tonn/m3.day
Reference
P, A
STR
20
90
0.444
[van Answgen and Marais, 2001]
Sao Bento Brazil
A, P, Pr
STR
20
580
0.138
[van Answfen and Marais, 2001; Dew et al. 1997, 2]
Olympia Greece
C
STR
20
15,936 (3 moduli of 41,328 m3 each)
0.048
[van Answegen and Marais, 2001]
Amantaytau Uzbechistan
Complex sulphides
STR
20
23,376 (4 moduli of 6x974 m3 each)
0.047
[van Answgen and Marais, 2001]
Wiluna Australia
P, A, Stb
STR
20
6x470 = 2,820 m3
0.045
[van Answgen and Marais, 2001 ; Dew et al., 1997 ; Brown et al., 1994]
Ashanti Sansu Ghana
A, P, Pr, Mrc
STR
20
16,200 (3 moduli of 6x900 m3 each)
0.0444
[van Answegen and Marais; 2001; Dew et al., 1997; Nicholson et al., 1994]
Cagliari Italy
P, A
Three stage Biorotor
40
0.045 m3 (3 moduli of 0.015 m3 each)
0.051
Plant and Location
Ore minerals
Fairview South Africa
P = Pyrite; A = Arsenopyrite; Pr = Pyrrhotite; Mrc = Marcasite; Stb = Stibine; C = complex Cu, Zn, Pb, As, Fe Sulphides
Bioleaching Applications
not provide all the information about bioleaching kinetics that could be gleaned from continuous operation. Thus with the aim of identifying all the factors affecting bioleaching kinetics, a continuously operating device was designed and repeatedly tested with a view to carrying out bioleaching tests on a pilot scale. Preliminary testing demonstrated that with a three-stage system consisting of three identical cylindrical barrels the flowsheets shown in Figures 1 and 2 were suitable for the most common mineral sulfide concentrates. Configuration (a), shown in detail in Figure 1, is the most convenient for difficult to leach ores (f. i. pyrite); of the flowsheets sketched in Figure 2, (b) and (c) are suitable for ores more amenable to leaching, like sphalerite; (d) has been satisfactorily tested for minerals such as chalcopyrite, that require an intermediate grinding step for complete bioleaching. Basically, the biorotor modules are similar to those described in detail in earlier papers. A module consists of a cylindrical barrel fitted, on its inner surface, with lifters and filled with the suspension practically up to the horizontal cylinder axis; as the barrel rotates the suspension is lifted and discharged by each lifter. The continuous sequence of laminar flowing films thus produced favours a gentle, though thorough, mixing and an effective atmospheric oxygen transfer to the suspension. Mixing and kLa depend on rotation speed, as shown by Figure 3, and on the angle formed by the lifters with the barrel radius. The design features of the new prototype used for continuous operation differ in certain respects from the batch device described in earlier papers [Loi, Trois and Rossi, 1995 and 1997]: the angle of the lifters has been modified and each barrel has been provided with a spiral scoop type feeder. This type of feeder ensures that the suspension flows by gravity from one unit to the next, doing away with the need for pumping which, apart from the associated power costs, is highly detrimental to the microbial cells.
Figure 1. Configuration (a) of flowsheet of the three-stage biorotor with three drums arranged in series. 1 = concentrate bin; 2 = stirred tank for culture medium; 3 = screw feeder; 4 = peristaltic pump; 5 = spiral scoop-type feeder; 6 = flowmeters; 7 = valves; 8 = air compressor; 9 = air inlets; 10 = stage No. 1; 11 = stage No. 2; 12 = stage No. 3; 13 = final thickener
433
Bioleaching Applications
Figure 2. Alternative flowsheets. (b): one final module fed by the outputs of two modules operated in parallel; (c): two parallel modules fed by one initial module; (d): initial bioleaching stage followed by a regrinding stage of the thickened output solids and by two modules arranged in series. B = module; T = thickener; M = grinding mill
Figure 3. Variation of kLa versus barrel rotation speed 4.
MATERIALS AND METHODS
4.1 The concentrate The case history of the Fairview Plant, located in South Africa, is well documented [van Answegen and Marais, 2001]. This plant was the first in the world to introduce, in 1974, biohydrometallurgical processing for pre-treating complex gold-bearing arsenic sulphide minerals and has continued to operate successfully up to the present time. For this reason, it was decided to purchase 100 kilos of the concentrate processed at the Fairview Plant. The mineralogical components, as determined by X-ray diffractometry, of the concentrate are, in order of abundance, quartz, pyrite, arsenopyrite and illite. The 434
Bioleaching Applications
concentrate arrived in Cagliari in polythene bags and was very moist. It was dried in a thermostated oven at 30°C and dry ground in a ceramic ball mill to –75 µm. Although in the form of a dry powder, the concentrate was very sticky and flowed with great difficulty. For this reason, the continuous feed of the first stage posed some problems and two alternative feeding devices were tested: (i) a small silos with a screw feeder that delivered the dry concentrate to the scoop feeder box and a peristaltic pump that delivered to the same box the 9K medium in the ratio for the required solids concentration; or (ii) a 25 dm3 stirred tank reactor with no air injection, where the suspension was prepared in the required percent solids concentration and a peristaltic pump transferred the suspension from the reactor tank to the scoop feeder box. Both systems are sketched in Figure 1, but proved unsatisfactory, owing to the difficulty of adjusting the small flow rates required (in the order of a few grams of solids per minute). It was necessary to manually adjust the feeders quite frequently, so, at this stage of testing, the device can be better defined as a fed-batch reactor.
4.2 The inoculum A mixed microflora, consisting of Acidithiobacillus ferrooxidans and Leptospirillum ferrooxidans strains, isolated from the drainage of the complex sulphide ores mine of Fenice Capanne, Tuscany, Italy, and bearing the conventional name "FC", was adapted to the Fairview concentrate in an STR; adaptation was very slow, requiring fifteen transfers. 4.3 The drums The geometry was the same for the three drums: 300 mm inner diameter and 540 mm in length. In each drum 12 equally spaced lifters were installed. The rotation speed was in the range from 1.05 rad s-1 to 1.36 rad s-1. The temperature of the suspension ranged from 32°C to 35°C. 4.4 Monitoring pH and Eh of pulp samples were determined daily using a conventional potentiometer. The solids were separated from the liquor by centrifugation. Ferrous and ferric iron in the leach liquors were determined by the α-α’dypiridil method; the solid phases (feed and bioleaching residues) were investigated by X-ray diffractometry and their composition was determined by quantitative chemical analysis. Power consumption was also recorded daily. 5.
RESULTS OF CONTINUOUS OPERATION TESTS Although the performance of the batch device was found to be quite encouraging, continuous operation at a suitable pilot scale, using a mineral concentrate, possibly well characterized and currently used in a commercial operation of proven performance, was considered the best procedure for properly evaluating the advantages of this bioreactor with respect to the STR. Notwithstanding the difficulties in regular feeding mentioned above, the results obtained were quite encouraging and considered worthy of reporting. The solids concentration was easily maintained at 40%, and, for a total useful volume of 45 dm3 and a feed rate of 4 grams per dm3 per hour, complete bioleaching was achieved, the solids in the output of the final stage consisting of quartz and illite with only traces of iron and arsenic.
435
Bioleaching Applications
1,50
600
1,40
550
1,30
500
1,20
450
1,10
400
1,00
350
Eh [mV]
pH
The last row of Table 1 shows that the performance parameters of the three-stage Biorotor are apparently poorer than those of only two plants out of seven. It should be borne in mind, however, that no published data specifying the extent of the partial bioleaching the minerals had undergone in those plants were available. The composition of the liquor effluent from the final stage, once the steady state had been achieved, was rather unusual and particularly interesting: total iron concentrations were about 50 g.dm-3, pH around 0.9 and Eh usually lower than 550 mV. Iron seems to be associated to an organic compound possibly an EPS. Considerable difficulties are being encountered in the chemical analysis of the liquors flowing out of the drums. Analytical determinations carried out on the same samples with different techniques exhibit systematic deviations. Therefore the existence of competing equilibria in the solutions, such as, for instance, the formation of high stability constants complexes with organic compounds, cannot be excluded. The results reported here should be considered valid insofar as they indicate a trend but their absolute values have to be considered with caution. Further investigations are under way and the results will be reported in a forthcoming paper. Figure 4 shows the pH and Eh plots of the liquors flowing out of the three drums arranged in series and in steady state:
300
0,90 0
1
2
3
4
Drum number pH, day "A"
pH, day "B"
pH, day "C"
pH, day "D"
Eh, day "A"
Eh, day "B"
Eh, day "C"
Eh, day "D"
Figure 4. Plots of pH and Eh values of liquors flowing out of the three drums arranged in series in steady state. Each couple of lines refers to one day sampling In agreement with the observations reported by other researchers [Breed, Dempers and Hansford, 1999], disruption of the process caused by irregular or inadequate aeration, without interruption of mixing, slows down kinetics considerably, to the extent that the level of microbial growth is reduced. It may be expedient to empty the device and recommence the process with a fresh suspension and a new inoculum of suitably adapted microorganisms. Therefore it is recommended to maintain a well-adapted and uncontaminated microbial population as standby in a conventional STR, to be used when necessary as new inoculum. The total power requirement was about 1000 kWh per tonne of feed. 436
Bioleaching Applications
6.
COMMENTS AND CONCLUSIONS The three-stage biorotor was found to perform quite satisfactorily and its operating characteristics appear to be competitive with the conventional reactors currently used in commercial biohydrometallurgical operations. Tests showed the Achille’s heel of the process to be the regularity of the feed: with a regular feed the already very good performance is likely to improve significantly, widening the gap between the Biorotor and conventional reactors. The fact that even though complete bioleaching of the mineral had been achieved ferrous iron is still present in the third stage, thus confirming the redox potential of the liquor, is worth mentioning and can probably be explained by the presence of arsenic. The high iron concentration in the effluent liquor and the absence of iron oxyhydroxides in the solid residue may appear to contradict the well-established principles concerning the physical chemistry of aqueous ferric and ferrous iron. However, it may be hypothesized that, owing to the almost shearless mixing that occurs in revolving drums, the EPS of the microbial cells is not damaged during their activity and can fully perform its function [Gehrke et al., 1997; 1998; Tributsch, 2001]. The iron ions may remain entrapped by the EPS free floating in the liquor after cell lysis. Analytical chemistry investigations, currently being conducted, are expected to elucidate this point thus enabling the authors to provide a more complete chemical description of the processes occurring in the different flowsheets proposed. The figure of total power requirement may be misleading: in effect, it should be taken into account that the mechanical losses become very significant in pilot scale models; no power data but only percent cost analyses have been found in the papers reporting on commercial plants and this did not enable the authors to make any comparison. However, power seems to account for more than 60% of overall operating costs of STR’s and consists of the power required for mixing in the STR’s and for the blowers that provide the required aeration [Brown, Irvine and Odd, 1994]. Calculations, carried out with the well-documented design procedures [Rossi, 2001] and applied to the above-mentioned plants, seem to provide slightly higher values.
ACKNOWLEDGEMENTS The Authors wish to express their appreciation to Dr. Cristina Trois, of the University of Natal, Durban, South Africa, for her kind assistance with the purchase and shipment of the Fairview Mine concentrate, to Prof. Salvatore Pretti and Ms. Rosa Porcu for carrying out the X-ray determinations, to Miss Cristina Licheri and Miss Marzia Fantauzzi for the analytical work carried out in the framework of their graduate theses and to their tutor, Prof. Antonella Rossi. They also wish to thank the Management of the Fairview Mine for granting the permission to purchase the concentrate. This work was carried out with the financial support of the Italian Ministry for Education, Universities and Research in the framework of the Research Project of National Relevance "Unconventional Bioreactors". REFERENCES 1. Bailey, A.D., and Ollis, D.F., Biochemical Engineering Fundamentals, McGraw-Hill, New York, (1986). 2. Breed, A.W., Dempers, C.J.N., and Hansford, G.S., South African Institute of Mining and Metallurgy, (1999) 23. 437
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3. Brown, Allan R.G., Irvine, W., and Odd, Paul, A.R., Biomine ’94, Australian Mineral Foundation, Glenside, SA, paper XVI, (1994) 16.1. 4. Chisti, M.Y., Airlift Bioreactors, Elsevier Applied Science, Amsterdam, (1989). 5. Ciezskowski, J. and Dylag, M., Mixing, (1988) 421. 6. Crundwell, F.K., Minerals Engineering, 9(10) (1996) 1081. 7. Crundwell, F.K., International Biohydrometallurgy Symposium, IBS-2001, Ciminelli, V.S.T. and Garcia Jr., O. (Eds.), Vol. A, Elsevier Amsterdam, (2001) 149. 8. Crundwell, F.K., Holmes, P.R., and Harvey, P.I., Electrochemical Proceedings, 96(6) (1996), 239. 9. Escobar, B., Huerta, G., and Rubio, J., World Journal of Microbiology and Biotechnology, 13 (1997) 593. 10. Fowler, T.A. Holmes, P.R., Crundwell, F.K., Appl. Environ. Microbiol., 65(7) (1999) 2987. 11. Gehrke, T., Hallmann, R., Kinzler, K., and Sand, W., Appl. Environ. Microbiol., 65 (1997) 159. 12. Gehrke, T., Telegdi, J., Thierry, D., and Sand, W., Appl. Environ. Microbiol., 64, 7 (1998) 2743. 13. Holmes, P.R. Fowler, T.A.Crundwell, F.K., J. Electrochem. Soc., 146(8) (1999) 2906. 14. Kinzler, K., Gehrke, T., Telegdi, J., and Sand, W., International Biohydrometallurgy Symposium, IBS-2001.BIOHYDROMETALLURGY "Fundamentals, Technologies and Sustainable Development", Ciminelli, V.S.T. and Garcia Jr. (Eds.), Vol. A, Elsevier, Amsterdam, (2001) 191. 15. Lee, J.C., Ali, S.S., and Tasakorn, P., Fourth European Conference on Mixing, BHRA Fluid Engineering Centre, Cranfield, England, (1982) 399. 16. Liu, M.S., Branion, R.M.R. and Duncan, D.W., Proceedings of the International Symposium Warwick, 1987, Norris, P. and Kelly, Don., P. (Eds.), Science and Technology Letters, Kew, Surrey, U.K., (1989) 375. 17. Loi, G., Trois, P. and Rossi, G., Vargas, T., Jerez, C.A., and Wiertz, J.Y., (Eds.), Biohydrometallurgical Processing, Vol. 1, The University of Chile, Santiago, (1995) 253. 18. Loi, G., Trois, P., and Rossi, G., Bioreactor/process Fluid Dynamics, BHR Group, (1997) 183. 19. Nicholson HM, Smith GR, Stewart RJ, Kock FW, and Marais HJ, Biomine ’94, Australian Mineral Foundation: Glenside, South Australia,. (1994). 2-1. 20. Nienow, A.W., and Conti, R., Chem. Eng. Sci., 33 (1978) 1077. 21. Ogut, A., and Hatch, R., The Can. J. Chem. Eng., 66 (1988) 79. 22. Ragusa, S., Ph. D. Thesis, The University of New South Wales, School of Biological Technologies, Sydney, Australia, (1990). 23. Rossi, G., Biohydrometallurgy, McGraw-Hill Book Company, Hamburg, (1990). 24. Rossi, G., Hydrometallurgy, 59 2-3 (2001) 217. 25. Rushton, J.H., and Bimbinet, J.-J., Can. J: Chem. Eng., 46 (1960) 16. 26. Tributsch, H., Hydrometallurgy, 59 (2-3) (2001) 177. 27. van Answegen, P.C., and Marais, H.J., Mineral Biotechnology, Kawatra, S.K. and Natarajan, K.A. (Eds.), SME, Littleton, Colorado, USA, (2001) 121. 28. Warmoeskerken, M.M.C.G. and Smith, J.M., Chem. Eng. Sci., 40 (1985) 2063.
438
15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Use of biosurfactants for the mineral surfaces modification Zygmunt Sadowski, Irena Maliszewska, Izabela Polowczyk Wroclaw University of Technology, Wybrzeze Wyspianskiego 27, 50-370 Wroclaw, Poland
[email protected] Abstract Modification of surface properties of various minerals can be a key for the mineral separation. Biomodification of the mineral properties was realised by the adsorption of biosurfactants, which were produced by Pseudomonas aeruginosa. In this study, it is shown that Pseudomonas can grow in the presence of minerals and produce a biosurfactant with substantially changes of the surface tension of supernatant. Measurements confirmed that an interaction of all used minerals with the supernatant caused a decrease of zeta potential. As expected, the ieps of mineral particles were shifted to lower pH values after the interaction with biosurfactant. Bio-pretreatment of the minerals has affected on the settling properties of mineral suspension. The settling results showed a strong stability for hematite and kaolin suspensions at the present of the high biosurfactant concentration (10% v/v).
Keywords: biosurfactant, adsorption, mineral suspension, stability, settling, zeta potential, Pseudomonas aeruginosa, kaolin, hematite, dolomite, chalk 1.
INTRODUCTION The problem of solid surface bio-modification is still an open issue despite some attempts to solve it. During the past decade the mineral beneficiation realised by chemoautotrophic bacteria is the most widely applied process for copper and gold recovery (Somasundaran et al., 2000 and Sharma et al., 2001). Biosurfactants are mainly produced in aqueous media from the carbon sources by growing microorganisms. Their use has been restricted to specific application. Commercially they are used as emulsifier reagent for hydrocarbon (Bognolo, 1999). Many microorganisms produce effective biosurfactants which reduce the interface tension between oil and brine to less than 0.01 mN/m. Biosurfactants are easily biodegradable and are particularly suited for bioremediation of oil dispersion. The biosurfactants affect the rate of hydrocarbon biodegradation in two ways: by increasing solubilization and by changing the affinity between microbial cells and hydrocarbon (Zang and Miller, 1995). Removal of entrapped organic liquid (hexadecane) can be enhanced by the use of biosurfactants. The in situ biodegradation of entrapped contaminants by rhamnolipid biosurfactant was investigated (Bai et al., 1997 and Herman et al., 1997). A detailed understanding of the biosurfactant role in modification of the mineral surface is currently lacking. The aim of the work described in this paper is to investigate 439
Bioleaching Applications
how biosurfactants produced by Pseudomonas aeruginosa affect the behaviour of mineral suspensions.
2.
MATERIALS AND METHODS
2.1 Microorganism The bacterial strain used in our experiments was Pseudomonas aeruginosa isolated from the soil samples. Cells were grown in 250 ml Erlenmeyer flasks containing 50 ml of liquid medium. Growth experiments were carried out with the medium consisted: 20 g/l mannitole; 0.05 M NH4NO3; 0.03 M KH2PO4; 0.04 M Na2HPO4.7H2O; 8 10-4 M MgSO4; 7 10-6 M CaCl2; 4 10-6 M Na2EDTA; 10 mg/l FeSO4.7 H2O. The chemicals were used as received without further purification. The strain was cultured in a rotary shaker (100 rpm) at the room temperature. The samples were taken every day, centrifuged (4000rpm at 10 min) and the supernatants were used for surface tension measurements. The biosurfactants synthesized by P. aeruginosa are most probably a mixture of rhamnolipids and glycolipids. The mineral sample (2 g) was added to the medium before sterilisation. The number of living cells in the cultures was determined by the standard colony counting method. 2.2 Minerals In this study, the pure mineral samples of hematite, kaolin, dolomite and chalk (calcite) were used. Kaolin was supplied by Surmin-Kaolin mine (AKW Group) (Poland). The average particle diameter was 1.1 µm. Hematite was purchased from Ward’s Natural Science Rochester, NY. (USA) and was ground to the size -40µm in a laboratory mill. Dolomite and chalk powders with the particle size specifications given as 90 w% < 40 µm were supplied by the Department of Geology University of Wroclaw (Poland). 2.3 Surface tension measurements Surface tension measurements were carried out by the ring method with a K10T tensiometer (Kruss, Germany). Surface tension measurements are a common tool to monitor the growth of microbial culture. Each value represents the mean of five measurements. All glassware was cleaned in chromic acid and washed in Mili-Q water. 2.4 Adsorption measurements In the experiment for the biosurfactant adsorption, 2 g of mineral was added to the biosurfactant solution. The concentration of biosurfactant solution was changed from 1 to 0.1 of an initial concentration. After 12 hours equilibration, the surface tension of supernatant was measured. From the difference of the surface tensions between the initial solution and the equilibrium solution, the adsorption has been calculated. 2.5 Zeta potential measurements Electrophoretic measurements were carried out with a particle micro-electrophoresis NicompTM 380 ZLS apparatus (Santa Barbara, California, USA). Measurements were made for diluted suspensions, obtained by adding a small quantity of mineral particles to the solution. The ionic strength of dilute suspensions was maintained at 10-3 M using NaCl. The samples were ultrasonicated for 2 min before measurements.
440
Bioleaching Applications
2.6 Sedimentation experiments The mineral suspensions were prepared by adding 2-gram mineral samples to the Andreasen pipete. Agitation and pH conditions were the same as for the biosurfactant free suspensions. Sedimentation measurements were performed using an Andreasen pipette. The stability of mineral suspension was calculated from the relationship:
Stability (% ) =
Mi − M f Mi
100
Mi - initial concentration of solid (t=0) Mf - concentration of solid after 3, 5, 10 and 15 min.
3.
RESULTS AND DISCUSSION The production of the biosurfactants was associated with the cell growth. For biomodification purposes, the mineral particles were inoculated with the bacterium. The growth curves were obtained for Peudomonas aeruginosa with out or the presence of various minerals. The relationships between the cell quantity and time are presented in Figure 1 for four minerals. 1.0E+11
Bacterial count [cells/ml]
1.0E+10
1.0E+9
1.0E+8 whitout mineral hematite kaolin
1.0E+7
chalk dolomite
1.0E+6 0.00
50.00
100.00
150.00
Time [h]
200.00
250.00
Figure 1. Growth curves of Pseudomonas aeruginosa with and without of mineral particles The surface tension changes of supernatants during the microbial growth for various minerals are shown in Figure 2. Mineral particles may attain an electrical charge, depending upon the pH of aqueous suspensions and the concentration of ions. In the presence of biosurfactant molecules some changes in the electrical double layer should be expected. Figure 3 presents the zeta potential data, which were collected during the growth of microorganism cells for the investigated minerals.
441
Bioleaching Applications
70 whitout mineral pH=6.9-7.2 hematite pH=6.1-6.4 kaolin pH=7.0-7.6
60
chalk pH=7.6-8.0
Surface tension [mN/m]
dolomite pH=7.8-8.3
50
40
30
20 0.0
50.0
100.0
150.0
200.0
Time [h]
250.0
Figure 2. The surface tension changes of supernatants during the microbial growth 40.0 hematite pH=6.1-6.4 kaolin pH=7.0-7.6 chalc pH=7.6-8.0 dolomite pH=7.8-8.3
Zeta potential [mV]
20.0
0.0 0.0
50.0
100.0
150.0
Time [h]
200.0
250.0
-20.0
-40.0
Figure 3. Zeta potentials of mineral particles as a function of microbial growth time As it can be seen, the mineral particles started with positive potential. Then, the positive potential steadily decreased. The zeta potential reversal was observed at the 9th day. Zeta potential provides an effective measurement of the potential at the solid-solution interface. The zeta potential of mineral particles was measured to determine the effect of the biosurfactant on the mineral surface charge density. The isoelectric point (iep) of the mineral is determined as the condition under, which the zeta potential is equal zero. The isoelectric point is an important characteristic of a solid-liquid interface. 442
Bioleaching Applications
Table 1. Isoelectric points of minerals pH of isoelectric point with bacteria without bacteria 4.5 8.3 5.2 7.0 8.0 8.5 9.5 12.5
Mineral Hematite Kaolin Chalk Dolomite
The effect of biosurfactant addition on the zeta potential of the minerals was different among the examined minerals. After an interaction with the supernatant, the iep of the minerals was shifted to a lower pH value. For hematite the iep was shifted to pH 4.5 after interaction with biosurfactant. A small shift of iep was observed for the chalk particles. Similar behaviour of minerals has been observed by Deo and Natarajan (Deo et al., 1998). 24.0 hematite pH=6.1-6.4 kaolin pH=7.0-7.6 chalc pH=7.6-8.0
20.0
dolomite pH=7.8-8.3
∆γ
16.0
12.0
8.0
4.0 0.0
10.0
20.0
γ −γ aq
30.0
40.0
eq
Figure 4. The adsorption isotherm of biosurfactant on the four mineral samples To observe the variation in adsorbed amounts with the biosurfactant concentration, the adsorption experiments were carried out. The results shown in Figure 4 reveal that the adsorption (the surface tension different, ∆γ = γi - γeq) with increasing the biosurfactant concentration (γaq – γeq). The sequence of adsorbed amount in all four minerals is given below: Kaolin > Hematite > Chalk > Dolomite. The effect of the biosurfactant addition on the stability of fines is shown in Figure 5. Generally, a number of coagulation mechanisms including charge neutralisation, bridging and hydrophobic interactions can be used to the explanation. These results suggest that bio-surfactants can be utilised as an effective reagent to stabilise as well as to destabilise of mineral suspension. It can be seen that lower amount of biosurfactant are needed to reach the fast rate of destabilisation of chalk, kaolin and dolomite suspensions. At the higher dosage of biosurfactant (10% v/v), the mineral suspensions are become more stabile. It is observed that high stabilisation occurs for both hematite and kaolin suspensions. 443
Bioleaching Applications
100.0
100.0 Kaolin pH=7.0-7.6 without broth
80.0
10% of broth
60.0
Stability [%]
Stability [%]
80.0
Hematite pH=6.1-6.4
40.0
without broth
2% of broth
60.0
10% of broth 2% of broth
40.0 20.0
0.0
20.0 0.0
4.0
8.0
12.0
16.0
0.0
Time [min]
4.0
8.0
12.0
100.0 Chalk pH=7.6-8.0
Dolomite pH=7.8-8.3
without broth
without broth
80.0
80.0
10% of broth
10% of broth 2% of broth
2% of broth
60.0
Stability [%]
Stability [%]
16.0
Time [min]
100.0
40.0
60.0
40.0
20.0
20.0
0.0
0.0 0.0
4.0
8.0
Time [min]
12.0
16.0
0.0
4.0
8.0
Time [min.]
12.0
Figure 5. Stability of mineral suspension at the presence of biosurfactant The interaction energy between two identical particles depends on the zeta potential and retarded Hamaker constant. Zeta potential value of about (plus or minus) 40 mV assures an energy barrier that prevents fast coagulation (Kosmulski, 2001). As seen from Fig. 4 the adsorption of biosurfactant onto hematite and kaolin was bigger than the adsorption onto chalk and dolomite. The results of zeta potential measurements with minerals in broth solutions clearly indicated the value about -40 mV for both hematite and kaolin particles (Fig. 3).
4.
CONCLUSION In this study, the attention was focused on the stability of mineral particles suspended in the various biosurfactants concentration solutions. There are two primary conclusions that can be drawn from this work. Substantial changes occurred in the zeta potential of mineral particles after their long contact with the culture broth. Observations of the mineral suspension stability after the biopretreatment demonstrate that biosurfactants are able to stabilize of hematite and kaolin suspensions more or less effectively.
444
16.0
Bioleaching Applications
ACKNOWLEDGEMENTS The financial support of this work was provided by the National Science Committee (KBN) Poland through the grant no. 5 T12B 003 22. The authors also thank Prof. A.Sokolowski for his valuable comments and the surface tension measurements. REFERENCES 1. Bai, G., Brusseau, L.M., Miller, R.M., (1997), Influence of a Ramnolipid biosurfactant on the transport of bacteria through a sandy soil, Applied Environmental Microbiology, 63, pp. 1866-1873. 2. Bognolo, G., (1999), Biosurfactants as emulsifying agents for hydrocarbons, Colloids Surfaces, 152, pp.41-52. 3. Deo, N., Natarajan K.A., (1998), Studies on interaction of Paenicillus polymyxa with iron ore minerals in relation to beneficiation, Int. J. Miner. Process., 55, pp 41-60. 4. Herman, D.C., Zhang, Y., Miller, R.M., (1997), Rhamnolipid (Biosurfactant), Effects on cell aggregation and biodegradation of residual hexadecane undwe saturated flow conditions, Applied Environmental Microbiology, 63, pp. 3622-3627. 5. Kosmulski, M., (2001), Chemical properties of material surfaces, Marcel Dekker, Inc. New York-Basel, pp. 223-282. 6. Sharma, P.K., Rao, Hanumantha, K., Forssberg, K.S.E., Natarajan, K.A., (2001), Surface chemical characterisation of Paenibacillus polymyxa before and after adapting to sulfide minerals, Int. J. Miner. Process., 62, pp. 3-25., 7. Somasundaran, P., Deo. N., Natrajan, K.A., (2000), Utility of bioreagents in mineral processing, Minerals Metallurgical Processing, 17, pp.112-116. 8. Zhang, Y., Miller, M.R., (1995), Effect of Rhamnolipid (Biosurfactant) structure on solubilization and biodegradation of n-alkanes, Applied Environmental Microbiology, 61, pp. 2247-2251.
445
C HAPTER 2 Bioremediation Environmental Applications
15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
A new bench scale restoration method for a mercury-polluted soil with a mercury resistant Acidithiobacillus Ferrooxidans strain SUG 2-2 Atsunori Negishia, Terunobu Maedab, Fumiaki Takeuchic, Kazuo Kamimurad and Tsuyoshi Sugiod a
Technical Research Institute, Hazama Corporation, 515-1 Nishimukai, Karima, Tsukuba 305-0822 b Civil Chemical Engineering Corporation, 3411 Sanuki-machi Ryugasaki 301-0033 c Administration Center for Environmental Science and Technology, Okayama University, Tsushima Naka 1-1-1, Okayama 700-8530 d Graduate School of Natural Science and Technology, Science and Technology for Energy Conversion, Okayama University, Tsushima Naka 1-1-1, Okayama 700-8530, Japan
Abstract A mercury resistant Acidithiobacillus ferrooxidans SUG 2-2 can reduce mercuric ion (Hg2+) with ferrous iron as an electron donor under acidic conditions to give volatilizable metallic mercury (Hg°). A bench scale restoration method for a mercury-polluted soil was studied with strain SUG 2-2 in order to obtain basic information for the removal of mercury from a mass of mercury-polluted soils by A. ferrooxidans. 10 kg of a mercurypolluted soil that contains 1000 mg of mercury was placed inside a cylinder made of vinyl chloride resin (30 by 60 cm). SUG 2-2 cells (104 mg of protein) suspended in 500 ml of diluted sulfuric acid (pH 2.5) and 500 ml of diluted sulfuric acid (pH 2.5) containing 5 g of ferrous sulfate were added to 10 kg of mercury-polluted soil. The cylinder was rotated once per min by a motor and the mercury volatilized from the polluted soil was trapped. 500 ml of diluted sulfuric acid (pH 2.5) containing 5 g of ferrous sulfate was added once every week into the cylinder to supply electrons to reduce mercury. The total mercury in 10 kg soil after 4 week’s operation at 30°C was 300 mg, indicating that 70% of the total mercury in the original mercury-polluted soil was removed by the treatment. This restoration method using mercury resistant A. ferrooxidans cells for mercury-polluted soil is promising because there was no need to consume much heat energy to volatilize metallic mercury from the polluted soil (above 360) and a comparatively large amount of mercury was removed within a shorter operation time. Optimum temperature and concentration of resting cells of SUG 2-2 to volatilize Hg° from mercury-polluted soil are also described. 1. INTRODUCTION Mercury and organo-mercurial compounds are highly toxic for almost all organisms because they have a strong affinity for thiol groups in proteins (1, 2). It has been known 449
Bioremediation Environmental Applications
that bacteria that are resistant to Hg2+ and/or organo-mercurial compounds have the ability to volatilize metal mercury (Hg°) from inorganic and organic mercurial compounds (1, 35). A wide range of Gram-negative and Gram-positive bacteria has mercury reductases that reduce Hg2+ with NADPH as an electron donor to give Hg (2, 6-9). Mercury reductase activity has been found in A. ferrooxidans cells (10-13). The genes involved in the volatilization of mercury have been cloned and characterized in detail (10, 14-18). Recently, we isolated a mercury-resistant A. ferrooxidans strain SUG 2-2 and clarified that in addition to NADPH-dependent mercury reductase activity, strain SUG 2-2 exhibits a novel Fe2+-dependent mercury volatilization activity (19). Cytochrome c oxidase purified from A. ferrooxidans SUG 2-2 cells catalyzed the reduction of Hg2+ with Fe2+ (20). A. ferrooxidans is an obligate acidophile and has an activity to volatilize mercury under acidic conditions. The bacterium is an obligate chemolithotroph and does not require organic compounds as an energy or carbon source for cell growth and cell maintenance. Moreover, A. ferrooxidans, in general, is resistant to heavy metals except the heavy metals compounds mercury chloride, silver nitrate (21, 22), sodium molybdate (23), and sodium tungstate (24). Considering these unique properties, a A. ferrooxidans strain resistant to mercury will be useful for the removal of mercury from mercury-polluted soil especially under acidic conditions, in which low concentrations of organic compounds, but high concentrations of heavy metals are present. We have shown that A. ferrooxidans SUG 2-2 cells volatilized Hg° from 1 g of mercuric chloride-polluted soil in water acidified with sulfuric acid (25). In this work, we develop a bench scale restoration method for a mercury-polluted soil using mercury resistant A. ferrooxidans SUG 2-2 cells. 2.
MATERIALS AND METHODS
2.1 Microorganisms, medium, and growth conditions The iron-oxidizing bacteria used in this study were A. ferrooxidans strain SUG 2-2 (19). These bacteria were cultivated at 30°C under aerobic conditions in ferrous iron medium (pH 2.5) containing 30 g FeSO4.7H2O, 3 g (NH4)2SO4, 0.5 g K2HPO4, 0.5 g MgSO4.7H2O, 0.1 g KCl, and 0.01 g Ca(NO3)2 per liter. Resting cells were prepared as follows. A ferrooxidans strain SUG 2-2 was grown in 70 liter Fe2+ medium (pH 2.5) under aeration for a week. The culture medium was filtered with Toyo no. 2 filter paper to remove the bulk of the ferric precipitates and then centrifuged using a Hitachi 18PR-52 continuous-flow rotor at 15,000 × g and a flow rate of 200 ml/min. Harvested cells were washed three times with 0.1 M β–alanine-SO42- buffer (pH 3.0), and two times with 1.6 mM sulfuric acid (pH 2.5), and used as resting cells in this study. 2.2 Analysis of mercury volatilized from mercury-polluted soil 50-ml culture flask with a screw cap contained 19 ml 1.6 mM sulfuric acid containing cell suspension of A. ferrooxidans SUG 2-2 (0.01 mg) and mercury-polluted soil (1 g). A small test tube containing 2 ml KMnO4 solution was inserted in the 50-ml reaction flask to trap the Hg° volatilized from the reaction mixture. The KMnO4 solution used (100 ml) was composed of 10 ml of a solution containing 0.6 g of KMnO4, 5 ml of concentrated H2SO4, and 85 ml deionized water. After the reaction mixture was shaken at 30°C and 100 rpm, the concentration of Hg° trapped in the KMnO4 solution was measured by cold-vapor atomic absorption spectroscopy.
450
Bioremediation Environmental Applications
2.3 Protein content Protein was measured by the biuret method (26), using crystalline bovine serum albumin as the standard. 3.
RESULTS AND DISCUSSION
3.1 Volatilization of mercury from 1 g of mercury-polluted soil by resting cells of A. ferrooxidans SUG 2-2 To obtain the basic information on mercury volatilization from a large mass of mercury-polluted soil using resting cells of A. ferrooxidans SUG 2-2, optimal conditions for mercury volatilization were determined using 1 g of mercuric chloride-polluted soil (1.5 µg Hg/g soil). We have shown that the optimum pH for mercury volatilization was 2.5 when 1 g of mercuric chloride-polluted soil containing 7.5 nmol of Hg was incubated for 10 days in 20 ml of water acidified with sulfuric acid (pH 2.5) containing resting cells of SUG 2-2 (0.01 mg of protein) and 3% ferrous sulfate (25). Under the conditions described above, approximately 4.1 nmol of mercury was volatilized. However, mercury was not volatilized from the reaction mixture without resting cells, or with the resting cells boiled for 10 min (data not shown). In these control experiments, the amount of Fe2+ oxidized for 10 days of incubation was quite low compared with that done with active cells. The amount of mercury volatilized from the mercuric chloride-polluted soil depended on the concentration of Fe2+. The maximum mercury volatilization was obtained when 3% ferrous sulfate was added to 20 ml of water acidified with sulfuric acid (pH 2.5) containing resting cells (0.01 mg) (25). Addition of a large amount of 5% ferrous sulfate inhibited cell’s activity to volatilize mercury. The optimum temperature for mercury volatilization was 30°C when 1 g of mercuric chloride-polluted soil was incubated for 10 days in 10 ml of acidic water (pH 2.5) containing resting cells (0.01 mg of protein) and 3% ferrous sulfate (Fig. 1). The effect of temperature on the mercury volatilization from the mercuric chloride-polluted soil was also studied in 20 ml salt solution (pH 2.5). The same optimum temperature and a similar level of mercury volatilization were obtained. The salt solution used for this purpose contained 3 g (NH4)2SO4, 0.5 g K2HPO4, 0.5 g MgSO4.7H2O, 0.1 g KCl and 0.01 g Ca(NO3)2 per liter of distilled water. The temperature which gave the maximum mercury volatilization activity corresponded well with the optimum pH for iron oxidation by this bacterium. The amount of mercury volatilized from 1 g of mercuric chloride-polluted soil depended on the concentration of resting cells incubated in 20 ml of acidic water (pH 2.5) (Fig. 2). The biggest mercury volatilization was obtained when 0.01 mg of cell protein was added to the reaction mixture. Addition of ten times more cells (0.1 mg of protein) to the reaction mixture decreased the amount of mercury volatilized. Although we cannot explain the reason why resting cells more than 0.01 mg decreased the amount of mercury volatilization, we are speculating the reason as follow. A rapid oxidation of Fe2+ in the reaction mixture was observed in the early step of incubation when 0.1 mg of protein was added to the reaction mixture, compared with those using 0.001 and 0.01 mg of protein because of high iron-oxidizing activity of the cells. As a result, the Fe2+ required for the reduction of Hg2+ was rapidly oxidized at the early step of the incubation without reducing Hg2+.
451
Bioremediation Environmental Applications
Figure 1. Effects of temperature on the volatilization of mercury from mercurypolluted soil by resting cells of A. ferrooxidans SUG 2-2 ■, salt solution (pH 2.5); □, water acidified with sulfuric acid (pH 2.5)
Figure 2. Effects of concentration of resting cells on the volatilization of mercury from 1 g of mercuric chloride-polluted soil by resting cells of A. ferrooxidans SUG 22 Symbols: mercury volatilization (●, 0.001 mg of protein; ■ , 0.01 mg of protein; ▲ , 0.1 mg of protein) : concentration of Fe2+ in the reaction mixture(○, 0.001 mg of protein; □ , 0.01 mg of protein; , 0.1 mg of protein) 452
Bioremediation Environmental Applications
The amount of mercury volatilized from mercury-polluted cells increased in proportion to that of mercury-polluted soil added to the reaction mixture (Fig. 3).
Figure 3. Effects of soil concentration on the volatilization of mercury from mercuric chloride-polluted soil by resting cells of A. ferrooxidans SUG 2-2 Symbols: ●, 0.1 g; ▲, 0.5 g ; ■, 1.0 g ; ▼, 5 g;
, 10 g of mercuric chloride-polluted soil
3.2 Bench scale volatilization of mercury from 10 kg of mercury-polluted soil with resting cells of A. ferrooxidans SUG 2-2 The result that mercury in 10 g of mercury-polluted soil did not inhibit cell’s activity to volatilize Hg from the soil supports an application of this mercury removal method to a large mass of mercury-polluted soil. Therefore, a bench scale restoration method for a mercury-polluted soil was studied with strain SUG 2-2 in order to obtain basic information for the removal of mercury from a large mass of mercury-polluted soils. 10 kg of a mercury-polluted soil which contains 1000 mg of mercury was placed inside a cylinder made of vinyl chloride resin (300 mm in diameter and 600 mm in length) (Fig. 4A and 4B). The cylinder flanged on both side was rotated once per minute by a motor. The air in the cylinder was circulated by an air pump through holes made on both sides of the cylinder. An installation to trap the mercury volatilized from the soil was placed between the pump and the cylinder. The apparatus described above was operated for 4 weeks in a closed system at 30°C. A. ferrooxidans SUG 2-2 cells (104 mg of protein) suspended in 500 ml of water acidified with sulfuric acid (pH 2.5) and 500 ml of water acidified with sulfuric acid (pH 2.5) containing 5 g ferrous sulfate was added to 10 kg of mercury-polluted soil. 5 g ferrous sulfate dissolved in 500 ml of diluted sulfuric acid (pH 2.5) was added once every week into the cylinder to supply ferrous iron to reduce mercury. The total mercury in 10 kg of mercury-polluted soil was 300 mg after 4 weeks of operation, indicating that 70% of the total mercury in the original mercury-polluted soil was removed by the treatment with resting cells of A. ferrooxidans SUG 2-2.
453
Bioremediation Environmental Applications
Figure 4A. Apparatus for restoration of mercury-polluted soil
Figure 4B. Photograph of apparatus for restoration of mercury-polluted soil An application of mercury resistant A. ferrooxidans cells to mercury removal from mercury-polluted soil is useful especially for the removal of Hg under acidic conditions compared with heterotrophic and neutorophilic mercury reducing bacteria. In this work, we developed a bench scale restoration method for a mercury-polluted soil using a mercury resistant A. ferrooxidans SUG 2-2 cells and show that the bacterium could volatilized 70% of total mercury in the 10 kg of mercury-polluted soil (1000 mg Hg/10 kg soil) by incubating four weeks under acidic conditions. This restoration method for mercury-polluted soil is promising because there is no need to consume much heat energy to volatilize mercury as a metal (above 360°C) from mercury-polluted soil, and a comparatively large amount of mercury can be removed for a short operation time. REFERENCES 1. 2. 3. 4. 454
J. B. Robinson, and O. H. Tuovinen, Microbiol. Rev., 48 (1984) 95. A. Velasco, .P. Acebo and F. Flores, Extremophiles, 3 (1999) 35. A. O. Summer, and S. Silver, Ann. Rev. Microbiol., 32 (1978) 637. S. Silver and T. K. Misra, Ann. Rev. Microbiol., 42 (1988) 717.
Bioremediation Environmental Applications
5. S. Silver and M. Walderhaug, Microbiol. Rev., 56 (1992) 195. 6. S. Silver and L. T. Phung, Ann. Rev. Microbiol., 50 (1996) 753. 7. J. Schottel, A. Mandal, D. Clark, S. Silver, S., and R. W. Hedges, Nature, 251 (1974) 335. 8. J. L. Schottel, J. Biol. Chem., 253 (1978) 4341. 9. K. Babich, M. Engle, J. S. Skinner and R. A. Laddaga, Can. J. Microbiol., 37 (1991) 624. 10. G. J. Olson, W. P. Iverson and F. E. Brinckman, Current Microbiol., 5 (1981) 115. 11. G. J. Olson, F. D. Porter, J. Rubinstein and S. Silver, J. Bacteriol.151, (1982) 1230. 12. J. E. Booth and J. W. Williams, J. Gen. Microbiol., 130 (1984) 725. 13. F. Takeuchi, K. Iwahori, K. Kamimura and T. Sugio. J. Biosci, Bioeng., 88 (1999) 387. 14. D. G. Rawlings and T. Kusano, Microbiol. Rev., 58 (1994) 39. 15. T. Kusano, G. Ji, C. Inoue and S. Silver, J. Bacteriol., 172, (1990) 2688. 16. C. Inoue, K. Sugawara, T. Shiratori, T.Kusano and Y. Kitagawa, Gene, 84 (1989) 47. 17. C. Inoue, K. Sugawara and T. Kusano, Gene, 96 (1990) 115. 18. C. Inoue, T., Kusano and M. Silver, Biosci. Biotech. Biochem., 60 (1996) 1289. 19. K. Iwahori, F. Takeuchi, K. Kamimura and T. Sugio, Appl. Environ., Microbiol., 66 (2000) 3823. 20. T. Sugio, K. Iwahori, F. Takeuchi, A. Negishi, T. Maeda and K. Kamimura, J. Biosci. Bioeng., 92 (2001) 44. 21. T. Sugio, T. Tano and K. Imai, Agric. Biol. Chem., 45 (1981) 2037. 22. K. Imai, T. Sugio, T. Tsuchida and T. Tano, Agric. Biol. Chem., 39 (1975) 1349. 23. K. Y. Ng, M. Oshima, R. C. Blake and T. Sugio, Biosci. Biotechnol. Biochem., 61 (1997) 1523. 24. T. Sugio, H. Kuwano, A. Negishi, T. Maeda, F. Takeuchi and K. Kamimura, Biosci. Biotechnol. Biochem., 65 (2001) 555. 25. F. Takeuchi, K. Iwahori, K. Kamimura, A. Negishi, T. Maeda and T. Sugio, Biosci. Biotechnol. Biochem., 65 (2001) 1981. 26. E. Layne, Methods Enzymol., 3 (1957) 447.
455
15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
A novel type of microbial metal mobilization: cyanogenic bacteria and fungi solubilize metals as cyanide complexes Helmut Brandl, Marion Stagars and Mohammad A. Faramarzi University of Zurich, Institute of Environmental Sciences, Winterthurerstrasse 190, CH-8057 Zurich, Switzerland (
[email protected]) Abstract A bacterial strain (Chromobacterium violaceum) as well as two fungal species (Pleurotus ostreatus, Boletus satanas) were cultivated under cyanide-forming conditions (glycine-rich growth media) in the presence of metal-containing solids such as nickel powder, platinum wire, automobile catalytic converter, or electronic scrap. All three organisms were able to mobilize nickel from fine-grained nickel powder. Copper and platinum cyanide were detected during the treatment of spent automobile catalytic converters by C. violaceum. From highly complex metal-containing wastes (e.g. electronic scrap), however, metal mobilization as cyanide complexes by cyanogenic microorganisms was hardly detectable in spite of the high metal content. The reason is probably the presence of compounds consuming or adsorbing cyanide. Keywords: HCN, metal cyanides, cyanogenic microrganisms, Chromobacterium 1.
INTRODUCTION It is known that in the presence of cyanide nearly all transitions metals (except lanthanides and actinides) form well-defined complexes (1). These complexes often show very good water solubility and are characterized by a high chemical stability. Besides, it is also well established that a variety of cyanogenic bacteria and fungi can produce hydrocyanic acid (HCN) (2, 3). However, until today, a combination of this chemical knowledge with microbiological principles regarding metal solubilization and the formation of water-soluble cyanide complexes has been considered only marginally: Besides a few reports on the biological solublization of gold by C. violaceum (4-6) the ability of cyanogenic microorganisms to mobilize metals from solid materials has not been investigated. A variety of microorganisms are able to solubilize metals from solids by redox processes, by the formation of organic and inorganic acids, and by metal complexation due to the formation of metabolic intermediates (7, 8) Their physiological abilities are used on an industrial scale for the biological extraction and recovery of metals from metalcontaining ores, a process usually termed bioleaching or biomining (7, 9). Microbially formed sulfuric acid is the main inorganic acid found in these leaching environments. It is formed by autotrophic sulfur-oxidizing microorganisms such as Acidithiobacilli. In addition, a series of organic acids are formed by bacterial (as well as fungal) metabolism 457
Bioremediation Environmental Applications
resulting in organic acidolysis, complex and chelate formation (10). No commercial process has been developed so far to mobilize metals by heterotrophic microorganisms. HCN is formed by a number of bacteria, fungi, cyanobacteria, and algae. An overview is given in Table 1. HCN does not appear to have a role in primary metabolism and is generally considered a secondary metabolite. Cyanogenesis has an ecological role and may offer the producer a selective advantage (11). Besides, a variety of plants is also known to form cyanogenic glucosides (12). Table 1. Selected microorganisms known to form cyanide (n.d., not determined, because cyanide has not been determined quantitatively, although formation is known) Maximum cyanide concentration (µM)
Reference
Chromobacterium violaceum
1900
(13)
Chromobacterium violaceum
297
(14), this work
Pseudomonas aeruginosa
220
(15)
Pseudomonas fluorescens
250
(16)
Pseudomonas pyocyaneus
4
(17)
Rhizobium leguminosarum
n.d.
(18)
Agaricus sp.
n.d.
(19)
Amantia sp.
n.d.
(19)
Boletus sp.
n.d.
(19)
Boletus satanas
102
(14), this work
Clitocybe geotropa
n.d.
(20)
Collybia maculata
n.d.
(21)
Lepiota sp.
n.d.
(19)
Marasmius oreades
n.d.
(22)
Neurospora crassa
19
(23)
Pholiota aurea
n.d.
(20)
Pleurotus sp.
n.d.
(19)
Pleurotus ostreatus
429
(14), this work
Anacystis nidulans
17
(24)
Plectonema boyanum
0.1
(24)
Chlorella vulgaris
155
(23)
Group Bacteria
Fungi
Cyanobacteria Algae
Organism
We are presenting the first report on the biological formation of metal cyanide complexes other than gold. The objectives of the project were (i) the quantitative determination of HCN formed by bacteria (C. violaceum) and fungi (Pleurotus ostreatus, Boletus satanas); and (ii) the microbiological treatment of metal-containing solids and the recovery of water soluble metal cyanides (e.g. as nickel complex).
458
Bioremediation Environmental Applications
2.
MATERIAL AND METHODS
2.1 Organisms and culture conditions C. violaceum was obtained from the German Collection of Microorganisms and Cell Cultures (DSMZ), Braunschweig, Germany (strain DSMZ 30191). Cells were cultured in a complex medium containing (in g/l) L-glutamate (4.4); KH2PO4 (1.4); Na2HPO4.7H2O (2.1); MgSO4.7H2O (0.2); FeCl3.6H2O (0.005); glycine (0.75) and L-methionine (1.5) under sterile conditions in baffled Erlenmeyer flasks and incubated at 30°C on a rotary shaker at 150 rpm. Long-term storage was carried out in 15% glycerol at -80°C. Bacterial growth was monitored by the absorbance at 600 nm. P. ostreatus and B. satanas were kindly supplied by Ivan Travnicek (Inst. of Plant Biology, University of Zurich, Switzerland). Organisms were cultivated in a defined medium containing (in g/l) maltextract (5); yeast-extract (15); D-glucose (15) and glycine (0.15) under sterile conditions in baffled Erlenmeyer flasks and incubated at 30°C on a rotary shaker at 150 rpm. Growth was monitored by measuring pH and by determining cell dry weight. For leaching experiments, different amounts (1 to 10 g/l) of solid materials (nickel powder, platinum wire, automobile catalytic converters, powdered electronic scrap) were added to the medium. 2.2 Chemical analyses Cyanide was analyzed colorimetrically at 601 nm using the pyridine barbituric acid colorimetric method (25). Metal analyses were performed by reversed phase high-pressure liquid chromatography (rP-HPLC). Metal-complexed cyanides were separated at 40°C on a hydrophobic C-18 column. The eluent consisted of 60 mM tetrabutylammonium hydroxide (TBAOH); 150 mM phosphoric acid; 25% acetonitrile; 2.34 mM sodium perchlorate·H2O. Flow rate was adjusted to 0.9 ml/min. Metal cyanides were measured by UV detection according to the extinction maxima: Fe, Cu, Ag, Au and Pt at 230 nm, Ni at 267 nm. 3.
RESULTS AND DISCUSSION
3.1 Cyanide formation The presence of nickel in the growth media (1 g/l) reduced growth (measured as optical density) of C. violaceum to a certain extent (Fig. 1a, b). Growth in the absence of nickel was accompanied by an increase in pH from 4 to 8 between 20 and 30 hours (Fig. 1). Cyanide was mainly formed in the late exponential phase (7.7 mg/l) and showed a slight decrease during the stationary phase which is probably due to degradation, adsorption, or the presence of cyanide consuming compounds in the growth medium (26). As shown in Fig. 1, the presence of nickel showed no influence on total cyanide formation. The concentration curves are identical. Fungal species (P. ostreatus, B. satanas) formed cyanide in the same order of magnitude as compared to C. violaceum. B. satanas was less efficient than P. ostreatus (Fig. 2). Also here, total cyanide decreased significantly during stationary growth phase. Growth was not influenced by the presence of nickel (1 g/l) as determined by dry biomass (data not shown).
459
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Figure 1. Growth and formation of cyanide by C. violaceum in the absence (a; open symbols) or presence (b; solid symbols) on powdered elemental nickel (1 g/l added to the growth medium). Points represent mean values of duplicate samples 12
total cyanide (mg/l)
10 8 6 4 2 0 0
2
4
6
8
time (d)
Figure 2. Formation of cyanide by P. ostreatus (triangles) and B. satanas (diamonds) in the absence (open symbols) or presence (solid symbols) of powdered elemental nickel (1 g/l added to the growth medium). Points represent mean values of duplicate samples 3.2 Metal solubilization as cyanide complexes The HPLC method developed allowed the separation and detection of several metal cyanide complexes. In general, C. violaceum mobilized more nickel from nickel powder as compared to P. ostreatus or B. satanas under the same conditions. Bacterial nickel solubilization proceeded steadily to reach a maximum level of approximately 850 µg/l after 4 days of incubation before decreasing. In comparison, P. ostreatus and B. satanas reached the maximum level of nickel-complexed cyanide of approximately 550 µg/l already after 1 to 2 days of incubation (Fig. 3). Substantial formation of nickel cyanide was observed even after 4 hours.
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Figure 3. Formation of nickel cyanide from powdered elemental nickel (1 g/l added to the growth medium). z C. violaceum; P. ostreatus; B. satanas. Points represent mean values of duplicate samples Total nickel mobilization was in the range of 1% of the total solid material applied for all experiments, which is rather low and not very efficient. However, results show that microbial metal mobilization as cyanide complexes is principally feasible. Using spent automobile catalytic converters as solid material, C. violaceum was able to mobilize copper and platinum. In preliminary experiments, platinum cyanide as well as copper cyanide was detected after an incubation of 29 hours (Fig. 4). In contrast, only copper cyanide was found in samples of P. ostreatus cultures.
Figure 4. Chromatographic separation of metal cyanides formed by C. violaceum grown on spent automobile catalytic converter Dust-like residues from the mechanical recycling of used electronic equipment (e.g. computers) represent a highly complex metal-containing matrix (27). However, metal mobilization as cyanide complexes by cyanogenic microorganisms was poorly detectable in spite of the high metal content. The reason is probably the presence of compounds consuming or adsorbing cyanide. 4.
CONCLUSIONS Over all, the results represent a novel type of microbial metal mobilization based on the ability of certain microbes to form HCN. The findings might have the potential for a microbially based industrial application regarding the treatment of metal-containing solids or the biological remediation of metal-polluted soils since metal cyanides can be separated 461
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by chromatographic means and easily be recovered by sorption onto activated charcoal. However, cyanogenic microbes have to be comprehensively evaluated to fully exploit the potential to form water-soluble metal complexes from solid metal-containing solids. Growth and cyanide formation have to be optimized as well as metal mobilizing efficiencies. The presence of cyanide-consuming compounds in the growth medium has to be considered especially. Typically, cyanide is formed during growth only during a short time period (early stationary phase) (28). A significant cyanide formation can be obtained only in certain growth media under specific conditions (29). Iron is known to stimulate cyanide formation by gram-negative bacteria (30). Magnesium stimulates the formation by gram-positive bacteria, copper and zinc the formation by fungi. There is a huge potential to find novel cyanogens, which can be industrially applied. It is known that in soils – and especially in the plant rhizosphere – cyanogenic microorganisms represent a large part (50%!) of the soil microbial community (31). Pseudomonas strains (e.g. P. aeruginosa, P. fluorescens) seem to be particularly suited. Work on these organisms is in progress. ACKNOWLEDGEMENTS We thank Christina Kägi and Tobias Rosenberger for the excellent technical assistance. REFERENCES 1. 2. 3. 4.
B.M. Chadwick and A.G. Sharpe, Adv. Inorg. Chem. Radiochem. 8 (1966) 83. M. Greshoff, Pharmaceutisch Weekblad 46 (1909) 1418. B.J. Clawson, C.C. Young, J. Biol. Chem. 15 (1913) 419. S.C. Campbell, G.J. Olson, T.R. Clark and G. McFeters, J. Ind. Microbiol. Biotechnol. 26 (2001) 134. 5. E.N. Lawson, M. Barkhuizen and D.W. Dew, In: Biohydrometallurgy & the Environment Toward the Mining of the 21st Century, Vol. 9A. R. Amils and A. Ballester (eds), Elsevier, Amsterdam, 1999; pp. 239. 6. A.D. Smith and R.J. Hunt, J. Chem. Tech. Biotechnol. 35B (1985) 110. 7. H. Brandl, In: Biotechnology, Vol. 10, Special processes. H.J. Rehm and G. Reed (eds.), Wiley-VCH, Weinheim, 2001, pp. 191. 8. M. Ledin and K. Pedersen, Earth Sci. Rev. 41 (1996) 67. 9. K. Bosecker, FEMS Microbiol. Rev. 20 (1997) 591. 10. J. Berthelin, In: Microbial Geochemistry. W.E. Krumbein (ed.), Blackwell, Oxford, 1983, pp. 223. 11. C. Blumer and D. Haas, Arch. Microbiol. 173 (2000) 170. 12. A. Nahrstedt, N. Erb and H.D. Zinsmeister, In: Cyanide in Biology. B. Vennesland, E.E. Conn, C.J. Knowles, J. Westley, F. Wissing (eds.), Academic Press, London, 1981, pp. 461. 13. D.F. Niven, P.A. Collins and C.J. Knowles, J. Gen. Microbiol. 90 (1975) 271. 14. M. Stagars, Masters thesis, Univ. of Zurich, Zurich, Switzerland 2001. 15. P.A. Castric, Can. J. Microbiol. 21 (1975) 613. 16. R. Askeland and S.M. Morrison, Appl. Environ. Microbiol. 45 (1983) 1802. 17. H. Lorck, Physiol. Plant. 1 (1948) 142. 18. H. Antoun, C.J. Beauchamp, N. Goussard, R. Chabot and R. Lalande, Plant Soil 204 (2000) 57. 462
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19. M. Locquin, Bull. Mens. Soc. Linn. Lyon 12 (1943) 151. 20. E. Bach, Physiol. Plant. 1 (1948) 387. 21. R. Singer, The Agaricales in Modern Taxonomy, Koltz Scientific Books, Konigstein, 1986. 22. J.B. Lebeau, E. Hawn, J. Phytopathol. 53 (1963) 1395. 23. E.K. Pistorius, H.S. Gewitz, H. Voss and B. Vennesland, Biochim. Biophys. Acta 481 (1977) 384. 24. B. Vennesland, E.K. Pistorius, H.S. Gewitz, In: Cyanide in Biology. B. Vennesland, E.E. Conn, C.J. Knowles, J. Westley and F. Wissing (eds), Academic Press, London, 1981, pp. 349. 25. Anonymous, Deutsche Industrienorm, DIN 38 405 1981. 26. P.A. Fagan, Dissertation, Univ. of Tasmania, Sidney, Australia, 1998. 27. H. Brandl, R. Bosshard and M. Wegmann, Hydrometallurgy 59 (2001) 319. 28. C.J. Knowles and A.W. Bunch, (1986) Microbial cyanide metabolism. Adv. Microb. Physiol. 27 (1986) 73. 29. R. Michaels and W.A. Corpe, Cyanide formation by Chromobacterium violaceum. J. Bacteriol. 89 (1965) 106. 30. D.G.Kleid, W.J. Kohr W.J. and F.R. Thibodeau, Processes to recover and reconcentrate gold from its ores. US patent 5,378,737, 1995. 31. R.J. Kremer and T. Souissi, Cyanide production by rhizobacteria and potential for suppression of weed seedling growth. Curr. Microbiol. 43 (2001) 182.
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
An approach to cyanide degradation in wastewater of gold ore processing Valentina Podolskaa, Zoya Ulberga, Nicolai Pertsovb, Ludmila Yakubenkoa and Beishen Imanakunovc a
F. Ovcharenko Institute for Biocolloidal Chemistry, National Academy of Sciences of Ukraine, 42 Vernadsky blvd., 03142 Kyiv, Ukraine b Russian Research Centre of Molecular Diagnostics and Therapy 8 Simpheropolsky blvd., 113149 Moscow, Russia c Institute of Chemistry and Chemical Technology NAS of Kyrghyzstan 267 Chuy Av., 720071 Bishkek, Kyrghyzstan Abstract This investigation deals with the development of a complex biotechnology of cyanide degradation in effluents and pulp, containing free cyanide and cyano-metal complexes. The complex approach includes three types of treatment. They are: the microbial degradation, the electrostimulation and the sorption. The microbial strain Pseudomonas fluorescens B5040 isolated from the slime of a tail pond is resistant to a high concentration of cyanide. The strain has shown ability to destruct free cyanides and complex cyanides containing copper, silver, nickel, zinc, cadmium to ammonia and carbon dioxide. The biosuspension treatment under dc field action allows the 30-50% reduction of the process duration due to cyanide complexes activation and the microbial cells electrostimulation. The sorption cleaning at the process completion allows to avoid the recalcitrant cyanide discharge to environment. Simultaneous use of the microbiological and physico-chemical treatment methods allow considerably reduce the treatment time, enhance purification efficiency and destruct cyanides directly in a pulp. Keywords: metal complexed cyanide, wastewater, microbial degradation, electric field 1.
INTRODUCTION Cyanation of oxidized ores and ore concentrates is now most widespread technology used for the gold extraction in an industrial scale. Sodium cyanide is an effective but nonselective solvent. Thus, during extraction many by-products from ores are retrieved in addition to precious metals. During the treatment of sulfide and oxidized minerals or after the precipitation of gold by zinc powder cementation, complexed cyanide of copper, iron, nickel, cadmium and silver can get into sewage. These are compounds with varying toxicity and stability; some of them are toxic likely sodium cyanide. In contrast to free cyanides, which can decompose under the action of external factors, complexed cyanides are more stable as to these conditions and may remain unaffected for a long period of time, later causing secondary pollution. Therefore, the development of sewage disposal 465
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technologies for gold-extracting plants requires consideration of complexed cyanides with CN- concentration usually varying from 10 to 50 ppm. In the majority of cases the existing technologies that deal with the purification of cyanide-containing sewage fail to reach the sanitary guideline. These technologies have significant shortcomings including the formation of new toxic compounds. The degradation of cyanide using microbes appears to be a more favorable approach as compared with other methods using reagents in purification for reasons of environmental safety and their low cost. An aerobic industrial process that effectively removes cyanide, metal complexed cyanide, thiocyanates and ammonia from cyanidation wastewaters at low concentration up to 10 ppm has been developed by Homestake Mine, USA. The main species of microbial consortium was Pseudomonas paucimobilis Mudlock [1]. The problem of interaction between bacteria and cyanide complexes of transition metals has a number of specific features. First of all, these features are connected with the influence of cyanides on physiologic activity and energetic parameters of cells [2]. Furthermore, the metal being released during the complexed cyanide destruction not only causes the additional contamination of effluents with heavy metal ions but also exerts a toxic effect on the degrading agent. Few papers reported the biodegradation of metal complexed cyanides by fungi [3], in activated sludge [4], and demonstrated the existence of a number of Pseudomonas species able to degrade nickel and copper complexes [5]. The present paper investigates the factors governing the microbial removal of free and complex cyanide from aqueous systems. The special emphasize was put on external electric and electromagnetic fields. The given paper gives the attempt to combine the advantage of the physico-chemical and biological methods for the cyanide destruction of cyanide-containing effluents. 2.
MATERIALS AND METHODS The Pseudomonas fluorescens RCIM B-5040 strain isolated from the slime of tail deposit of gold-extracting plant in Uzbekistan was used in this study. Its minimum inhibitory concentration values determined by WASP plate method were 7 mM (CN-) in the rich medium and were two times less (3 mM) in the medium. Bacteria have been cultivated under aerobic conditions for 18 hours in 200 ml of 5M medium (0.2% KH2PO4 + 0.1% K2HPO4 + 0.05% Na2CO3 + 0.03% MgSO4 + 0.01% NaCl + 0.2% glucose + 0.05% peptone + 0.01% NaCN) at the temperature of 28°C. Two methods were used for the determination of cyanide: distillation for the total cyanide including ferrocyanides, measurement and WAD CN method for the simple and metal complexed cyanides (copper, zinc, silver), excluding ferrocyanides, measurement. The colorimetric technique was used for the determination of the free cyanides, thus formed, using pyridine and barbituric acid. Aminoacid analysis of cultural liquid was carried out with BIOTRONIC-LC500 instrument (Japan) using ninhydrin sorbent. The experiments on an electric field influence on the cyanide removal rate by bacteria were carried out in the following way. Suspension containing P. fluorescens RCIM B5040, cyanide complex, and 5M medium was placed in vessels and incubated on the shaker. The platinum foil electrodes were introduced into the vessels and dc electric field was applied. Ag/AgCl electrode was used as a reference one. For treating operation the
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impulse potentiostat PI-50-1-1 (Russia) was used. No electric treatment was applied to the control vessels. Irradiation of cells by electromagnetic field was conducted in the following way. 0.5 ml of cell suspension concentrated by centrifugation were placed in radio-transparent plastic container and electromagnetic field was applied. The thickness of layer irradiated did not exceed 2 mm. The time of irradiation was 20 min at density power of 10 mWt. The source of coherent oscillation in the frequency interval 37.5-78.3 GHz was used. 3.
RESULTS AND DISCUSSION
3.1 Biochemical interaction between the microbial cells and metal complexed cyanides Table 1 demonstrates the kinetic relationships of destruction of cyanides being included in the composition of complex salts in the suspensions of P. fluorescens RCIM B-5040 cells. As it is seen from this data, the cyanides are most easily destructed in the solutions of Na2Zn(CN)4 salt. Cyanides to be complexed with copper and silver were destructed slower. After 48 hours of contact with bacteria the potassium ferrous-cyanide was only slightly attacked and less than 5% degradation was observed. Table 1. Kinetics of metal complexed cyanide biodegradation (mg/l) Contact time, hrs 0 4 18 28 48
K4Fe(CN)6 [CN-] [Fe2+] 4.3 10.7 4.2 10.7 4.1 10.5 4.0 10.4 3.9 10.2
NaAg(CN)2 [Ag+] [CN-] 10.7 5.2 8.2 4.3 2.5 1.7 2.1 1.0 2.2 0.4
Na3Cu(CN)4 [Cu+] [CN-] 5.5 9.25 5.2 1.9 5.0 <0.1 4.9 <0.1 4.2 <0.1
Na2Zn(CN)4 [Zn2+] [CN-] 20.8 13.4 6.4 1.4 0.9 1.4 0.5 1.0 0.29 <0.1
There was noted remarked the specific reaction of cells to the adding of the cyanosilver complex into the bacterial suspension. In four hours of contact between cells and dicyanoargentate the pH value decreased by 2 units from 7.3 to 5.2. The higher the concentration of complex anion was, the more pH value reduced. In most cases the addition of cyano-copper complex resulted in slight acidation of culture liquid following the first day of contact and in pH rise during the second day of the contact. There were no substantial variations in the pH value in the case of zinc and iron cyanide complexes. The amino-acid analysis of culture liquid being formed during the process of P. fluorescens RCIM B5040 growing on cyanide-containing substrates was also performed. The results obtained indicate the substantial differences of amino-acid composition depending on the compound added thereto, although there were some regularities of this process. It was revealed that the growth of cells on the nutrient medium with NaCN was accompanied by the accumulation in the metabolite of aminoacids of increased content of basic groups. In this case basic groups and acidic ones constituted about 51% and 6.29% of total content, respectively. The interaction of cells with Na3Cu(CN)4 and NaAg(CN)2 was accompanied by 2.5 fold increase of the content of acid and aliphatic groups, mainly, due to the presence of aspartic and glutamic amino-acids. At the same time, the content of basic groups was 5-6-fold reduced. The increase of contact time from 24 to 48 hours resulted in two-fold increase of acid group content. The metal concentration and the cyanide content in bacterial suspensions were controlled simultaneously. It has been shown that on interacting between the cell and 467
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NaAg(CN)2 the substantial reduction of silver concentration occurred correlating to that of cyanide. The ratio of [CN-] and [Ag+] was really about 2.3 during the 48-hours period of incubation. As to cuprocyanide the biodegradation of cyanide was accompanied by the accumulation of cuprous ions in solution. The reduction of copper concentration in the solution was significantly lower as compared with silver complex. The substantial reduction of zinc concentration in Na2Zn(CN)4 solution during the cells cultivation is connected with zinc accumulation by microbial cells in smaller extent but is associated with the formation slightly-soluble precipitates in greater extent. Thus, from the data obtained on the rate of cyanide destruction by P. fluorescens RCIM B5040, the metal complexed cyanides are placed in the following row: NaCN > Na2Zn(N)4 > Na3Cu(CN)4 > NaAg(CN)2 > K4Fe(CN)6. Earlier for the Pseudomonas fluorescens Harris and Knowles demonstrated that cyanide is converted to NH3 and CO2 by an enzyme system with the properties of dioxygenaze [6]. One could assume that the mechanisms of the interaction between the microbial cells and cyano-metal complexes is associated with the displacement of equilibrium of complex anion dissociation due to cyanide ligand assimilation by bacteria according to one of the mechanisms of enzyme kinetics. The rate of destruction is the higher, the lower is pH value. At pH drop the dissociation of complexes is increased and at the rise of pH the dissociation is reduced. As it was demonstrated in [7], that at pH 5 the equilibrium concentration of silver in 0.01 M solution of NaAg(CN)2 constitutes 1.5×10-6 M and at pH 7 it reduces to 6.8×10-8. This regularity is, however, true only at low concentration of complexes in the solution. With the concentration increase the hydrolysis can take place. Thus, the equilibrium of the complex dissociation reaction may be displaced towards the formation of dissociation products due to the removal of metal ions in the form of difficulty soluble hydroxide. It is noteworthy that the pH value should not be outside the limits of physiologic pH values for specific bacteria culture, which is in the range pH from 5.5 to 9.5. Physiologic response of cells to the addition of some cyanides, in particular, copper and, especially, silver cyanides into the medium reveals in pH decrease. Consequently, it displaces the equilibrium of anion dissociation and facilitates assimilation of cyanides by the cells. As follows from Table 2, the rate of microbial destruction of discussed complexes is well correlated with the equilibrium concentration of metal and ligand, calculated from instability constant Kins. The apparent discrepancy between the low rate of microbial assimilation of ferrous-cyanide complex (Table 1) and its low durability (Table 2) may be well explained by the kinetic inertness of this complex. [8]. The obtained results demonstrate that the higher is the equilibrium concentration of metal and ligand in a solution and the faster this equilibrium achieved, the higher the rate of destruction of said complex is. Table 2. Equilibrium ion concentrations (M) into 1×10-4 M solutions at pH 7 NaAg(CN)2 Na3Cu(CN)4 K4Fe(CN)6 Na2Zn(CN)4
Kins [7] 8×10-22 5×10-31 1×10-24 1.3×10-17
[Me] 6.8×10-8 2.2×10-6 1.0×10-4 1.0×10-4
[CN-] 1.4×10-7 8.8×10-6 6.0×10-4 4.0×10-4
Procedure for performing calculations is described in [7]. The accumulation in solution of excess amount of metal ions to be released after assimilation of cyanides by the cells results in the reduction of complex dissociation 468
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degree and influences the inhibiting effect on microorganisms. Metals can be removed from the contact medium by one of the below-mentioned techniques: accumulation by cells, production of stable complexes with the products of metabolism, formation of insoluble or low-soluble sediments, etc. As we have noted, the first of the above mentioned mechanisms is true for NaAg(CN)2. This specific interaction between the cell and cyanide complex of silver reduces the toxic effect of metal and rises the threshold of its toxicity for bacteria. The second mechanism is most probably realized during the contact of cells with Na3Cu(CN)4. Copper mainly remains in the solution in the form of organic complexes. The aeration process enhances the displacement of redox-potential to positive region as well as the mineralization of organic compounds and the copper removal from the solution as water insoluble sediments. Zinc is easily hydrolyzed in neutral and weakly alkaline pH region. The presence of phosphate compounds contributed to a sediment formation. Contrariwise, under the oxygen deficit conditions due to formation of metal-organic composition with the metabolites in the solution the watersoluble zinc compounds accumulated. One could come to conclusion that the intensive aeration is needed to realize the biochemical treatment of cyanide-containing solutions including the transition metal cyanides. 3.2 Electric field treatment The experiments on cyanide destruction by the microbial cells P. fluorescens RCIM B5040 in 5M medium under the dc and impulse electric field treatment were carried out. In Fig.1 shows the kinetic relationships of NaAg(CN)2 and K4Fe(CN)6 destruction, respectively. These relationships are: assimilation with bacteria (curve 2); treatment by impulse electric field (curve 1); assimilation with bacteria together with the electric field treatment (curve 3). It has been demonstrated that the highest rate of cyanide destruction was observed in the third variant where the electric stimulation of microbial destruction took place. The treatment both by the direct current field and by the impulse electric field gave closely agreed results. Some electrochemical investigations allowed to explain the nature of the discovered phenomenon. In the cyclic volt-ampere curves of NaCN there were not found peaks corresponding to direct anode oxidation of cyanides. Unlike to simple cyanide, the reversible waves corresponding to the reduction of metal complexed cyanide have been found in the cyclic volt-ampere curves in the solution of silver and copper complexed cyanides at the potential values being -0.51 V and -0.70 V (vs. Ag/AgCl), respectively (curves are not shown). The preparative electrolysis of NaAg(CN)2 in undivided electrochemical cell revealed low reduction of cyanide concentration at the potential of working electrode being displaced by 0.2 V to the anode region in relation to the potential reduction, namely at -0,3 V. The results obtained are presented in Table 3. At the potential of -0.3 V (vs. Ag/AgCl) the effect of electrochemical destruction of metal complexed cyanide under the dc field treatment in microbial suspension has been reduced to minimum. Therefore, the data given in Fig. 1 are not connected directly with the electrolysis of complex salts. It is possible to suppose that the irreversible displacement of the equilibrium of cyanide complexed anion dissociation due to cyanide removal from the reaction medium is enhanced while treating the bacterial suspension by electric field of low intensity. The application of electric field accelerates the biodestruction of weak and moderate complexes and initiates the biological destruction of the strong and inert complexes. Total effect of the microorganisms and electric field action has the synergetic character, i.e., considerably exceeds the sum of each factor action separately. This is especially 469
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noticeable for the iron cyano-complex. As it follows from data of Table 1 and Fig. 1a (curve 2), the bacteria practically do not destruct the potassium ferrous-cyanide ions. Electric field action alone for two days brings about the 24% destruction of ferrocyanides. At the same time, the dc field treatment in the presence of bacteria results in the 93% complex ferrous-cyanide destruction. Thus, the given phenomenon can be considered as the stimulation effect of the process in the microbial cyanide destruction. The electric field stimulation action was observed while applying electric field intensity excluding the electrochemical cyano-metal complex destruction.
Figure 1. Kinetics of K4Fe(CN)6 (a) and NaAg(CN)2 (b) destruction. 1 - dc field treatment; 2 – bacterial treatment; 3 - simultaneous dc field and microbial treatment. Potential applied:–0,3 V (Ag/AgCl) Table 3. Preparative electrolysis of NaAg(CN)2 in 5M medium (CN, ppm) Time, hrs 0 2 6 10 16
Undivided electrochemical cell -0.3 V -0.7 V 9.2 9.6 9.0 8.4 8.9 6.2 8.4 4.0 7.6 0.8
Divided electrochemical cell Cathode compartment Anode compartment -0.7 V -0.7 V 9.1 9.1 7.9 8.9 5.6 8.4 3.4 7.8 0.2 6.9
We have also discovered that the rate of the assimilation of NaAg(CN)2 with the bacteria to be treated by the electric field appeared to be by 25% higher than that of the assimilation with the native cells. During the subsequent multiple seeding of electrically treated bacteria on cyanide-containing media the considerable biomass increase exceeding that of control tests was observed. 3.3 Electromagnetic field treatment Low intensive magnetic field of millimeter range at the certain frequencies can influence stimulatively both the eukaryote and procaryote organisms [9]. It was noted that both the stimulation and inhibition of the physiological bacteria activity take place in narrow frequency range that indicates the resonant character of the influence [10]. The target of the given investigation was whether it is possible to intensify the microbial cyanide destruction using electromagnetic field and to decrease of cyanide toxic action on the culture. We have performed the experiments on the influence of electromagnetic 470
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irradiation (EMI) in frequency range from 37.5 to 78 GHz on the efficiency of NaAg(CN)2 microbial destruction. Suspension of P. fluorescens 5040 cells was subjected to the irradiation by the millimeter EMI in NaCN solution and was used as inoculum into medium containing NaAg(CN)2. Fig. 2 shows the data on the efficiency of NaAg(CN)2 microbial destruction depending on the electromagnetic field frequency and incubation duration. It is seen that in one day (curve 2) the microbial destruction of silver cyanide complex by the irradiated cells was approximately by 30% higher as compared with the control (curve 1), which contain non-irradiated cells. At the frequency 54 GHz the destruction was maximum and reached 53%. In two days (curve 3) the destruction by native and irradiated cells practically equalized. However, for the sample irradiated by field with 54 GHz frequency the destruction index was the highest. In three days (curve 4) deference in the destruction between the irradiated and non-irradiated cells was not observed any more. The bacteria growth within the hole interval of the irradiation as well as in control differed not more than by 8%, therefore, this index was not the indicative for the assessment of the biological action of millimeter EMI. Probably, this is explained by toxic influence of sodium dicyanoargentate on the culture, stipulated by silver. Irradiation of inoculum showed the initial stimulating influence on the cyanodestructing activity of P. fluorescens B5040 strain but not on its growth. However, the frequency region close to 54 GHz, that corresponded to the highest stimulated influence was observed. Addition investigations have allowed to suppose that the mechanism of the millimeter irradiation action associated with its influence on the processes running in the membrane of a bacterial cell. The matter is that simultaneously with the increase in the destructive activity the increase the hydrolytic activity of the membrane ATPhase and the cell surface charge were detected (data are not given) [11].
Figure 2. The kinetics of the NaAg(CN)2 microbial destruction as a function of the electromagnetic field frequency. The cell inoculum was irradiated before seeding Because the EMI stimulating effect is limited in time it could be used in two ways: (i) under the continuous irradiation condition of microbial suspension by the millimeter EMI in the range 50-60 GHz, (ii) in periodic mode but with less time, 20 min. Besides, the specificity of the millimeter irradiation is associated with the EMI absorption by water. This operation demands the treatment in suspension thin layer, that makes its applicability enough complicated in practice for the time being.
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3.4 Technological approaches The performed investigations have shown the possibility for the complex approach based on the creation of highly effective strain, optimization of the physiological and biochemical indexes of biodestruction, and the intensification of these parameters under the action of physico-chemical factors. This method allows to destruct cyanides directly in a tail pulp containing free and complexed cyanides and thiocyanates. The process includes the following sequential operations: growth of the seed inoculate; move of the inoculate into bioreactors with influent, sequential increasing biomass, and simultaneous cleaning under the conditions of periodic cultivation (starting the process); biological oxidation and simultaneous procedure of the periodic treatment of the bio-mineral suspension under dc electric field action; additional purification of influents from calcitrant complexes using the sorbent and microorganisms. Table 4. Results of wastewater (pulp) biological oxidation (ppm) Total cyanides Thiocyanates Hexacyanoferrates Gold Silver Copper Iron
Influent 74.5 58.0 36.3 0.26 1.7 6.25 5.2
Effluent 0.1-0.6 0.1 1.8 0.02 0.2 0.11 1.3
The process can be carried out both periodically and continuously. To realize the process the pilot mobile plant with the production rate 20m3 per day was designed and tested at a gold mining factory. The content of the main components in an effluent and the results of its treatment in the pilot plant are given in Table 4. After the treatment the cyanide and thiocyanates content decreased up to the sanitary guidelines and the additional gold and silver extraction took place. 4.
CONCLUSIONS 1. The present study of some metal complexed cyanide demonstrated that the efficiency of cell-complex interaction is determined, on the one hand, by the degree of bacteria adaptation to cyanides and heavy metals and, on the other hand, by the form of complex salt in the solution. Such parameters as pH value, excess concentration of metal and cyanide are capable to cause the reversible displacement of the equilibrium of the anion dissociation. In the process of microbial degradation of cyanide and on treating the microbial suspension by dc electric field the irreversible displacement of equilibrium of dissociation due to the removal of cyanide from the reaction medium occurs. 2. It was revealed that the effect of cyanide destruction acceleration is not a simple combination of the processes of electrolysis and biodegradation, but a complex system of interrelated phenomena, which include the electrostimulation and electroadaptation of microorganisms to toxic compounds, displacement of dissociation equilibrium due to the removal of one of the components from the reaction medium, etc. It is well understood that the application of external electric field could cause both the intensification of the process of assimilation of cyanides by cells and the death of bacteria culture as well as the deterioration of process characteristics due to incorrectly chosen electro-treatment conditions. 472
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3. The possibility to destruct cyanides and thiocyanates up to the sanitary guidelines directly in a pulp, minimum use of reagents as well as the possibility to extract additionally gold and silver makes the developed process not only ecologically attractive but economical method for the cyanide destruction in effluents from gold-mining factories. The positive results obtained during the process test in the pilot plant indicate the possibility of the complex approach combining the microbial treatment with physicochemical actions. The research was partly supported by the ISTC kr-556 project. REFERENCES 1. R.M.R. Brannion and H.G. Ebner (Ed.), J.L. Whitlock, T.I. Mudder, Fundamental and Applied Biohydrometallurgy, Elsevier, Amsterdam, 1986. 2. L.N. Yakubenko, V.I. Podolska, V.E. Vember, V.I. Karamushka, Colloids Surfaces, A: Physicochem. Engineer. Aspects, 104 (1995) 11. 3. M. Barclay, A. Hart, C.J. Knowles, J.C.L. Meeussen, V.A. Tett, Enzyme Microb. Technol., 22 (1998) 223. 4. M.M.M. Goncales, A.F. Pinto, M. Granato, Environ. Tecnol., 19 (1998) 133. 5. C.Boucabeille, A. Boris, P.Olliver, G. Michel, Environ. Pollution, 84 (1994) 59. 6. R.E. Harris, A.W. Bunch, C.J. Knowles, Sci. Prog. Oxf., 71 (1987) 293. 7. V.E. Shpak, V.I. Podolskaya, Z.R. Ulberg, E.A. Shpak, Colloid Journal, 57 (1995) 102. 8. B.M. Chadwick, J.G. Sharpe, Advances in Inorganic Chemistry and Radiochemistry, 8 (1966) 83. 9. H. Frohlich (Ed.), Biological Coherence and Response to External Stimuli, Springer Verlag, 1988. 10. A.A. Kataev, A.A. Alexandrov, L.I. Tikhonova, G.N. Berestovsky, Biophys. (Russia), 38 (1993) 446. 11. V.I. Podolska, G.V. Ponezha, L.N. Yakubenko, T.G. Gruzina, N.I. Grishchenko, Phys. of Alive, 10 (2002) 56 (in Russian).
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Available options for the bioremediation and restoration of abandoned pyritic dredge spoils causing the death of fringing mangroves in the Niger Delta Elijah I. Ohimain Rohi Biotechnologies Ltd, 104DAirport Road, Warri, Delta State, Nigeria Abstract In the Niger Delta, estuarine sediments/soils containing pyrite are often dredged to create access for the exploration and exploitation of petroleum resources. The unconfined disposal and abandonment of the resultant sulfidic spoils along canal banks has resulted in environmental degradation principally through soil acidification, heavy metal pollution, flooding, mangrove die back, erosion, and siltation, succession to freshwater plant species, and altered topography and hydrology. Consequently, former mangrove areas have been converted to either bare spoil heaps, grassland or freshwater forest after several years of natural weathering. This paper discusses ways to avert further degradation through bioremediation, restoration and rehabilitation of the affected areas. The socio-economic consequences of exposed sulfidic spoils in the estuarine ecosystem are also discussed. Keywords: acidification, acid sulphate soils, dredging, mangrove, Niger Delta, pyrite, rehabilitation, restoration, sulfidic dredge spoils/sediment 1.
INTRODUCTION The Niger Delta, which occupies over half of the entire Nigerian coastline possesses the largest mangrove forest in Africa and is also one of the largest wetland in the world. Despite the recognized value of mangroves for shoreline protection, as nursery grounds and source of food for commercial and sports fisheries [1], the acreage of wetland habitat in many areas has been reduced by anthropogenic modifications [2]. Outstanding among these modifications is dredging, with concomitant spoil disposal, and in Niger Delta, this appears to be the most important single cause of alteration of tidal wetlands. Dredging in this estuarine ecosystem is often carried out to create safe navigable accesses for resource exploitation particularly oil and gas. During dredging, mangrove sediments and soils are removed, placed along canal banks mostly upon fringing mangroves and abandoned, thus killing the mangroves. Several hectares of mangroves fringing most of the creeks where dredging has taken place has been killed likewise. The extent of impacts on the mangroves have not been reported neither has the quantity of abandoned spoils been quantified. For instance, a company dredged about 2 ha of mangrove in order to create access for oilwell drilling, a further 2.4 ha of mangrove was killed as a result of dredge spoil dumping [3]. A major oil producing company in the delta generated approximately 20 million cubic metres of spoils between 1990 and 1996 [4]. It 475
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is expected that the amount of abandoned spoils may have increased considerably taking into account the activities of other oil companies, the nearly 50 years of such operations in the delta and the observed high sedimentation/siltation rates which often necessitates frequent maintenance dredging. The sediments and soils of the mangrove zone have been reported to contain reduced iron sulphides particularly pyrite [5]. When present in the natural anoxic and undisturbed state under mangrove cover, sedimentary pyrites are known to be innocuous, but their disturbance through dredging and spoil disposal often initiates a cascade of oxidative reactions leading to estuarine acidification. Estuarine sediment acidification and mine spoil drainages have been reported to cause the death of vegetation [6], fish/aquatic biota [7], change in water quality [8] and heavy metal pollution [9-13). Furthermore, the usual practice of placing unconfined sediments continuously along the canal bank beyond tidal inundation has led to the creation of artificial levee. In the process, several kilometers of undulating spoil heaps now characterize the once low-lying intertidal landscape. The resultant change in the topography and hydrology of the area often prevent site recolonization by native mangrove species. After several years of weathering only acid and metal tolerant plants become established particularly grasses and sedges followed by some freshwater species. Worldwide, the disposal of pyritic dredge spoils is a major challenge because of the risk of environmental degradation [10, 11, 13-15]. However, the natural microbial succession and colonization of acidic and heavy metal-laden mine spoil heaps [16-17] and dredge spoils [12] underscores their importance in the management of these wastes. Several microorganisms have been isolated from highly acidic environments such as bacteria [18], fungi [19], algae [20] and diatoms [21]. Some of these organisms have been studied for the bioremediation of metal contaminated wastes [22] through biosorption and changes in redox state. Others have considered sulphate reduction processes for the bioremediation of acidity and heavy metals [23-24]. Lovley and Coates [22] had indicated that bioremediation of metals is still at the research stage with little large-scale application. The aim of this paper therefore, is to present the results of a preliminary laboratory study on the bioremediation of acidic and metal laden dredge spoil leachates with the potential of large-scale application through restoration of site hydrology that will enhance bacterial sulphate reduction and to permit volunteer mangrove recruitment. The environmental and social impacts arising from the abandonment of pyritic spoils are also highlighted. 2.
ENVIRONMENTAL IMPACTS OF ABANDONED SULFIDIC SPOILS The abandonment of unconfined dredge spoils has led to a number of environmental impacts such as direct burial and destruction of fringing mangroves and associated fauna, change in topography and hydrology, siltation of navigable canals, flooding and suffocation of mangroves, degradation of water quality, habitat fragmentation and alteration of vegetation (i.e. conversion from mangrove to bare spoil heap and succession to grassland and freshwater vegetation). This problem is often compounded when the spoils contain sulfidic materials particularly pyrite as in the case of the Niger Delta, it could lead to severe acidification with attendant consequences including heavy metal pollution, vegetation dieback, reduced plant/animal/agricultural productivity, corrosion of steel, concrete, and other engineering structures, degradation of surface and ground water quality, mortality of estuarine biota especially fishes and bioaccumulation of pollutant [8, 25-26]. 476
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Typically, the tall red mangrove species, Rhizophora racemosa fringes the banks of the numerous creeks in the mangrove zone of the delta. During dredging, sediment, soil and vegetation along the proposed right of way (ROW) are removed, dumped over bank beyond tidal influences and then abandoned. This has led to the alteration in the topography and hydrology of the area [27]. As inundation becomes less frequent with increasing elevation, the soil chemistry and hydrology are also altered, resulting in the alteration of vegetation and organisms inhabiting the area. It has been reported that the hydrology of wetlands is highly sensitive spatially to even minor changes in topography and associated tidal regime [28-29]. With no tidal and freshwater reaching the artificial spoil levee, the salinity of the pore water will now depend on rainfall and rising ground water levels. The salinity will change depending on season; it will become concentrated during the dry season and diluted during the wet season. After several years of leaching/weathering the spoil will become relatively less saline, which tend to favour the growth of invasive species. Typically, the spoils are placed continuously as canal banks, which form barriers to water flow and thereby causing excessive flooding of the mangroves in the backswamp. Mangroves are known to be sensitive to excessive flooding, and are often killed in the process [30]. Furthermore, acidic spoil leachates often drain into the backswamp and stagnate there. The leachates from spoils in the Niger Delta have been shown to contain high levels of heavy metals [12]. High acidity, heavy metals and altered topography, hydrology and salinity regimes may have contributed to the lack of natural re-vegetation in most of the dredge spoil dumpsites. This phenomenon has led to the creation of canopy gaps and vast wasteland (devoid of vegetation) in the otherwise sheltered estuarine ecosystem and several years after dumping, grasses and freshwater swamp forest communities develop in an otherwise mangrove swamp forest. 3.
SOCIO-ECONOMIC IMPACTS The availability of suitable land for housing and farming is a major challenge in the mangrove areas of the delta. The soils are typically low lying (0.8 – 1.2 m above mean sea level) and are seasonally flooded and tidally inundated daily. The major occupation of the natives is fishing. They therefore found elevated abandoned dredge dumps attractive for the establishment of houses, fishing camps and home gardens. Fruit trees such as pawpaw (Carica papaya), mango (Magifera indica), Avocado pear (Persea Americana), Coconut (Cocos nucifera) and pineapple (Ananas comosus) are commonly cultivated on matured spoils close to human dwellings. Vegetable and other food crops such as okra (Abelmoschus esculenta), bitterleaf (Vernonia amygalina), flutted pumpkin (Telfaria ocidentalis), cassava (Manihot sp.), cocoyam (Colcasia esculenta) and plantain (Musa sp.) are also cultivated on elevated levees or degraded dredged spoil dumps [27]. Some of these crops particularly pineapples have been reported to be tolerant to acid sulphate soils and have been successfully cultivated on elevated sulfidic spoils in other countries [3132]. In the process of their occupation of abandoned dredge spoil dumps; some natives now reside dangerously close to oilfield installations. Most of the crops planted on dredge spoils with the exception of plantain and pineapples often suffer from poor yields. Beyond this, there is the risk of heavy metal toxicity and bioaccumulation. Mangrove plants have been reported to bioaccumulate heavy metals [33] so are crops grown on sulfidic dredge spoils [34-35].
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4.
OPTIONS FOR BIOREMEDIATION AND RESTORATION Soil acidification has been regarded as a major challenge to mangrove restoration worldwide [36-38]. The pH of dredge spoil leachates is often below 3.5 [12]. Such low pH has been found to inhibit mangrove seedling growth [38]. The accompanying high concentration of heavy metals may compound the problem. Ohimain [25-26] had suggested proper handling of dredge spoils to prevent acid formation. This involves the selective placement of spoils in such a way as to avoid/minimize contacts between the causative agents of acidification, namely, Acidithiobacillus, pyrite, water and air. In this paper, emphasis will be focused on the bioremediation of acidic and metal laden leachates followed by restoration of site hydrology to permit field application of bioremediation in large scale and to enhance natural mangrove recruitment. 4.1 Bioremediation In a preliminary laboratory study on the bioremediation of spoil leachates, sulphate reducing bacteria (SRB) was isolated from the bottom layer of recently dredged spoils using modified Baar’s medium [39]. The study was carried out under two different pH regimes, a set of leachates, which had a pH of 2 were sterilized using autoclave, while the other was adjusted to pH 6 using 1.0 M NaOH (microcosm 2) prior to sterilization. The leachates were fed into a 500 ml separation funnel inoculated with pure cultures of SRB under reducing conditions and incubated in the dark at a temperature of 28°C for 180 days. Samples were collected monthly from the top of the separation flask into sealed universal bottles using sterile hypodermic syringes. The samples were analyzed for sulphate (turbidimetric/colorimetric), pH, redox potential (Russell’s pH/mV equipped with platinum electrodes in combination with Ag/AgCl reference electrodes), sulphide (iodometric methods), heavy metals (Cd, Cr, Cu, Ni, Mn and Zn) (using atomic absorption spectrophotometer) and SRB population [39]. At the beginning of the experiment (day 0), leachates were characterised with high sulphate (5200 mg/l), acidic pH (microcosm 1 only), redox potential (+280mV), while sulphide was not detected. The medium became blackened after 30 days, thus indicating initiation of sulphate reduction activities. The intensity of the blackening increased thereafter up to day 90 when the medium separated into 2 distinct layers, consisting of an upper clear aqueous layer (supernatant) and a lower dark solid layer (pellets). As the experiment progressed, the supernatant became clearer while the pellets became darker. The increased blackening appears to have correlated with the population of SRB. In both microcosms, as the population of SRB increased, the level of sulphate decreased with a correspondingly increase in sulphide and pH values and a decrease in redox potential (Fig. 1). Heavy metal levels of the aqueous layer followed this pattern; it decreased rapidly from Day 30 to Day 180. Apart from the initial lag period observed within the first 30 days of the experiment in microcosm 1 (Fig. 2), the heavy metal removal efficiency was similar in both microcosms (Table 1), which suggests that the SRB was probably acid tolerant. Using acid-tolerant SRB has the added advantage of reducing the cost of alkaline pre-treatment prior to bioremediation. It was observed from correlation statistics that SRB population was inversely related to sulphate concentration and directly related sulphide concentration and pH [12], it therefore follows that the microbiological process for the bioremediation of acidic and metal laden leachates depends on the reduction of soluble sulphates to insoluble sulphides, leading to the neutralization of sulphate acidity (i.e. increase in alkalinity) with resultant heavy metal precipitation [12]. Apart from the direct formation of insoluble metal 478
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sulphides, the increased pH as a result of sulphate reduction can provide an additional mechanism for metal removal since some metals such as Zn, Co, Ni, Mn sulfides are more soluble at low pH than at neutrality or high pH [23].
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Figure 1. Changes in SRB population, sulphate, sulphide, Redox potential and pH during bioremediation Table 1. Heavy metal concentration of dredged spoil leachates and precipitation efficiency after 180 days bioremediation treatment
Copper Cadmium Chromium Nickel Manganese Zinc
Initial leachate metal concentration, mg/l 82.8 122.0 53.2 75.3 171.0 113.2
Precipitation efficiency, % Microcosm 1 Microcosm 2 91 92 90 90 94 96 98 100 99 100 98 99
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180 Initial pH 2
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Figure 2. Changes in heavy metal concentration during bioremediation using acid tolerant SRB 4.2 Mangrove restoration and rehabilitation Ohimain [25-26, 40] had suggested the restoration and rehabilitation of abandoned spoil banks to prevent environmental degradation particularly through acidification. Details of mangrove restoration and rehabilitation techniques are presented elsewhere [4043], here emphasis will be placed on the restoration of soil salinity, hydrology and topography and also provide a means of creating anaerobic conditions in the spoil banks to permit the growth of SRB for the removal of sulphate acidity and heavy metals (bioremediation), which will permit natural mangrove recruitment. Prior to spoil abandonment, the topography of the Niger Delta wetlands is 0.8 – 1.2 m above mean sea level [44]. To permit natural mangrove recruitment, the restoration of normal tidal exchange and residence time, site topography and drainage, and freshwater inputs are necessary [37]. This will require site excavation (and grading to pre-spoil disposal elevation) and back filling into disused canals [45] especially those linking dry or exhausted oilwells or other unsuccessful hydrocarbon prospects. This will obviously restore the hydrology, and as the site becomes tidally inundated once again, it will permit removal of oxidation products [37] and since the brackish water of the Niger Delta is well buffered with pH ranging from neutral to slight alkaline (7.0-8.4) [40, submitted] the acidity is expected to decline afterwards [46]. Tidal inundation will permit the soils return to anoxic condition found in natural undisturbed mudflats/mangroves [37] and encourage the growth of estuarine bacteria [47] particularly SRB [12]. White et al. [48] suggested rehabilitation of wetlands by re-flooding. Flooding is expected to reverse acidity through microbial catalyzed reactions in which sulphate is reduced to sulphide [49]. Ainodion et al. [50] had used hydrological restoration to successfully re-establish mangroves in the Niger Delta. 480
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5.
CONCLUSION The unconfined disposal of sulfidic dredge spoil is one of the major challenges of the oil industry operating in the Niger Delta. The practice of creating abandoned spoil banks in the estuarine ecosystem has caused a plethora of environmental problems, including direct destruction of fringing mangroves, acidification, heavy metal pollution and alteration in site topography, hydrology and salinity, which prevents volunteer mangrove re-colonization. Through natural weathering processes, the resultant bare spoil dumps continues to release acidic and metal laden leachates into the environment. The result of the preliminary laboratory bioremediation studies appears promising for the removal of acidity and heavy metals. The field application of bioremediation will require the restoration of the normal site hydrology, which will create the required anaerobic conditions for the removal of acidity through bacterial sulphate reduction. Hydrological restoration also has the added advantage of restoring salinity while encouraging tidal buffering and recruitment of volunteer mangrove seedlings. REFERENCES 1. A.K. Semesi and K. Howell. The Mangroves of the Eastern African Region. United Nations Environmental Programme (UNEP) Kenya, 1992. 2. J.E. Drifmeyer and W. E. Odum. Environ. Concentration. 2: (1975) 39. 3. UNICAL CONSULT, Study on the effect of drill slot on aquatic life in Cawthorne Channel. Report submitted to SPDC, Nigeria, 1994. 4. Ade Sobande and Associates, Dredging Impact Study. Report submitted to SPDC, Nigeria, 1998. 5. B. Anderson, Report on The Soils of The Niger Delta Special Area. Niger Delta Development Board, Port-Harcourt Nigeria, 1966. 6. A. Schippers, P-G, Jozsa, W. Sand, Z.M. Kovacs, M. Jelea, Geomicrobiology J. 17 (2000) 151. 7. J. Sammut, M.D. Melville, R.D. Callinan and G.C. Fraser, Aus. Geogr. Studies, 33 (1995) 89. 8. J. Sammut and R. Lines-Kelly, An Introduction to Acid Sulphate Soils. Natural Heritage Trust, Australia Seafood Industry and Environment Australia, 2000. 9. M. Astrom, Environ. Geol., 36 (1998) 219. 10. S.R. Stephens, B.J. Alloway, A. Parker, J.E. Carter and M.E. Hodson, Environ. Pollut., 114 (2001) 407. 11. M. Astrom, J. Geoc. Expl., 73 (2001) 181. 12. E.I. Ohimain, Bioremediation of heavy metal contaminated dredge spoil from a mangrove ecosystem in the Niger Delta. Ph.D. Thesis, Univ.of Benin, Nigeria, 2001. 13. P. Peltola and M. Astrom, Sci. Tot. Environ. 284 (2002) 109. 14. M.W. Clark and D.M. McConchie, In Proceedings of the 5th International acid sulphate soils conference, Australia, (2002) 6. 15. S.Y. Demas, A.M. Hall, D.S. Fanning, M.C. Rabenhorst, E.K. Dzantor, In Proceedings of the 5th International acid sulphate soils conference, Australia, (2002) 8. 16. C.L. Brierley, Hydrometallurgy 59 (2001) 249. 17. D.B. Johnson, S. Rolfe, K.B. Hallberg and E. Iversen, Environ. Microbiol. 3 (2001) 630. 18. E.I. Robbins, Hydrobiologia, 433 (2000) 61. 19. S. Gross and E. I. Robbins, Hydrobiologia, 433 (2000) 91. 20. W. Gross, Hydrobiologia, 433 (2000) 34. 21. D.M. DeNicola, Hydrobiologia, 433 (2000) 111. 481
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22. D.R. Lovley and J.D. Coates, Curr. Opinion Biotech. 8 (1997) 285. 23. C. White and G.M. Gadd, Microbiology .142 (1996) 2197. 24. B.C. Hard, S. Friedrich and W. Babel Microbiol. Res.152 (1997) 65. 25. E.I. Ohimain, A paper presented at the National Conference on Environmental Science and Technology held at Greensboro September 8–10, 2002. (2002a). 26. E.I. Ohimain, A paper presented at the DPR International Conference on Health, Safety & Environment in Oil and Gas Exploration and Production held in Abuja Nigeria, December 2002. (2002b). 27. E.I. Ohimain, A paper submitted to J. Soils Sediment (2002c). 28. C. Hughes PhD thesis University of Newcastle, Australia (1998). 29. W.J. Mitsch, and J. G. Gosselink, Wetlands. Van Nordstram, New York, 2001. 30. K. Kathiresan and B. L. Bingham, Advances in Marine Biology, 40: (2001): 81. 31. L.Q. Minh, T. P. Tuong, H. W. G. Booltink, M. E. F. van Mensvoort and J. Bouma, Agric. Wat. Manage. 32 (1997). 32. N.J. Stevenson, R.R. Lewis and P.R. Burbridge: In W. Streever (Ed.), An International Perspective on Wetland Rehabilitation. Kluwer Academic Publications, The Netherlands (1999) 277. 33. L.D. Lacerda, Trace Metals Biogeochemistry and Diffuse Pollution in Mangrove Ecosystems, ISME, Japan, 1998. 34. R.D. Delaune, and C.J. Smith, J. Environ. Qual. 14 (1985) 164. 35. R.G.V. Bramley and D.L. Rimmer, J. Soil science. 39 (1988): 469. 36. N.J. Stevenson, Coastal Management 25 (1997). 423. 37. U.L. Kaly and G.P Jones, Ambio 27 (1998): 656. 38. Van Dessel, J.P and P.S.Omoku, A paper presented at the international conference on Health, safety and Environment in Oil and Gas Exploration and productions, Jakarta, (1994) 437. 39. C.O. Obekwe and N. Okonkwo, Microbios. Letters. 21 (1982) 113. 40. E. I. Ohimain, A paper submitted to Wetlands Ecol. Manage. (2002d). 41. C.D. Field, Restoration of Mangrove Ecosystems. ITTO/ISME, Japan, 1996. 42. C.D. Field, Mar. Pollut. Bull. 37 (1998) 383. 43. C.D. Field, Hydrobiologia 413 (1999) 47. 44. P.C. Nwilo and A. Onuoha, Environmental impacts of human activities on the coastal areas of Nigeria. In: L. F. Awosika, A. C. Ibe and P. Shroader (Eds): Coastlines of West Africa. American Society of Civil Engineers, New York, 1993, 220. 45. R.E. Turner and B. Streever, Approaches to Coastal Restoration: Northern Gulf of Mexico. SPA Academic Publ. The Hague, The Netherlands 2002. 46. B. Indraratna, W.C. Glamore and G.O. Tularam, Geotech. Geol. Engr. 20 (2002) 181 47. D.M. Alongi, Hydrobiologia 285 (1994) 19. 48. White, M.D. Melville, B.P. Wilson and J. Sammut, Wetlands Ecol. Manage. 5 (1997) 55. 49. D. Dent, Acid Sulphate Soils: A Baseline for Research. ILRI Pub.39, Wageningen, the Netherlands, 1986. 50. M.J. Ainodion, C.R. Robnet, T.I. Ajose, A paper presented at the Society for Petroleum Engineers (SPE) International Conference on Safety and Environment in Oil and Gas Production held in Kuala Lumpur, Malaysia, 20-22 March 2002. SPE Paper 74033 (2002).
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Bacterial reduction of TcO4- under the haloalkaline conditions T. Khijniaka*, N.N. Medvedeva-Lyalikovaa, M. Simonoffb a
Institute of Microbiology, Russian Academy of Sciences, 7/2 Prospect 60-letiya Oktyabrya, Moscow 117811, Russia b Laboratoire de Chimie Nucleaire Analytique et Bioenvironnementale CNRS UMR 5084, Le Haut Vigneau, BP 120, 33175, Gradignan Cedex, France Abstract Among long-lived radionuclides, technetium (99Tc) is notable for its long half-life (2.13 x 105 years), its extreme mobility in the environment and its tendency for bioconcentration. Wastes from fuel reprocessing have an extremely high pH and high salinity after neutralization. There are no published studies concerning TcO4- reduction under alkaline conditions. In the present work haloalkaliphilic bacteria isolated from sodalake environments were used for reduction of TcO4- under anaerobic alkaline conditions, when acetate was served as ē-donor. After 2 month incubation portion of the reduced Tc reaches 80% (from 0.25 mM initial) and Tc(III) was detected spectrophotometrically in the medium besides Tc(V) and Tc(IV). Unlike neutral pH, under alkaline conditions reduced Tc probably formed aqueous complex TcIVO(OH)3(CO3)-. TcO4- reduction was not detected in the control experiments without bacteria or ē-donor, or in the aerobic conditions. The microbial reduction has been suggested as a potential mechanism for the removal of Tc from contaminated environments or waste streams. Keywords: haloalkaliphilic bacteria, Halomonas, reduction of pertechnetate, pH 10, Tc(VII), Tc(IV), Tc(III) 1.
INTRODUCTION The problem of radioactive wastes arises from experimental explosions of nuclear weapons, wastes from nuclear fuel cycle reprocessing plants and the use of isotopes for medical purposes. Featured are eigth elements that constitute some of the most prevalent metals and radionuclides found in Department of Energy (USA) waste: cesium, chromium, lead, mercury, plutonium, uranium, strontium and technetium. Among longlived radionuclides, technetium (99Tc) is notable for its long half-life (2.13 x 105 years), its extreme mobility in the environment and its tendency for bioconcentration. During reprocessing, Tc is solubilized from nuclear spent fuels and is present in all waste streams as the pertechnetate anion (TcVIIO4-), which is enters to environment [1]. In anoxic
* Financial and human support from the French Ministry of Education and Research, CNRS, Bordeaux Region and from PACE (Programme sur l’Aval du Cycle Electronucléaire) is greatefully acknowledged. Also, this work was supported in part by grants 02-04-48196 from the Russian Foundation for Basic Research.
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sediments, dissimilatory reduction may play an important role in directly alterating the form and mobility of technetium. However, the complex chemistry of Tc under reducing conditions required careful consideration of both aqueous and solid-phase products resulting from dissimilatory reduction, which in turn will be strongly dependent upon the electron donor and solution composition. Pertechnetate is weakly sorbed by most soils and subsurface sediments [2]. Tc can form several reduced species, including Tc(VI), Tc(V), Tc(IV) and Tc(III), which interact strongly with rocks, minerals and organic ligands [3]. But solubility of reduced technetium can be stabilized and increase in aqueous phase by complexing ligands as carbonate. Therefore, it is important to identify Tc forms accurately to predict the long-term stability and environmental mobility of microbially reduced Tc following in situ bioremediation. Microbial metal reduction results in the precipitation of a low valence, reduced, form of the element, and has therefore been proposed as a strategy to treat contaminated waters [4]. Sulfate-reducing bacteria Desulfovibrio desulfuricans [5, 6], metal-reducing bacteria Shewanella putrefaciens [7] and Geobacter sulfurreducens [8] and Escherichia coli [9] were found to be capable of reducing Tc(VII) at neutral pH. Under acidic conditions, Acidithiobacillus ferrooxidans and A. thiooxidans can also reduce Tc(VII) to low-valency forms [10]. However, there is no information about technetium reduction under alkaline conditions. Earlier, a group of alkaliphilic halotolerant heterotrophic bacteria from alkaline environments in Siberia, Kenya, Buryatiya and Mongolia was shown to oxidize sulfur compounds to tetrathionate [11, 12]. Currently, the collection of these bacteria includes about 20 strains, which according to 16S rRNA sequencing, belong to the genus Halomonas in gamma Proteobacteria [11-13]. The aims of this work were to determine whether haloalkaliphilic soda-lake isolates would reduce Tc(VII) under anaerobic haloalkaline conditions and to assess a mobility and solubility of reduced form of Tc under haloalkaline conditions. 2.
MATERIALS AND METHODS
2.1 Reagents All reagents were of analytical grade and purchased from Sigma-Aldrich. 2.2 Organisms and culture conditions Facultative anaerobic bacterial strains (Se1, Se 3, Se4, Se5, Se D) were obtained from D.Yu.Sorokin (Institute Microbiology RAS, Moscow, Russia) and were isolated from Mongolian and Kenyan soda lakes. Strain Mono, also facultative anaerobic bacterium, was isolated from Mono Lake (California, USA) by N.N.Lyalikova. The nutrient medium contained: Na2CO3 - 13 g, NaHCO3 - 4 g, NaCl - 50 g, K2HPO4 - 0.5 g, MgSO4 - 0.1 g, yeast extract - 0.1 g, NH4Cl - 0.1 g, sodium acetate - 2.8 g, 2 ml of microelement’s solution [14] per L of distilled water. The final pH of the medium was 10. All cultivations were conducted under anaerobic static conditions at 30°C. Medium was dispensed (4.5-9 ml aliquots) into 17 ml serum bottles, sealed with butyl rubber stoppers, deaerated with N2 and autoclaved. Cells aerobically pre-grown on acetate were washed twice with the same medium and aliquots (0.5-1 ml) and electron donor and acceptor (sodium acetate and sodium pertechnetate) were then transfered to serum bottles. The concentration of TcO4was 0.25 mM. Bacterial protein measurement was performed by Bradford assay [15].
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2.3 Spectrophotometric determination of different Tc species in the cultural supernatants UV-visible spectra of technetium were recorded in the digital form, using a Beckman DU-50 Series double-beam diode array UV-visible spectrophotometer in the range of 250 to 700 nm. Identification of bands was carried out according to [16] in carbonatebicarbonate medium, where the band at 290 nm is characteristic for Tc(VII), the band at 512 nm for Tc(IV) and at 630 nm for Tc(III). When Tc(III) was recorded, experimental bottles had not been opened and placed directly to spectrophotometer, in all other cases the solution had been transfered to quartz cuvettes. 2.4 Chromatographic separation and identification of Tc Samples (0.2 ml) were centrifuged with Sigma 3MK (Germany) at 5000 g for 10 min. Different Tc species were separated on Whatman 3MM chromatography paper with 0.3 N HCl [17], 0.9% NaCl [18] or 100% acetone [19] as the mobile phase. Chromatography paper was impregnated with 5-10 µl aliquots. After separation air-dried chromatograms were exposed to phosphor storage screen for 16 h. Chromatograms were visualized and quantified with the PhosphorImager technique [20]. 2.5 Extraction techniques for determination of Tc species For comparison with chromatographic Tc determination, several extraction techniques were used to differentiate between Tc species in the cultural supernatant. To co-precipitate Tc(IV) and Tc(V), FeCl3 (final concentration 10 mg ml-1) and concentrated NH4OH (3 drops) were added to the sample (0.5 ml). Precipitate was washed 3 times and dissolved with concentrated HNO3 (0.5 ml). The amounts of Tc in the washes and the precipitate were determined. To extract Tc(VII), 1 ml of tetraphenylarsonium chloride in chloroform (0.05 M) was added to an equal volume of sample [21]. After 1 min of shaking, the phases were separated. The amounts of Tc in the organic and aqueous phase were determined. To extract the Tc(V) species, 4% (w/w) 8-hydroxyquinoline in chloroform [22] was added to the supernatant. After 5-10 min of shaking, the phases were separated. The amounts of Tc in the organic and aqueous phases were again determined. In all cases the radioactivity of each phase (organic, aqueous, precipitate, washes) was determined with a Beckman liquid scintillation analyzer and Insta Gel coctail, Packard Bioscience Company. 3.
RESULTS AND DISCUSSION
3.1 Spectrophotometric analysis of cultural supernatant. Haloalkaliphilic bacteria were tested for the ability to reduce TcO4- under the alkaline anaerobic conditions. Under these conditions the carbonate-bicarbonate medium became progressivly pink (after 23 h), then colourless (10-12 days) with time as Tc(VII) was reduced by bacteria. Spectrophotometric analysis indicated the presence of Tc(IV) (welldefinied band at 512 nm) and Tc(III) (very wide band with maximum at 628 nm) in the bacterial spent media (Fig. 1b,c). The spectral data were in reasonable agreement with the reference values of Tc(IV) and Tc(III) at λmax=512 nm and λmax=630 nm, respectively [16]. No reduced species of technetium were detected in the control solutions without bacteria or with dead biomass (Fig.1a). As soon as colourless bacterial spent medium was exposed to air, the pink colour reappeared. At the same time, spectrophotometric data showed that the band at 628 nm had completely disappeared and the intensity of the band at 512 nm had slightly increased (Fig. 1d). 485
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Figure 1. Spectrophotometric studies of TcVIIO4- reduction. Spectra were recorded with strain Mono at cell protein concentration 0.035 mg/ml. Different forms of technetium could be determined as Tc(IV) at 512 nm and Tc(III) at 628 nm according to [16]. a - control experiment without bacteria; b - bacterial medium after 23 hours of exposition; c - after 12 days of exposition; d - case c exposed to air If incubation of bacteria was continued under the mentioned conditions, the pink colour disappeared again. Tc(III) is very sensitive to oxygen and immediately reoxidises to Tc(IV) [16], that’s why in our experiments this form of technetium was detected only transiently. Spectrophotometric analysis gives a quick, precise but only qualitative characteristic of reduced technetium forms. For detailed characteristic it is necessary to use other methods. 3.2 Tc(VII) reduction by haloalkaliphilic bacteria In order to determine whether metal-reducing capacity is wide spread among haloalkaliphilic bacteria isolated from different soda-lakes environments, six strains from genera Halomonas were tested for the ability to reduce TcO4- under haloalkaline anaerobic conditions. All soda-lake isolates were able to reduce technetium and the profile of the resulting Tc species was very similar (Fig.2). Reduced technetium species could be separated, identified and quantified, for example, using extraction technique or paper chromatography. The results of extraction clearly show that after 4 days of incubation 486
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under anaerobic alkaline conditions, an average of 76% of the 0.25 mM of pertechnetate had been reduced by haloalkaliphilic cultures (Se1, Se4, Se5, Se D) to Tc (IV) and Tc(V) species (Fig.2a). Note that pertechnetate ion was not reduced in the control experiments without bacteria or with dead biomass (Fig. 2a). Also, no technetium reduction was observed under the aerobic conditions. No Tc was precipitated from the medium over a 34 hours period (bacterial spent medium was centrifugated at 10000 rpm for 20 min, the technetium measurements were done before and after centrifugation). After 3 days of incubation, about 6% was co-precipitated from the medium with excess of carbonate. After 2 months, the amount of reduced technetium reached 80%: from that, 25% was precipitated and 55% remained in solution. The reduced technetium species could be differentiate from TcO4- and TcO2 with the help of chromatographic technique: multiple paper chromatography (PC) with 0.3 N HCl [17], polar 0.9% NaCl [18] and acetone [19].
Figure 2. Reduction of pertechnetate by halophilic bacteria. a – the amount of reduced Tc after 4 and 60 days of cultivation. b – the amount of pertechnetate in the bacterial medium. The data were received by methods of extraction with (a) (Ph)4AsCl and (b) coprecipitation with Fe(OH)3. D.b. – dead biomass Paper chromatography with 0.3 N HCl showed presence of only Tc(VII) (Rf = 0.7), Tc(V) (Rf = 0.0) and Tc(IV) (Rf = 0.9) in the bacterial spent medium (Fig.3, lines 1, 2, 3). In case of PC/saline (Fig.3, lines 4,5,6) technetium was found as Tc(VII) (Rf = 0.7), Tc(IV)complex (Rf = 1.0) and Tc(IV)O2 (Rf = 0.0), which shows no movement in this solvent system, whereas in case of acetone as a mobile phase all reduced technetium remained at the origin (Rf = 0) and only TcO4- was moved (Fig. 3, lines 7, 8, 9). After acetone this chromatogramm was treated for the second time with 0.9% NaCl to have another proof of 487
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technetium valency. Fig. 3 (lines 10, 11) shows that technetium, which remained at the origin in acetone system, now separated for two spots with Rf = 1.0 and Rf = 0.0. We suggested that spot with Rf = 1.0 represent an electronegative Tc(IV)-complex and spot with Rf = 0.0 represent mostly electroneutral TcO2. As for the structure of Tc(IV)complex, we can suggest that in our experiments reduced technetium existed as a complex TcIVO(OH)3(CO3)-. According to literature data there exist several aqueous complexes : TcO(OH)2 (aq), TcO(OH)3-, Tc(OH)2CO3 and TcIVO(OH)3(CO3)-. Within the pH range from 8 to 11, the anionic complex TcIVO(OH)3(CO3)- is the dominant one [23].
Figure 3. Chromatographic separation of technetium species in the bacterial supernatant after cultivation with haloalkaliphilic bacteria. Lines 1, 4, 7 – control, lines 2, 3, 5, 6, 8-11 – bacterial supernatant 3.3 Kinetics of technetium reduction by haloalkaliphilic bacteria Although the physiological properties of strains showed some differences, the patterns of technetium (VII) reduction were about the same (see 3.2.). For that reason, only one strain was used for kinetic studies. The kinetics of Tc(VII) reduction were monitored over a 34-h period with pre-grown cells of strain Mono (Fig. 4). Microbial reduction of pertechnetate could be described as a two-phase process: first, very fast, where about 40% of pertechnetate was reduced within 23 h and second, very slow, when technetium reduction continued at a very low rate. In addition, dependence of Tc-reduction on amount of biomass was tested. The results show direct correlation between the amount of bacteria and technetium reduction: as the concentration of bacteria in the experiment increases the amount of reduced technetium also increases (Fig. 4a,b). Finaly, after 34 h of incubation 55% remained as Tc(VII), 36% was found as Tc(IV) and 8% as Tc(V).
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Figure 4. Kinetics of microbial reduction of technetium by strain Mono over period of 34 h. (a) – amount of pertechnetate and (b) – reduced technetium in the bacterial medium. Quantification was performed using PhosphorImager and verified with extraction techniques.Values represent the mean of four replicas. ● – without bacteria, ○ – with dead biomass ▲, ∆, □, ■) – bacterial experiment, were cell protein concentration was 0.017 and 0.070 mg/ml, respectively. (▲, ■) -Tc(VII), (∆, □) reduced technetium. 3.
CONCLUSION The results of this research may be important for the fate and transport of technetium in the environment. Haloalkalophilic heterotrophic microorganisms could provide a perspective for the biotechnological treatment of low level radioactive waste. Our isolates were capable of using formate, acetate, lactate, methanol and ethanol as electron donors and pertechnetate, nitrate, selenite, selenate, tellurite, chromate and elemental sulfur as electron acceptors [24]. This wide variety of substrates and resistance to high concentration of technetium (1.5 mM [24]) make such bacteria attractive for biotechnological applications. However, the formation of highly electronegative soluble Tc(IV) carbonate complexes indicates that it may be necessary to reassess current concepts of Tc transport in anaerobic, carbonate enriched ground waters, where Tc mobility has been considered to be controlled by the low solubility of TcO2. REFERENCES 1. T.M. Beasley, P.R. Dixon and L.J. Mann, Environ. Sci. Technol. 32 (1998) 3875. 2. M. Masson, F. Patti, C. Colle, P. Roucoux, A. Grauby and A. Saas, Health Phys. 57/2 (1989) 269. 3. K.E. Guerman, V.F. Peretrukhin, L.I. Belyaeva and O.V. Kuzina, J. Nuclear Biology and Medicine 38/3 (1994) 406. 4. D.R. Lovley, Annu. Rev. Microbiol. 47 (1993) 263. 5. J.R. Lloyd, J. Ridley, T. Khizniak, N.N. Lyalikova and L.E. Macaskie, Appl. Environ. Microbiol. 65/6 (1999) 2691.
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6. J.R. Lloyd, A.N. Mabbett, D.R. Williams and L.E. Macaskie, Hydrometallurgy 59 (2001) 327. 7. R.E. Wildung, Y.A. Gorby, K.M. Krupka, N.J. Hess, S.W. Li, A.E. Plymale, J.P. McKinley and J.K. Fredrickson, Appl. Environ. Microbiol. 66/6 (2000) 2451. 8. J.R. Lloyd, V.A. Sole, C.V.G. Van Praagh and D.R. Lovley, Appl. Environ. Microbiol. 66/9 (2000) 3743. 9. J.R. Lloyd, G.H. Thomas, J.A. Finlay, J.A. Cole and L.E. Macaskie, Biotechnology and Bioengineering 66/2 (1999) 122. 10. N.N. Lyalikova and T.V. Khizhnyak, Microbiologia 65/4 (1996) 468. 11. D.Yu. Sorokin, A.M. Lysenko and L. Mityushina, Microbiologia 65 (1996) 370. 12. D.Yu. Sorokin and L. Mityushina, Microbiologia 67 (1998) 93. 13. D.Yu. Sorokin, T.P. Tourova, A.M. Lysenko and J.G. Kuenen, Appl. Environ. Microbiol. 67 (2001) 528. 14. N. Pfennig and K.D. Lippert, Arch. Microbiol. 55/3 (1966) 245. 15. M.M. Bredford, Anal. Biochem. 72 (1976) 248. 16. J. Paquette and W.E. Lawrence, Can. J. Chem. 63 (1985) 2369. 17. S.K. Shukla, J.Chromatogr. 21 (1966) 92. 18. M.R.Pillai, K. Kothari, S. Banerjee, G. Samuel, M. Suresh, H.D. Sarma and S. Jurisson, Nuclear Medicine and Biology 26 (1999) 555. 19. T. Maksin, J. Vucina, D Djokie and D. Jankovic, J.Radianal. Nucl.Chem. 243/3 (2000) 669. 20. J.R. Lloyd and L.E. Macaskie, Appl. Environ. Microbiol. 62/2 (1996) 578. 21. K. Shwochau, Technetium. Chemistry and Radiopharmaceutical Applications. WilleyVCH, 2000. 22. L.L.-Y. Hwang, N. Ronca, N.A. Solomon and J. Steigman, Int. J. Appl. Radiat. Isot. 35 (1984) 825. 23. T.E. Eriksen, P. Ndalamba, J. Bruno and M. Caceci, Radiochimica Acta 58/59 (1992) 67. 24. T.V. Khijniak, N.N. Medvedeva-Lyalikova and M. Simonoff, FEMS Microbiology Ecology, 44/1 (2003) 109.
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Biodegradation of cyanides under saline conditions by a mixotrophic Pseudomonas putida N.K. Bipinraj, N.R. Joshi and K.M. Paknikar# Microbial Sciences Division, Agharkar Research Institute, G.G. Agarkar Road, Pune 411 004, India Abstract An alkalophillic strain of Pseudomonas putida able to utilize thiocyanate (TC), free cyanide (potassium cyanide-KCN), and metal cyanides such as tetra cyanocuprate (TCC) and tetracyanonickelate (TCN) at pH 7.5 was isolated from highly alkaline (pH 9.5), saline (4% NaCl) and infertile soil using Starkey’s medium by enrichment culture technique. The culture grew by utilizing cyanides as nitrogen source under autotrophic as well as heterotrophic conditions. The cells grew autotrophically by using 1% thiosulphate or 4% tetrathionate or 1% ferrous sulphide or 1% sulphur as energy source. Heterotrophic growth was observed when supplied with 10mM of glucose. Degradation of all forms of cyanide, viz. 2 mM potassium thiocyanate (TC), 0.5 mM tetra cyanocuprate (TCC) as well as tetracyanonickelate (TCN) and 0.2 mM KCN, was studied under heterotrophic as well as autotrophic and saline (4% NaCl) conditions. Degradation was observed only when the medium was supplied with sulphide or glucose was supplied in the medium. At a cell density of 109 cells/ml the culture degraded 99% TCC and 92% TCN within 4 hours, 95% KCN and 96% TC within 6 hours when supplemented with glucose. In the presence of ferrous sulphide as energy donor, degradation was achieved at the level of 81% for KCN and 91% for TC within 6 and 9 hours respectively. These studies demonstrate autotrophic as well as heterotrophic degradation of different forms of cyanides under saline conditions by a mixotrophic strain of Pseudomonas putida. 1.
INTRODUCTION Despite its toxicity cyanide and thiocyanate are introduced into the environment by biological as well as industrial processes. Biologically cyanide is produced by synthesis from cyanogenic glycosides in plants and thiocyanate by biological cyanide detoxification (1, 2). Thiocyanate is a persistent environmental contaminant because it is nonhydrolysable and non-volatile (3). Cyanide is one of the most indispensable industrial chemicals. Large quantities of metal cyanides (TCC, TCN) however, are also released by mining industries in the extraction of metal, especially by plants processing gold ore (4), #
Corresponding author (E-mail:
[email protected]) NKB thanks Council for Scientific and Industrial Research (CSIR, India) for financial support.
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industries manufacturing organic chemicals, steel making as well as metal plating industries and in the coke industries (5). Several techniques are currently employed to detoxify cyanide-containing effluents. Chlorination is the most widely used method. This method, however, is environmentally hazardous and fails to bring the concentrations of CN-, SCN- and their complexes with metals to the permissible limits (6). Other physico-chemical methods, such as coppercatalyzed hydrogen peroxide oxidation, ozonation, electrolytic decomposition, etc. are highly expensive and are rarely used for treatment of metal cyanides. The role of microorganisms in the degradation of cyanide and thiocyanate has been studied for a long time and the advantages of biodegradation of cyanide are well documented (1, 7, 8, 9, 10). Bacterial degradation of cyanide involves hydrolysis of cyanide to carbon dioxide and ammonia whereas thiocyanate degradation involves three steps. In the first step thiocyanate is hydrolysed to cyanate and sulphide. The cyanate is further hydrolysed to carbon dioxide and ammonia while sulphide is oxidized to sulphate. In arid regions, where supplies of fresh water is the limiting factor, gold cyanidation plants have to use ground water that contains low amounts of dissolved organic compounds and invariably has a high salinity. As reported by Scott et al (9), in Western Australian gold fields water used for the cyanidation process to treat arsenopyrite-rich refractory gold ore has to be recycled to the biological oxidation plant. Thiocyanate, a byproduct of the cyanidation process, must be removed, as it is toxic to the sulphide oxidizing bacteria. In such a case, bacterial cultures exhibiting high efficiency of thiocyanate degradation under saline conditions would be beneficial to treat the water to be recycled. However there are very few reports on such bacterial strains. In this paper we report the degradation of different forms of cyanide under saline conditions by a thiocyanate degrading Pseudomonas putida. 2.
MATERIALS AND METHODS
2.1 Screening of cultures for thiocyanate degradation under saline conditions Thiocyanate degrading cultures were screened from cultures enriched for utilizing reduced sulphur compounds as energy sources. These cultures were isolated from a saline, alkaline, barren soil treated with sulphur (11) and were able to utilize thiosulphate, sulphur, tetrathionate and ferrous sulphide as the energy source. In all, 20 cultures were screened for thiocyanate degradation using modified Starkey’s medium. The medium consisted of 0.4% K2HPO4, 0.15% KH2PO4, 0.05% MgSO4, 0.03% yeast extract, 2 mM TC or ammonium sulphate as nitrogen source and 1% thiosulphate as energy donor. Salinity was imparted by adding 4g NaCl. All cultures were inoculated in 100 ml medium in a 250 ml conical flask and incubated at 30°C on an orbital shaker incubator for 72 hours. Thiocyanate degradation was monitored by checking the residual thiocyanate concentration by the method of Stafford and Callely (12). 2.2 Degradation of thiocyanate All degradation studies were carried out in 100 ml modified Starkey’s medium with 2.0 mM KSCN and 4% NaCl. The medium was inoculated with a 1% (v/v) cell suspension corresponding to 1 OD at 540 nm. The ability of the culture to use thiocyanate as nitrogen as well as energy source was checked in the medium, which contained only thiocyanate as sole source of nitrogen and energy. Thiocyanate degradation was also checked in presence of different energy sources, such as elemental sulphur (1%), sodium thiosulphate (1%), 492
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potassium tetrathionate (0.4%), ferrous sulphide (1%) and glucose (10 mM). The growth and thiocyanate degradation were monitored after 72 hours of incubation under shaking conditions at 30°C. Uninoculated controls were also monitored for thiocyanate degradation. 2.3 Degradation of different forms of cyanides under saline conditions In order to check the degradation of different forms of cyanide, the bacteria were grown in Starkey’s medium with 4% NaCl, 2 mM of KSCN as nitrogen source and 10 mM of glucose or 1% ferrous sulphide as energy sources. The culture was harvested after 72 hours of growth by centrifugation at 10,000 rpm for 20 minutes, cells were washed and suspended in 20 ml of Starkey’s basal salt solution. The experiment was carried out in two sets. One set with 10 mM of glucose, the other set containing 1% ferrous sulphide as the energy source. Each set had four aliquots of 20 ml Starky’s medium, which were supplied with 2 mM TC, 0.2 mM KCN, 0.5 mM TCC and 0.5 mM TCN separately. Appropriate controls were run with the experimental samples. Both the sets were inoculated with the washed cell suspension (cell density 109 cells/ml) of the culture at 1% v/v level. The samples were incubated at 30°C at 120 rpm for 10 hours on a rotary shaker incubator. Cyanide degradation was assessed by estimating the respective residual cyanides in cell free culture broth at regular intervals. KCN was estimated using barbituric acid reagent according to APHA-AWWA-WPCF, 1992 (13). TCC and TCN were analyzed by UV spectrophotometric method (14). 3.
RESULTS AND DISCUSSION
3.1 Screening of cultures for thiocyanate degradation under saline conditions All the 20 isolates were enriched in modified Starkey’s medium with ammonium sulphate as nitrogen and 1% thiosulphate as energy source supplemented with 0.03% yeast extract. However, only one culture, designated as N1 grew when ammonium sulphate was replaced with thiocyanate. The culture degraded 40% 2 mM TC within 72 hours. The organism was identified on the basis of its morphological, physiological and biochemical characteristics according to the Bergey’s manual of systematic bacteriology (1989) and was confirmed as Pseudomonas putida by 16s rRNA sequence analysis. 3.2 Degradation of thiocyanate When Pseudomonas putida was inoculated in the medium containing TC as the sole source of nitrogen and energy it failed to grow and degrade thiocyanate. This shows that the culture does not exhibit pure autotrophy for thiocyanate utilization, as has been reported in the case of Thiobacillus species (15, 16, 17). It was found that the Pseudomonas putida needs some energy source other than thiocyanate and 0.03% yeast extract for growth and degradation of thiocyanate. When the degradation was studied using different energy sources, in presence of glucose and ferrous sulphide the culture showed degradation with higher efficiency (90% and 70% respectively). When thiosulphate was provided as energy source the degradation was only 32% while in the presence of tetrathionate and sulphur the cells failed to degrade thiocyanate (Table 1). Based on these results ferrous sulphide and glucose were selected as energy donors for further studies.
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Table 1. Degradation of thiocyanate by Pseudomonas putida in the presence of different energy sources Energy source
Degradation
Glucose
90%
Ferrous sulphide
70%
Thiosulphate
32%
Tetrathionate
-
Sulfur
-
3.3 Degradation of different forms of cyanides under saline conditions At high cell density, 109 cell/mL and in the presence of 10 mM glucose P.putida culture degraded all forms of cyanide with very high efficiency (Fig. 1-4). Within four hours the culture degraded >99% of 0.5mM TCC, while TCN was degraded upto 92%. In the case of KCN and KSCN degradation was 95% and 96% within 6 hours. 0,6
TCN, mM .
TCC, mM.
0,6 0,4 0,2
0,4 0,2 0
0 0
2
4
6
8
0
10
2
4
Figure 1. Degradation of TCC by Pseudomonas putida in presence of glucose
10
Figure 2. Degradation of TCN by Pseudomonas putida in presence of glucose 6
125 100 75 50 25 0
KCN, mg/l .
TC, mg/l .
8
Time, h
Time, h
0
2
4
6
Time, h
8
10
Figure 3. Degradation of TC by Pseudomonas putida in presence of glucose 494
6
4 2 0 0
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Figure 4. Degradation of KCN by Pseudomonas putida in presence of glucose
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6
120
TC, mg/l .
KCN, mg/l .
When ferrous sulphide was used as energy source, the culture could degrade only KCN (81%) and TC (91%). The results are given in Fig.5-6.
4 2 0
80 40 0
0
2
4 6 Time, h
8
10
Figure 5. Degradation of KCN by Pseudomonas putida in presence of FeS
0
2
4
6
Time, h
8
10
Figure 6. Degradation of TC by Pseudomonas putida in presence of FeS
Different authors have reported thiocyanate degradation under autotrophic mode by Thiobacillus sp. (4) and heterotrophic mode by Pseudomonas stutzeri (2). Thiocyanate degradation under saline conditions has been reported by Stott et al (9). Metal cyanide degradation by a heterotrophic Pseudomonas sp was reported by reported by Y.B. Patil (10). However, the culture reported in our studies possess unique characteristics to degrade metal cyanides, thiocyanate and free cyanide under saline conditions with glucose and ferrous sulphide used as energy sources. 4.
CONCLUSIONS The results of this work show that Pseudomonas putida, used in our studies is able to degrade thiocyanate, potassium cyanide, metal cyanides, viz, tetracayno cuprate and nickelate with high efficiency under saline conditions (4% NaCl) when supplied with 10mM of glucose as energy source. The culture can degrade only potassium cyanide and thiocyanate efficiently with 1% ferrous sulphide as energy source. Water in the gold fields of arid regions only low amounts of organic compounds, has high salinity and invariably contains high concentrations of thiocyanate and various metal cyanides producing a restricted environment for bacterial growth. On this background the culture like Pseudomonas putida exhibiting high efficiency of cyanide degradation under saline conditions and metal resistance is unique. These properties make the culture an ideal candidate in designing a bioreactor to treat effluents from such gold fields. REFERENCES 1. 2. 3. 4.
J. Westley, Cyanide in biology., B.Vennesland (ed.), London: Academic Press. J. Stratford, A.E.X.O. Dias, and C.J. Knowles, Microbiol., 140 (1994) 2657. T.I. Mudder and J.L. Whitlock, Minerals Metallurgical Processing, (1984) 161. Y. Katayama, Y. Narahara, Y. Ioue, F. Amano, T. Kanagawa and H. Kuraishi, J.Biol. Chem., 267 (1992) 9170. 5. S. Basheer, O.M. Kut, J.E. Prenosil and J.R. Bourne, Biotechnol. Bioeng., 39 (1992) 629. 6. D.Y. Sorokin, T.P. Tourova, A.M. Lysenko, and J.G. Kuenen, Appl. Environ. Microbiol., 67 (2001) 528.
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7. G.I. Karavaiko, T.F. Kondrat’eva, E.E. Savari, N.V. Grigor’eva and Z.A. Avakyan, Microbiology, 69 (2000) 209. 8. Y. Katayama, A. Hiraishi and H. Kuraishi, Microbiology, 141 (1995) 1469. 9. M.B. Stott, L.R. Zappia, P.D. Franzmann, P.C. Miller, H.R. Watling and M.R. Houchin, Biohydrometallury and the Environment toward the mining of the 21st Century, R. Amils and A. Ballester (eds.), Process Metallurgy 9B, (1999) 809. 10. Y.B. Patil and K.M. Paknikar, Lett. Appl. Microbiol., 30 (2000) 33. 11. S.K. Polumury and K.M. Paknikar, Biohydrometallury and the Environment toward the mining of the 21st Century, R. Amils and A. Ballester (eds.), Process Metallurgy 9B, (1999) 717. 12. D.A. Stafford and A.G. Callely, J. Gen. Microbiol., 55, (1969) 285. 13. APHA-AWWA-WPCF, Standard methods for the examination of water and wastewater, Washington D.C., (1992). 14. G. Rollinson, R. Jones, M.P. Meadows, R.E Harris and C.J. Knowles, FEMS Microbiol. Lett., 40 (1987) 199. 15. F.C. Happold, K.I. Johnstone, H.I. Rogers and J.B. Youatt, J. Gen. Microbil., 10 (1954) 261. 16. J.B. Youatt, J. Gen. Microbiol., 11 (1954) 139. 17. Y. Katayama and H. Kuraishi, Can. J. Microbiol., 24 (1978) 804.
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Bioleach of a fluvial tailings deposit material indicates longterm potential for pollution S. Willschera*, T.R. Clarkb, R.H. Cohenc, J.F. Ranvilled, K.S. Smithe and K. Walton-Daye a
TU Dresden, Faculty of Forest, Geo and Hydro Sciences, Institute of Waste and Site Management, Pratzschwitzer Str. 15, D - 01796 Pirna, Germany b Little Bear Laboratories, Inc., 5906 McIntyre St., Golden, Colorado 80403 USA c Colorado School of Mines, Division of Environmental Science and Engineering, Golden, CO 80401, USA d Colorado School of Mines, Department of Chemistry and Geochemistry, Golden, CO 80401, USA e U.S. Geological Survey, M.S. 973, Denver Federal Center, Denver, CO 80225-0046, USA
Abstract Microbial leaching experiments were carried out with the aim of estimating the longterm potential of a fluvial tailings deposit material to develop an acid mine drainage (AMD) problem and to contaminate ground and surface waters with metal ions and arsenic. In addition to siliceous components, the mineralogy of the fluvial tailings deposit material included small fractions of unaltered sulfides such as sphalerite and chalcopyrite and secondary minerals such as plumbojarosite. The material was inoculated with a mixed culture of autotrophic and heterotrophic acidophiles, and the extent of metal mobilization under different sets of conditions was measured over the time. The leaching experiments were carried out in aerated stirred reactors and percolator columns. In the stirred reactors and percolators, an average of 20% to 40% of the Al, Zn, Cu, Cd and Cr were solubilized during the first day of each experiment. This clearly indicated that the material had been slowly weathering in situ and was consistent with the observed presence of plumbojarosite. Therefore, a substantial reservoir of acid-soluble metals existed already in the deposit. Complete mobilization of Al, Cr, As and Cu was achieved over a period of weeks in the stirred reactors. As the leach progressed, Pb was precipitated quantitatively as an insoluble sulfate. In the percolators, approximately 40% of the As and Cr, 66% of the Cu, 89% of the Cd, and 100% of the Al were solubilized in comparison to results for the stirred reactors. However, dissolution of Zn and Mn was much greater in the percolators in comparison to the stirred reactors. This indicated that more than one mechanism of metal mobilization was being observed under acidic conditions. Of interest was the propensity of the material to biooxidize despite the presence of relatively large concentration of partially mineralized organic carbon. Without mitigation, there may exist conditions suitable for more than one mechanism of naturally occurring bioleaching in this *
Author for correspondence: S. Willscher, e-mail
[email protected]
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kind of material. Assuming that 20-40% of this material had already weathered in the last 100-140 years, further AMD formation and pollution of ground and surface waters with metal ions could continue without mitigation over hundreds of years. Keywords: acid mine drainage (AMD), fluvial tailings deposit, acidophile heterotrophs, complexing organic matter 1.
INTRODUCTION Fluvial tailings deposits consist of mine waste and tailing materials that have been transported and deposited by natural fluvial processes some distance from their origin in a mining district. These deposits may contain pyrite and related sulfidic, metal and arsenic bearing minerals. Flood plains downstream from mining districts can contain abundant amounts of this material, particularly in older mining districts where mining activity occurred before implementation and enforcement of environmental regulations [1]. The weathering of such materials may cause contamination of ground and surface waters by acidity, dissolved metals and salinity, but the potential for such an occurrence is somewhat undefined, especially given a content of organic material. A previous study focused on the potential for chemical elution of metals, acidity and salinity from this tailings material [1]. However, the contribution of microbially enhanced leaching processes was not investigated. Microbial processes play an important role in the weathering of sulfidic rocks and minerals, causing acid mine drainage (AMD) [2-7]. It was the aim of this work to estimate the potential contribution of microbial weathering processes to the development of contamination from this type of material and to assess the long term risks of similar mining sites exposed to geomicrobial weathering processes. 2.
MATERIALS AND METHODS
2.1 Fluvial solids In addition to siliceous components, the mineralogy of the fluvial tailings deposit material included small fractions of unaltered metal sulfides such as pyrite, sphalerite and chalcopyrite, measured by XRD [1]. Minerals such as plumbojarosite, indicative of previous weathering, were also present. Sediment cores were homogenized, digested, and analyzed for metal content using inductively coupled plasma atomic emission spectroscopy (ICP-AES), as described in detail in [1]. Chemical data obtained following acid digestion of the mixed sample correlated well with former analyses of the particular core materials [1]. Its elemental composition was as follows: 6.77% Fe, 0.37% Al, 0.3% Zn, 0.3% Pb, and Mn, Cu, As, Cr and Cd in ppm amounts. The total sulfur content was 3.24%. An aqueous suspension of the material was acidic (pH 3.13 at 3.75% w/v). Unlike a typical metal sulfide feed, the fluvial tailings material also contained a fraction of plant derived organic material (total organic carbon of 3.13%) in various stages of decay. The tailing area itself was largely devoid of vegetation, presumably because of the low pH and the toxicity of the near surface and the plant material was likely imported by wind and periodic flooding events. The total organic carbon (TOC) was determined on a Coulometerics Total Carbon Analyzer, a high-temperature combustion method that titrates the carbon dioxide produced.
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2.2 Acidiphilic inoculum A mixed culture of Fe- and S-oxidizing mesophiles (Little Bear Laboratories), typically maintained on a mixture of elemental sulfur and pyrite, was adapted to the fluvial solids prior to use in the leaching experiments. The growth media and leach solution was a 1:10 diluted 9K basal salts medium [8] (pH < 2). A high oxidation-reduction potential, indicative of Fe-oxidation, was used as a qualitative indicator of culture viability. For the inoculum, the cells were separated by centrifugation of 500 ml culture suspension per each reactor experiment, washed and suspended in a 1:10 diluted 9K basal salt solution, and subsequently used for the leaching experiments. The leaching was carried out in open aerated stirred reactors and percolation columns. 2.3 Stirred reactors The fluvial solids were leached in batch mode using stirred and aerated one-liter round-bottom flasks. A solids concentration of 3.75% (w/v) was suspended with the inoculum in 500 ml of 1:10 diluted 9K basal salts medium. The suspension was stirred at 200 rpm at room temperature and aerated at a rate of 1200 cm3/min. 2.4 Percolation columns Standard counter-current leach columns were used as a second test system. To ensure adequate drainage and air dispersion, fluvial solids were co-mixed 1:1 with a commercial sand (ten-cycles of sulfuric acid pre-washing). Column charges (fluvial solids and sand co-mix) were then loaded into 5 cm (i.d.) polycarbonate columns over a sand support under-layer. The active bed height was 20 cm. Leach solution was applied at a rate of 20 ml/h. Each column was aerated at a rate of 1200 cm3/min. 2.5 Solution analyses Concentrations of mobilized species of general interest in metal sulfide biooxidations (Fe, Zn, As, Cu) as well as some from the standpoint of overall environmental or water quality impact (Mn, Cr, Cd, Pb and Al) were monitored as AMD developed in the test systems. Periodic sub-samples were centrifuged at 4000 rpm for 15 minutes, filtered by 0.2mym, and diluted into 10% HNO3 prior to analyses by ICP-AES. Spike recoveries of standards at regular intervals were used for analytical quality control. The dissolved organic carbon (DOC) was analysed by catalytic oxidation and subsequent determination of the CO2 formed (Sievers 800 TOC Analyzer). 3.
RESULTS AND DISCUSSION Microbial leaching experiments were carried out in two reactor systems to compare the leaching behavior of the solids over the time under different sets of conditions. It was suspected that metal solubilization would proceed more rapidly in the stirred reactors than in the percolators because of more effective mixing conditions (i.e. high transfer rates and friction forces). However, the percolators could represent a system more closely related to actual conditions in the soil/ sediment columns. 3.1 Leaching results in both reactor systems The pH of the leach solution in the stirred reactors decreased to an average of pH 1.0 after one month and then remained fairly constant until termination. The redox potential exceeded 700 mV after 10 days (Fig. 1). Viable cells were present at the end of the experiment. In the case of the percolator columns, an average final pH of 1.4 was 499
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measured. Perhaps because of limited mass transfer of oxygen, the redox potential in the percolators increased more slowly, rising only after 20 days and exceeding 700 mV only after 40 days (Fig. 2). 800
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Figure 2. Course of pH and redox potential in the percolator columns The stirred reactor system was a convenient method for determining in a relatively short time period the quantity of leachable metals, the kinetics of their solubilization, and the emission potential of acidity and salinity. The yield of oxidized Fe, S, As, Al, Cu and Mg in this system was nearly quantitative. In contrast, only 81% of the Zn content was solubilized. Over the course of the experiment, Pb was only marginally soluble (0.6%) and was precipitated quantitatively by microbially generated sulfate after 19 and 25 days. Pb precipitation was confirmed in subsequent microbial leaching experiments (data not shown here). Therefore, despite high bulk concentrations of Pb in the fluvial tailings, this metal was not mobile at higher sulfate concentrations in the surrounding water. However, it remains an environmental concern. The highest concentrations of solubilized metals were achieved in the percolator experiments, with 25 mg/l As, 11 mg/l Cd, 2.27 g/l Al, 2.26 g/l Zn, 84 mg/l Cu, 0.588 g/l Mn, 4 mg/l Cr, and 20 g/l Fe. At termination, the sulfate concentration was 63 g/l. 500
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Individual concentrations of these metals, arsenic, acidity and salinity are environmentally relevant and more so in combination. The experimentally determined concentrations of solubilized contaminants correlated well with data from ground and surface water samples collected during the field study [1]. It’s easy to imagine that these components were released from "hot spots" in the fluvial tailings deposits and penetrated into the surrounding environment. Microbial iron- and sulfur -oxidation in the percolators had an average yield of 40% of that observed in the stirred reactors (Fig. 3). This result was not entirely surprising because of the potential for mass transfer limitations in the percolator system. Likewise, the yield of As and Cr was also 40% of the yield in the stirred reactors. However, slightly higher yields were achieved with respect to Cu (66%), Cd (89%) and Mg (75%). Al and Zn had comparable yields to the stirred reactors. In these cases, leaching was nearly complete. As expected, manganese was solubilized to a greater extent in the percolators than in the stirred reactors. The yield ratio for lead is not shown in Fig. 3 because of its quantitative precipitation in all reactor experiments. 250
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Figure 3. Ratio of the solubilization yields in the percolators compared to the stirred reactors (set to 100%) The different yield ratios for the particular metals and As in the reactor experiments may be indicative of multiple solubilization mechanisms. Whereas As was directly solubilized by biooxidation of the arsenopyrite in the fluvial tailings material, mobilization of Cr may have been coupled to the formation of hydronium ions resulting from sulfideoxidation. This would explain why the yields and solubilization rates of these species are related to the yields and solubilization rates of Fe and S, respectively, resulting from biooxidation (Fig. 3). However, while Cu, Cd and Zn are also suspected to have been mobilized as a consequence of the sulfide biooxidation, ion exchange and organic complex formation may also have been involved. Zinc was solubilized in the percolators at nearly double the rate observed with the stirred reactors, although there may have been greater mass transport limitations in the columns. Therefore, while Cu may be mobilized primarily by biooxidation, Cd and Zn appeared to have been solubilized to a greater extent by other mechanisms. Obviously, Mg and Al were solubilized by ion exchange processes. Their release was somewhat lower than that observed with the stirred reactors. Even here, an enhanced solubilization by organic complex formation is imaginable. Greater mobilization of Mn in the percolators may have occurred in zones of locally lower redox potential. Especially the solubilization of Mn, Zn and Al is going comparable or better in the percolators than in the 501
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stirred reactors - the experimental data for Mn, Zn and Al were consistent with observed environmental measurement data showing high mobility of these metals in surface waters within the environs of such mining sites. As observed in the stirred reactors, Pb was quantitatively precipitated in the percolators at the 24th and 35th day of the experiment. However, even in the presence of sulfate, a marginal but environmentally relevant solubilization of Pb by ion exchange and complex formation may occur. This was also characteristic of environmental samples collected at a field site in a previous study (up to 1 mg/l Pb in the shallow ground water at "hot spots") [1]. 3.2 Dissolved organic carbon (DOC) DOC increased with time in both the stirred reactor and percolator systems (Fig. 4). The maximum concentrations measured were 358 mg/l and 1206 mg/l, respectively. Considering the replacement of the samples taken from the leaching solutions by fresh nutrient solution, the calculated maximum DOC concentrations were 422 mg/l and 1340 mg/l, respectively. The organic carbon dissolved in the percolator columns with an average yield of 0.68% of the total solid used and with a yield of 21.7% of the total organic carbon contained in the solid.
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Figure 4. Course of the DOC in the stirred reactor and percolator experiments UV spectroscopic and titrimetric analyses indicated that, especially in the percolators, the concentration of humic-like substances (such as fulvic acids) increased with increasing DOC (data not shown). The existence of such substances of medium molecular weight and complex functional groups could explain the enhanced solubilization of Zn, Mn, Al, and, to some extent, Cd and Cu in the percolator experiments. 3.3 Microbial life Although these systems were characterized by a high content of dissolved organic carbon, inhibition of the acidophilic biooxidation process wasn’t noted in any experiment. At the end of the experiments, relatively high colony counts (CFU) of heterotrophic acidophilic bacteria were present (Table 1). Possibly, they were eliminating potentially 502
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toxic low molecular weight organic compounds and preventing the inhibition of the autotrophic and acidophilic biooxidizers [9, 10]. Table 1. Colony forming units of acidophilic, autotrophic and heterotophic microorganisms in the reactor experiments (CFU/ml) Acidophilic heterotrophic bacteria 105 - 107
Acidophilic autotrophs 104 - 105
Acidophilic fungi 103 - 105
Furthermore, acidophilic or acid tolerant fungi detected in the reactor experiments are believed to have degraded the plant material in the fluvial tailings deposit to dissolved humic-like substances (Table 1). In this way, heterotrophic acidophiles seem to play an active role in the solubilization process of the metals. Apparently, the microbial community of acidophilic autotrophic biooxidizers, acidophilic heterotrophic bacteria and fungi act in concert to enhance metal solubilization under these conditions. It is conceivable then that these same processes occur in the fluvial tailings deposits within the floodplain. If so, the acidophilic microbial community would enhance the release of metal ions by several different mechanisms. 3.4 Initial solubility of the metals and environmental impact In both the stirred reactor and percolator systems, an average of 20% to 40% of the Al, Zn, Cu, Cd and Cr were solubilized during the first day of each experiment (Fig. 5). This clearly indicated that the material had been slowly weathering in situ and was consistent with the observed presence of plumbojarosite. Therefore, a substantial and readily leachable reservoir of acid-soluble metal compounds already existed in the deposit. 90 str
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Figure 5. Initial solubility of the elements in the tailing sediment Assuming that 20-40% of this material had already weathered in the last 100-140 years and without mitigation, further AMD formation and pollution of ground and surface waters with metal ions could continue over hundreds of years. Because of alternating flooding and the periodic infiltration of water, as well as penetration of oxygen from the air, conditions for microbially enhanced weathering under these circumstances are nearly ideal. The microorganisms apparently have everything necessary for their growth and weathering activity: oxygen, sufficient humidity, and even dissolved phosphate as an essential nutrient is slowly solubilized out of the material. Because of the periodic flooding, especially during the spring season, substances possibly inhibitory to the 503
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microorganisms (e.g. some metal species) would also be removed, new nutrients and essentials could be transported in, and the weathering of the fluvial tailings material could continue over long periods of time. 3.5 Mitigation measures The percolator experiments clearly demonstrated that the penetration capacity of the material is very important for the throughput of air and water to maintain the microbial weathering processes, as well as the transport of the contaminants out of the material. Measures to decrease penetration of air and water into the tailings, combined with the increase of the pH, should inhibit the microbial and geochemical weathering processes by a large extent. In this way, biooxidation processes, as well as the release of dissolved metals, As, acidity and salinity out of the fluvial tailings deposit should be avoided or largely mitigated. The site was recently remediated with 100 dry tons per acre compost and 100 tons per acre 3/8"- agricultural grade lime. The material was tilled to approximately one foot depth and seeded with field peas, Jose tall wheatgrass, Canadian wild rye, Canadian bluegrass (Reubens), alkali grass (Fults), and red top. Further investigations will show the success of the remediation measures. 4.
CONCLUSIONS Stirred reactors and percolator columns are convenient systems for the investigation of microbially enhanced weathering of mine waste materials. The stirred reactor was useful for rapidly assessing the solubilization potential of the contaminants, whereas the percolator columns provided a system more closely related to field conditions. Upon comparing results from both reactor systems, it was evident that there may exist more than one mechanism of naturally occurring bioleaching in this kind of material. In addition to biooxidation, ion exchange, oxide dissolution and complexation by microbially produced organic substances were observed. This complex solubilization process was carried out by a cooperative microbial mixed culture consisting of acidophilic biooxidizers, acidophilic heterotrophic bacteria and fungi. Such cooperation seemed to enhance metal dissolution. The results, especially with regard to Mn, Zn and Al, were consistent with observations of ground and surface water contamination at other related sites. The complex mobilization process observed here will be the subject of further investigations. Data from the percolator experiments indicated that measures for decreasing the permeability of the tailings to air and infiltration of water in combination with other measures to increase the pH should minimize the microbial and geochemical weathering processes by a large extent. The success of the recent remediation measures will be the subject of further investigations. ACKNOWLEDGEMENTS S.W. is grateful for a grant from the DBU (German Federal Foundation of Environment, Nr. 06000/693).
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REFERENCES 1. K. Walton-Day, F.J. Rossi, L.J. Gerner, J.B. Evans, T.J. Yager, J.F. Ranville, K. Smith, Effects of fluvial tailings deposits on soils and surface- and groundwater quality, and implications for remediation – Upper Arkansas River, Colorado, 1992 – 1996, U.S. Geological Survey, Water-Resources Investigations Report, 99 - 4273 2. V.P.B. Evangelou, Pyrite oxidation and its control, CRC Press, Boca Raton, Florida, U.S.A., 1995 3. G.R. Watzlaf, Mine drainage and surface mine reclamation, D.S. Brown, D.P. Hodel (eds.), Bureau of Mines, U.S. Department of the Interior, IC 9183 (1988), 109 4. H.L. Ehrlich, Geomicrobiology, Marcel Dekker, New York, 1990 5. G. Rossi, Biohydrometallurgy, Mc Graw Hill, Hamburg, Germany 1990 6. A.P. Harrison, Annu. Rev. Microbiol.38 (1984) 265 7. L.C. Bryner, J.V. Beck, D.B. Davis, D.G. Wilson, Ind. Eng. Chem. 46 (1954) 2587 8. M. Silverman, D. Lundgren, J. Bacteriol. 77 (1959) 642 9. D.B. Johnson, F.F. Roberto in: Biomining, D.E. Rawlings (ed.), Springer, Berlin 1997 10. D.B. Johnson, J. Microbiol. Meth., 23 (1995) 205
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Bioleaching of copper converter slag using A. ferrooxidans Seyed Baghery S.A.a∗, Oliazadeh M.b a
R & D Division, National Iranian Copper Industries Company, Sarcheshmeh Copper Complex, Rafsanjan, Iran b Faculty of Mining Engineering, Tehran University, Tehran, Iran
Abstract In this paper, bioleaching of copper converter furnace slag by means of mesophilic bacterium, A. ferrooxidans, has been investigated. At Sarceshmeh Copper Complex, Iran, 90% of converter slag is turned back to the reverberatory furnaces and the remaining quantity, which is estimated at 14000 tones annually, is dumped. The total copper content of slag is 6%, while 12% of the total copper is present as oxide. In this study, bioleaching was used to treat the slag in bench scale using shake flasks. Mineralogical analyses showed that sulfide copper exists mainly in the form of chalcocite and then chalcopyrite, which can be suitable minerals for bioleaching. A 5% pulp composed of pulverized slag and 0K medium was prepared and inoculated with active cells of A. ferrooxidans at pH of 1.9 and temperature of 32ºC. Eh, pH and the copper concentration of the media were measured everyday. Comparing the control and the culture media, it was concluded that acid consumption of control media is higher than the culture ones. Hence, acid production took place in the media containing bacteria. The Eh of the control and culture media reached 550 and 800 mV (SHE) respectively. Copper recovery from culture media was 77.5%, while this figure for the control medium was only 22.4%. The overall results revealed that the copper converter slag can be extracted by bioleaching process. Keywords: copper, converter slag, bioleaching 1.
INTRODUCTION Mineral biooxidation, also called bioleaching, applies certain microorganisms to oxidize sulfide minerals present in ores or concentrates. In this process metals, like copper are released into a dilute sulfuric acid solution for recovery by conventional methods [1]. Other suitable substances for bioleaching process are: reverberatory and converter furnaces slag, flotation tailings and mixed oxide/sulfide copper ores [2, 3]. Quit a few studies have been performed on slag bioleaching, a substance that is rich in copper mostly in the form of sulfide minerals [4]. This paper deals with the bioleaching of copper converter slag produced at Sarcheshmeh Copper Complex; where about the 90% of the slag is turned back to the reverberatory furnaces, and the remaining amount of 14000 tones per annum, is dumped. The dump is estimated to contain 200000 tones converter slag with a copper grade about 6% w/w. Some processes have been developed to treat the ∗ Corresponding Author: e-mail:
[email protected]
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copper converter slag like flotation, electric furnace and carbothermic reduction. Bioleaching, which is another alternative for slag treatment, is the subject of the present paper. 2.
MATERIALS AND METHODS This is the first biologically experiment done on the Sarcheshmeh copper converter slag, so it can be considered as a pre-feasibility test, hence some factors, like initial pH, pulp density, culture medium and the type of inoculated bacterium were considered as constant parameters. 2.1 Bacteria and culture medium A previously isolated adapted strain of A. ferrooxidans from Sarcheshmeh Copper Mine was used [5]. The optimum temperature, pH and culture medium for this strain were 32ْC, 1.9 and 9K medium respectively [6, 7]. 2.2 Slag A sample of copper converter slag was taken from the nearby slag dump. A representative sample was collected by sampling from different spots of the dump. 2.3 Methods pH and Eh values measured daily using a WTW pH/Eh meter model 323. Copper was analyzed by AAS method. Samples of 10 grams pulverized slag (80%< 100µ) and 185 ml 0K medium were added to 500 ml flasks and the initial pH was adjusted and maintained during the test to 1.9 using concentrated sulfuric acid. After incubation in a Kohner shaker incubator at 32ºC, a 5% inoculation of a fresh culture of A. ferrooxidans was added to the flasks. The experiment was run in three flasks, a culture medium with one replicate and a control one without bacteria. The control flask could also be considered as an acid leach medium. The speed of the incubator shaker was adjusted to 150 rpm. Periodically, some samples were taken and pH, Eh and copper concentration were measured. The value of pH was adjusted to 1.9 in cases of its increase. The test lasted for 161 hours. A sample of control medium was taken periodically and checked to ensure the absence of bacteria. After the test completion, the slag residue was washed, dried and used for chemical and mineralogical tests. 3.
RESULTS AND DISCUSSION Chemical and mineralogical analyses of the slag sample are shown in Table 1. Copper existed as chalcocite, chalcopyrite, bornite and native copper. There was also a great amount of iron compounds as fayalite (Fe2SiO4) and magnetite (Fe3O4) generated during the smelt process. Table 1. Chemical and mineralogical analyses of copper converter slag Element Fe Cut CuOX* Cu2S CuS CuFeS2 Cu5FeS4 CuN** FeS2 Fe2SiO4 Fe3O4 /Compound Amount 38.5 5.8 0.73 4.3 1.0 2.3 0.5 0.4 2.2 26.3 47.8 (%w/w) * Copper in acid leachable forms ** Native copper
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Figure 1 shows pH variations in culture and control media. Due to the presence of acid consuming compounds, changes of pH in the first days and in both media were high. But as it can be seen, the control medium consumed more acid than the culture one. It may be concluded that some acid producing reactions were taken place at the culture medium following bacterial activity. From the hour 110 up to the end of the test no sulfuric acid was added to the culture one, while the control one required acid addition to maintain the pH at the desired value. 4 3,5 3 pH
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Figure 1. pH changes in culture and control media during bacterial leaching of slag Figure 2 shows the redox potential variations during the process. After 47 hours incubation, bacteria started to grow (log phase) and the Eh increased at a rate much more than the control one. At the same time, copper concentration increased as well (Figure 3). By the time the Eh was stabilized, copper recovery was lowered. But still there were enough copper minerals to continue the process (Table 2). It may be because of achieving the high Eh value in which the chalcopyrite oxidation is stopped. There may be also another reason like the precipitation of ferric ion complexes because of the very high pH variations. 850 800 750 700 Eh(mV), 650 (Pt-SHE) 600 550 500 450 400 0
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Figure 2. Eh changes in culture and control media during bacterial leaching of slag According to Figure 3, after an initial high increase in copper concentration in the control medium (due to the presence of acid soluble copper compounds), it further increased in a very low rate (3 mg/l.hr) while the respective rate for the culture medium was 20 mg/l.hr. 509
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Figure 3. Copper concentration increase in culture and control media during bacterial leaching of slag Table 2 shows the chemical and the mineralogical analysis of the bioleached and the chemical leached slag samples. Total recovery of copper from the culture bioleached medium was 77.5% while this figure for the control medium was only 22.4%. The mineralogy analysis showed that still there was a considerable amount of copper especially in the form of chalcopyrite. Hence, some preparations should be applied to leach this mineral. Bioleaching at higher temperatures may be a suitable process. Table 2. The chemical and mineralogical analyses of the bioleached and the leached slag samples Element Fe Cut CuOX Cu2S CuS CuFeS2 Cu5FeS4 CuN FeS2 Fe2SiO4 Fe3O4 /compound Culture 34 1.3 0.02 0.2 0.0 3.2 0.0 0.0 1.2 19.7 41.5 Amount Med. (%w/w) Control 33 4.5 0.05 3.2 0.0 6.8 0.0 0.0 2 25.5 42.1 Med.
Similar studies previously performed by K. D. Mehta, et al, (1999) working on the bioleaching of copper converter slag in India indicated that 99% of copper was recovered in 1920 hours, while the recovery (77.5%) in the present study obtained in just 161 hours. It seems that Sarcheshmeh copper converter slag is a more amenable substrate to bioleaching process. 4.
CONCLUSIONS The results showed the possibility of copper extraction from Sarcheshmeh copper converter slag dump (which amounts to about 200000 tones) using native bacteria isolated from this mine. Comparing the obtained results with the other same studies, one can conclude that the Sarcheshmeh converter slag is a more amenable substrate to bacterial leaching process. It is necessary to measure the ferric to ferrous ratio throughout the process and also the effect of mixed mesophilic bacteria inoculation. Regarding the substantial amount of chalcopyrite left in the residue of bioleached slag, it is suggested to run some tests using extremely thermophilic bacteria, as these are weak iron oxidizers and will maintain the redox potential in the range, where chalcopyrite is more active and ready to oxidize.
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It is proposed for the next tests to inoculate the bacteria when the pH of the medium reaches its minimum variation. REFERENCES 1. Brierley, C. L. and A. P. Briggs, (2002), Selection and sizing of biooxidation equipment and circuits, in: Andrew L. Mular, Doug N. Halbe, Derek J. Barrat, (eds.) Mineral processing plant design, practice and control proceedings, 1540-1568. 2. F. Acevedo and J. C. Jentina, (1993), Bioleaching of minerals: a valid alternative for developing countries, Journal of biotechnology, 31, 115-123. 3. H. Brandi (1999), from waste to resources: Microbial metal recovery from slag and ash, The AusIMM proceedings, No.1. 4. K. D. Mehta, B. D. Pandey and Premchand (1999), Bioassisted leaching of copper, nickel and cobalt from copper converter slag, Materials Transactions, JIM, 40, 3, 214221. 5. S. A. Seyed Baghery, H. R. Hassani, (2001), Isolation and preliminary identification of some iron- and sulphur-oxidizing bacteria from Sarcheshmeh Copper Mine, in: S. T. Cimminelli and O. Garcia Jr (eds.), Biohydrometallurgy: Fundamentals, Technology and Sustainable Development, Part A, 393-396. 6. J. Salehi, R. Vagher and S. A. Seyed Baghery, Bioleaching of chalcocite ore of Dahane Sia Mine, (2002), 6th annual congress of Iranian Metallurgy Society proceedings, 253263. 7. Z. Manfi, A. R. Shahverdi, S. A. Seyed Baghery, M. Oliazadeh, (2002) Column bioleaching of the agglomerated low grade copper ore, the 1st symposium on the applications of biotechnology in Kerman province, Iran, 51.
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Biooxidation of mine tailings using a mixed bacterial population Mior Ahmad Khushairi Mohd. Zahari, Jefri Jaapar, Mohd Azri Bunyok, Shahrul Halim Sohor and Wan Azlina Ahmad* Biotechnology Laboratory, Chemistry Department, Faculty of Science, University Technology Malaysia, 81310 UTM Skudai, Johor, Malaysia Abstract In this study, mine tailings obtained from the Lubuk Mandi Gold Mine in Terengganu, Malaysia were biooxidized using a mixed culture comprising Thiobacillus thiooxidans (TT), Thiobacillus ferrooxidans (TF), Leptospirillum ferrooxidans (LF) and Caldibacillus ferrivorus (CF). XRD analysis of the tailings indicated the presence of pyrite (FeS2), hematite (Fe2O3) and arsenopyrite (AsFeS2). A substantial amount of carbon and sulphur, i.e. 0.28% and 0.18% respectively, were also detected. The biooxidation experiments were first carried out in shake flasks using a mixed culture of bacteria at varying ratios. Results from biooxidation studies in shake flasks showed that mixed culture consisting of TT, TF, LF and CF at a ratio of 3:1:1:3 decreased the percentage of preg-robbing by a factor of 3 compared to the control. Batch studies using a stirred tank reactor (STR) were also conducted prior to a continuous mode of operation (CSTR). In the batch mode, the biooxidation mixture comprised of 1.8L of modified CF medium, 10% w/v of tailings and 200mL of culture comprising TT: TF: LF: CF at the following ratios 1:1:1:1 and 3:1:1:3 respectively, and was run for four days. From this experiment, the solubilisation of iron for the 3:1:1:3 mixed culture was 0.5 times higher than the 1:1:1:1 mixture and 1.0 fold higher than the control. Highest gold recovery was observed in the 3:1:1:3 cultures i.e. 65%, compared to the 1:1:1:1 cultures (44%) and 21% for the control. 1.
INTRODUCTION Tailings refer to the processed ore generated from active and inactive mining sites, which often contain the same heavy metals and acid forming minerals as the ore. In Malaysia, large quantities of tailings are generated due to the closure of old mining areas. However, surveys on the amounts of tailings generated have not been carried out. Estimation of the amounts of tailings generated is based on extraction rate and grade of the ore. As an example, the closure of Mamut Copper Mine, in Ranau, Sabah had contributed about 160 million tonnes of tailings expected to contain precious metals such as copper and gold in an area of 400 hectares [1]. Another example of an inactive mining site is Lubuk Mandi Gold Mine, which is situated in the district of Marang, about 15 km south of Kuala Terengganu, Terengganu. One of the prime reasons for the decrease in mining activity is due to depletion of high-
*
[email protected]
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grade gold ore reserves and the abundance of low-grade carbonaceous ore. According to Sia Hok Kiang, only 40% gold recovery was achieved using the Carbon In Leach (CIL) and Carbon In Pulp (CIP) process [2]. It was suspected that the gold was trapped by carbon due to “preg-robbing”. One alternative to overcome this problem is using the roasting technique. However, it is not economical due to low concentrations of gold present in the ore. As a result, 1.3 million tonnes of tailings was left idle in the tailings dam [2]. 2.
MATERIALS AND METHODS
2.1 Preparation of growth media Four types of bacteria were used for this biooxidation study. Before biooxidation, bacteria were first grown in their respective media, TT in TT medium, TF and LF in 9K medium and CF in CF medium. For the biooxidation experiments, mixed culture was subcultured in modified CF medium supplemented with 10% tailings. The T. thiooxidans medium was prepared by dissolving NH4Cl (0.10g), KH2PO4 (3.00g), MgCl2.6H20 (0.10g), and CaCl2.2H2O (0.14g) in 1000mL of deionised water. The pH of the solution was adjusted to 4.2 using H2SO4 (5M) before autoclaving at 121°C, 103.42 kPa for 15 min. Sulphur, which had been autoclaved separately (112°C, 15 min) was then added to the medium at 5% (w/v) final concentration [3]. The 9K medium consists of two solutions, i.e. basal salts solution and ferrous sulphate solution. The basal salts solution was prepared by dissolving (NH4)2SO4 (0.30g), K2HPO4 (0.50g), MgSO4.7H2O (0.50g) and KCl (0.10g) in 700mL deionised water. The pH of the solution was adjusted to 2.0 by adding concentrated sulphuric acid. It was then sterilised by autoclaving at 121°C, 103.42kPa for 15 min. The ferrous sulphate solution was prepared by dissolving 44.22g of FeSO4.7H2O in 300mL deionised water. The pH of the solution was adjusted to 1.8 and filter sterilised using a 0.45µm cellulose acetate membrane. The filter-sterilised ferrous sulphate solution was then added aseptically into the sterilised basal salt medium. For the basal salts medium, the procedure is similar to the preparation of basal salts for 9K medium. The difference is the final volume was made to one litre instead of 700mL. The CF medium consists of four solutions, i.e. basal salts solution, 20 mM ferrous sulphate, 0.1M potassium tetrathionate solution and glycerol (10mM final concentration) [4]. The basal salts solution was prepared by dissolving (NH4)2SO4 (1.50g), K2HPO4 (0.50g), MgSO4.7H2O (5.00g), KCl (0.50g) and Ca(NO3)2.4H2O (0.14g) in 1000mL deionised water. The pH of the solution was adjusted to 2.0 by adding concentrated sulphuric acid. It was then sterilised by autoclaving at 121°C, 103.42 kPa for 15 min. The modified CF medium consists of only the basal salts solution without any addition of ferrous sulphate solution, potassium tetrathionate solution and glycerol. 2.2 Culture preparation and adaptation studies Before adaptation of bacteria can be made, the preparation of culture was first carried out in their respective media. Under normal circumstances, iron-oxidizing bacteria (T. ferrooxidans and L. ferrooxidans) were subcultured (10% inoculum) into 250mL flasks containing 25mL fresh 9K medium and incubated in an orbital shaker at 200 rpm and
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30°C. The sulphur oxidizing bacteria, T. thiooxidans was inoculated into T. thiooxidans medium [3]. For adaptation studies, growth of T. ferrooxidans and L. ferrooxidans were continued in the basal salts medium containing tailings instead of iron. In the case of T. thiooxidans, tailings were added to the T. thiooxidans medium instead of sulphur. For C. ferrivorus, tailings were added to the modified CF medium instead of iron, tetrathionate and glycerol. The bacteria first grown in their respective medium were then transferred into the basal salts medium (for T. ferrooxidans and L. ferrooxidans), TT medium and modified CF mediumcontaining increasing concentrations of the tailings. The final concentration of the tailings was 10% w/v. Bacterial activity was monitored periodically under the microscope using the hanging drop technique. Dissolved Oxygen (DO), Redox Potential (Eh) and pH of medium were checked using WTW MultiLab P4 meter. 2.3 Biooxidation of mixed culture in shake flasks Bacteria that have been adapted to 10% tailings material were mixed in varying ratios (Table 1) in a 2 L flask containing 200mL modified CF medium and supplemented with 10% (w/v) tailings. They were then incubated for 40 days in an orbital shaker at 200 rpm and 38°C. Table 1. The ratio and combination of bacteria used in biooxidation test in shake flasks Mixed Culture Control TT: CF TF: CF LF: CF TT: TF: LF: CF TT: TF: LF: CF TT: TF: LF: CF
Ratio – 1:1 1:1 1:1 1:1:1:1 3:1:1:1 3:1:1:3
2.4 Carbon degradation test The biooxidized tailings material treated with the mixed culture of bacteria in Section 2.3 were separated from the suspending medium. The tailings material was then washed and rinsed using distilled water, and then dried in the oven. Preg-robbing analysis was carried out by adding 50mL distilled water to the dried tailings in a 2L roller bottle. Soluble gold, 5 ppm was added to the slurry and pH adjustment to 10-11.5 was made using lime. NaCN, 0.05g was added to the mixture and the bottle was rolled at 50 rpm for 24 hours. The slurry was then filtered and the gold content in the filtrate was determined using AAS. 2.5 Biooxidation of mixed culture in stirred tank reactor (STR) A preliminary study on biooxidation of mine tailings by mixed bacterial culture was carried out in a stirred tank reactor (STR). The biooxidation study was run in a batch mode with 1.8L modified CF medium, 10% w/v of tailings and 200mL of culture (T. thiooxidans: T. ferrooxidans: L. ferrooxidans: C. ferrivorus, ratio 1:1:1:1 and 3:1:1:3). A comparison was also made with the control, which consisted of 2L deionised water with pH adjusted to 2.0 by adding concentrated H2SO4 and 10% w/v tailings only. The following conditions were set: temperature (38°C); stirring speed (350 rpm) and air flow rate of 25 Lh-1 and carried out for four days. Samples were taken every 8 hrs to determine 515
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total iron solubilised, pH, Eh and dissolved oxygen (DO). Figure 1 shows the schematic diagram for the bioreactor used in this study. Motor Exit air
Rotameter Compressed air
Filling port Cooling water in
Cooling water out
Stirrer
Sparger
Fermentor
Figure 1. Bioreactor used in the experiments
2.6 Gold and iron analysis The biooxidized tailings material treated with the mixed culture of bacteria were separated from the suspending medium. Iron content in the filtrate was determined using AAS (Philips PU9100X). The tailings were then washed and rinsed using distilled water, and dried in the oven. The tailings were then leached using cyanide for extraction of gold. Dried tailings, 200g was mixed with 466mL of water in a 2L bottle. The pH of the slurry was then adjusted to 10-11.5 by adding lime. NaCN, 0.4g was added to the slurry and the bottle was rolled at 50 rpm for 24 hrs. The slurry was then filtered. Gold content in the filtrate was determined using AAS. 3.
RESULTS AND DISCUSSION
3.1 Adaptation of culture with tailings Bacterial adaptation plays an important role in enhancing bioleaching rate. Adaptation enables the bacteria to work more efficiently under higher concentrations of metal ions and lower pH [5]. In microbial leaching of sulphide minerals, bacteria can tolerate higher concentrations of metal ions by exposing the bacteria to a gradual increase in the metal ion concentration. It was found that bacterial adaptation prior to bioleaching enhanced the degree of sulphide leaching 2-4-fold over non-adapted bacteria [5]. In this study, bacteria were adapted to the tailings by gradually increasing the tailings concentrations in the medium. Visual inspection under the microscope using hanging drop technique showed that bacterial activity and population was maintained by the gradual increase in the tailings concentration. The increase in total iron solubilised in the medium also indicated that the bacteria have been successfully adapted to the tailings. 3.2 Biooxidation of mixed culture in shake flask The use of mixed cultures in mineral processing is well documented and was shown to accelerate oxidation of sulphide minerals, compared to pure cultures [6,7]. Manipulation of moderately thermophilic acidophiles in bioleaching processes has also been reported to enhance the dissolution of finely ground mineral sulphide in stirred tank 516
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reactors [8, 9] and in bioheap processes [10]. In this study, a combination of mesophilic and moderately thermophilic acidophilic bacteria was used in the biooxidation process. 3.3 Carbon degradation test Table 2 shows the preg-robbing analysis for the biooxidized tailings carried out in shake flasks. Table 2. Carbon degradation tests for biooxidized tailings in shake flasks study Mixed Culture Control TT: CF TF: CF LF: CF TT: TF: LF: CF TT: TF: LF: CF TT: TF: LF: CF
Ratio – 1:1 1:1 1:1 1:1:1:1 3:1:1:1 3:1:1:3
Au in solution (ppm) 1.6 3.3 3.2 3.3 3.2 2.8 4.0
Au adsorbed (%) 68.2 33.6 35.2 34.8 35.6 43.2 19.8
Carbon degradation test was carried out to indicate the ability and the efficiency of the mixed culture in pre-treatment of the tailings. From the result obtained, it was observed that the highest percentage of gold in the form of gold-cyanide complex was adsorbed on the tailings in the control compared to the other experiments. From 5 ppm of the soluble gold added to the mixture, 3.4 ppm or 68.2% of gold was adsorbed on tailings due to pregrobbing phenomena. On the other hand, for experiments with mixed cultures, only a small percentage i.e. < 50% of gold was adsorbed to the tailings. From the results, it was shown that CF could decrease the adsorption of the gold-cyanide complex by deactivating the carbon present in the tailings. The apparent function of CF in this study is to modify the properties of the carbon in the tailings to prevent the removal of the solubilised gold. CF facilitates gold recovery by either utilising or modifying the gold adsorbing carbon of the tailings to a non-preg-robbing form. 3.4 Biooxidation of mixed culture in stirred tank reactor (STR) Biooxidation was first carried out in a batch mode prior to a continuous mode. Different conditions prevail in the shake flasks and fermentor; some of these include aeration, agitation and temperature. In the reactor, aeration was supplied via air sparging and agitation was provided by an impeller or by the motion imparted to the broth (liquid phase) by rising gas bubbles [11]. Temperature was maintained at a constant and uniform value by circulation of cooling water through coils in the vessel or in a jacket surrounding the vessel [12]. In this experiment, pH, Eh, DO and total iron solubilised were analysed to check the efficiency of the biooxidation process in bioreactor (STR). 3.5 DO profile Figure 2 shows the changes in DO value within the period of the biooxidation studies carried out in a stirred tank reactor, for both test and control.
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Dissolved oxygen, DO (mg/L)
8,50 8,00 7,50 7,00 6,50 6,00 5,50 5,00 0 8 16 24 32 40 48 56 64 72 80 88 96
Time (hr) Figure 2. Dissolved oxygen (DO) profile during biooxidation for control ( (TT: TF: LF: CF) ( ) and 3:1:1:3 (TT: TF: LF: CF) cultures ( )
), 1:1:1:1
From the results obtained, it was clearly shown that the DO profile in the control is significantly different to the 1:1:1:1 and 3:1:1:3 (TT: TF: LF: CF) system. For the control, the DO value increased rapidly (from 6.00 to 7.50 mg/L) for the first 8 hrs and decreased slightly until 48 hrs. After that, the DO value decreased with time until the end of the process. This profile was observed in the control system probably due to the chemical oxidation of pyrite by oxygen. In the presence of oxygen and water, pyrite was chemically oxidised to produce sulphate and ferrous iron. However, the oxidation of pyrite by oxygen is a very slow process, occurring just after 8 hrs of the process. The chemical oxidation of pyrite by oxygen can be represented as follows: (1) 2FeS2 + 7O2 + 2H2O → 2Fe2+ + 4SO42- + 4H+ The DO trend in the 1:1:1:1 and 3:1:1:3 system showed utilisation of oxygen for bacterial growth. It is interesting to note that the metabolism of the culture was affected by the concentration of dissolved oxygen in the broth [13]. An adequate supply of oxygen is a prerequisite for good bacterial growth and high activity of the leaching bacteria. This could be achieved by aerating and agitating the broth. In the test, the DO values range from 5.50 mg/L to 8.00 mg/L, showing that the reactor system provided sufficient oxygen for bacterial growth and biooxidation. It was suggested that the minimum DO concentration level to inhibit bacterial activity is between 0.29 mg/L to 1.50 mg/L [14, 15]. In the biooxidation process, there was some correlation between dissolved oxygen and the rate of iron biooxidation. As an example, it was suggested that the maximum biooxidation rate constant in shake flasks was affected by the oxygen transfer rate only at low aeration conditions [16]. From the experiment carried out in shake flask, the critical oxygen transfer rate and the maximum biooxidation rate constant were about 0.03 mmol O2 l-1.min-1 and 0.050 h-1 respectively [16]. An experiment carried out in an aerated, agitated fermentor (pH, 2.3 and temperature, 32°C), a fivefold increase in the volumetric oxygen mass transfer coefficient, kLa from 2.4 x 10-3 to 1.3 x 10-2 s-1, caused a moderate increase in the maximum oxidation rate from 9.65 to 12.6 mmol Fe.l-1.h-1 [17]. The oxygen 518
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gas-liquid transfer rate per unit of slurry volume V, OTRg→l (mol O2.l-1.s-1), was calculated from the following equation [18]: (2) OTRg→l = kLa (O2).(1-ξs).([O2]L* - [O2]L) In these equation kLa (O2) is the volumetric mass transfer rate coefficient of oxygen, [O2]L* is the dissolved oxygen concentration of the liquid phase in equilibrium with the gas phase, [O2]L is the dissolved oxygen concentration and ξs is the solid hold-up in the non-aerated slurry.
700
2,38
600
2,30
500
2,22
400
2,14
300
2,06
200
1,98
100
1,90
pH
Eh (mV)
3.6 pH and Eh profile Figure 3 shows the pH and Eh profile during biooxidation of mine tailings in STR.
0 8 16 24 32 40 48 56 64 72 80 88 96
Time (hr) Figure 3. pH and Eh profile during biooxidation of mine tailings in STR ); pH of 1:1:1:1 (TT:TF:LF:CF) system ( ), Eh of pH control ( ), Eh control ( 1:1:1:1 (TT:TF: LF:CF) system ( ); pH of 3:1:1:3 (TT:TF:LF:CF) system ( ) and Eh of 3:1:1:3 (TT:TF:LF:CF) system ( )
In the control, the Eh value decreased rapidly with increasing pH value. This trend continued until the end of the biooxidation process. This profile was in accordance with the chemical oxidation of pyrite taking place in the reactor. The chemical oxidation of pyrite by ferric iron is shown in Eq. 3: (3) FeS2 + 14Fe3+ + 8H2O → 15Fe2+ + 2SO42- + 16H+ From Eq. 3, ferric iron was utilised to oxidise the pyrite, resulting in a decrease in Eh value. The correlation between Fe3+/Fe2+ and Eh value can be shown by Eq. 4 [7]: Eh = E o +
RT ⎛ Fe3+ ⎞ ⎟ ln⎜ zF ⎜⎝ Fe 2 + ⎟⎠
(4)
From Eq. 4, it can be shown that the Eh value is dependent on the concentration of Fe3+ and Fe2+. The Eh value decreased as the pH value increases for the control as shown in figure 3. For the control, the pH value increased as acid (H+) was being consumed by the tailings due to the presence of alkaline gangue, carbonate and silicate minerals, which will increase the pH value [19][20]. 519
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On the other hand, a different profile was observed in the test. The pH values were found to increase during the 64 hrs and 80 hrs for the 3:1:1:3 and 1:1:1:1 (TT: TF: LF: CF) cultures, respectively, before decreasing to pH 2.00. However, there were no significant changes in the Eh values. This is probably due to the biooxidation behaviour of tailings by the mixed culture. It can be implied that oxidation of a pyrite by mixed culture via biological and chemical reaction is at equilibrium in both tests. This shows that the indirect mechanism was more dominant in the biooxidation process of the system [18]. Pyrite oxidation in a mixed culture is shown in Eq. 3, whilst the biological reaction is shown in Eq. 5: Fe 2 + + H + + 0.25 O 2
bacteria
Fe3+ + 0.5 H 2 O
(5)
By combining Eq. 3 and 5, the overall reaction in the mixed culture can be summarised as follows: FeS2 + 3.75 O 2 + 0.5 H 2 O → Fe3+ + 2 SO 24 − + H +
(6)
3.7 Iron solubilisation profile Figure 4 shows the iron solubilisation profiles of tailings in STR during the biooxidation process.
Iron solubilised (ppm)
7000 6000 5000 4000 3000 2000 1000 0 0
8 16 24 32 40 48 56 64 72 80 88 96 Time (hr)
Figure 4. Iron solubilisation profiles during biooxidation for control ( (TT:TF:LF:CF) ( ) and 3:1:1:3 (TT:TF:LF:CF) cultures ( )
), 1:1:1:1
From the results obtained, it was clearly shown that the use of mixed cultures could increase the efficiency of biooxidation of the tailings. The 3:1:1:3 (TT: TF: LF: CF) culture was the most efficient biooxidation system compared to the 1:1:1:1 (TT: TF: LF: CF) culture and the control. It was observed that the solubilisation of iron in the 3:1:1:3 (TT: TF: LF: CF) culture was 0.52 times higher than the 1:1:1:1 (TT: TF: LF: CF) culture. From the results also, it is interesting to note that the solubilisation of iron in the 1:1:1:1 (TT: TF: LF: CF) culture and 3:1:1:3 (TT: TF: LF: CF) culture was constant during 40 hrs of the biooxidation process. After that, the solubilisation of iron increased rapidly in both tests. However, the solubilisation of iron in the 1:1:1:1 (TT: TF: LF: CF) culture decreased suddenly after 48 hr and was constant until the end of the process. On the other hand, the solubilisation of iron increased until the 64th hr for the 3:1:1:3 (TT: TF: LF: CF) cultures. After that, the concentration of iron decreased rapidly and remained constant until the end of the process. This can be explained as follows: at the beginning of the leaching process, bacteria exhibited a long lag phase during which the metabolism of 520
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the bacteria is readjusted (acclimatized) to the new environment [5]. Biooxidation during this phase is usually slow, because bacteria obtain their energy mainly from the oxidation of sulphur rather than iron during this lag phase [21]. The following reaction occurred during the lag phase: So + 2 H 2 O + O 2
bacteria
SO 24 − + 4 H + + 2 e −
(7)
The bacterial activity increased after the lag phase due to enhanced leaching of sulphides as shown in Eq. 8. This can be summarised by the following reaction: bacteria 4 FeS2 + 15 O 2 + H 2 O 2 Fe 2 (SO 4 )3 + 2 H 2 SO 4 (8) The possible explanation for the decrease in iron solubilization rate may be due to the precipitation of iron in the reactor probably as jarosites. The formation of jarosite increased with temperature as described by Dutrizac [22]. 3.8 Gold analysis Table 3 shows the gold analysis for the biooxidized tailings carried out in a STR. Table 3. Gold analysis for biooxidized tailings Test Control 1:1:1:1 (TT: TF: LF: CF) 3:1:1:3 (TT: TF: LF: CF)
Percent of Gold Recovery (%) 21 44 65
From the results, it can be clearly seen that the use of microbial consortium can enhance gold recovery from the tailings. After cyanidation, 44% of gold was extracted from the biooxidized tailings treated with 1:1:1:1 (TT: TF: LF: CF) and 65% in 3:1:1:3 (TT: TF: LF: CF) cultures. A small percentage of gold was extracted in the control i.e. 21%. The gold extracted in the control could be due to gold found in the free form. From the results obtained, it was observed that highest recovery of gold was achieved in the 3:1:1:3 (TT: TF: LF: CF) cultures probably due to maximum iron solubilisation observed earlier. One plausible explanation is that most of the gold in the tailings was locked in a sulphide mineral, such as pyrite (FeS2) and arsenopyrite (FeAsS). Biooxidation by the mixed culture have successfully solubilised the iron, releasing it into solution, increasing the permeability of the tailings, and subsequently increasing the rate of gold recovery. Furthermore, the combination of iron- and sulphur-oxidising bacteria (TT, TF and LF) and the presence of the carbon-degrading bacteria has increased biooxidation and hence gold recovery. Iron and sulphur oxidising bacteria (TT, TF and LF) play a prominent role as pyrite oxidisers, resulting in solubilisation of iron, as ferric iron, and sulphide, as sulphate [23]. C. ferrivorus, a chemolithoheterotrophic bacterium, had played the most important role in enhancing the recovery of gold. This is because; C. ferrivorus was able to degrade the graphitic carbon compound in the tailings, which is the major cause of "preg-robbing". C. ferrivorus was able to degrade the organic carbon content in the tailings by directly metabolizing the carbon compound to a nonpreg-robbing form. It is important to stress that the oxidation of pyrite and degradation of carbon compound occurred simultaneously. 4.
CONCLUSIONS In conclusion, the 3:1:1:3 (TT: TF: LF: CF) culture was the best combination for the biooxidation process compared to the 1:1:1:1 (TT: TF: LF: CF) culture and the control. For future experiments, this combination and ratio of bacteria will be applied to the 521
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biooxidation process carried out in a continuous system (CSTR). Finally, it can be concluded that the mixed culture used in this study shows great promise in the biooxidation of mine tailings and in solving an important environmental problem. REFERENCES
1. The Star (2001). “New Lease of Life for Mamut Dam.” Star Publications (M) Bhd. March 14, 2001. 2. Sia Hok Kiang (2000). Personal Communication. 3. DSMZ, Deutsche Sammlung von Mikroorganismen und Zelkulturen GmbH (1993). “Catalogue of Strains 1993.” Germany, GmBH. 4. Adibah Yahya (2000). “Physiological and Phylogenetic Studies of Some Novel Acidophilic Mineral-Oxidising Bacteria.” University of Wales, United Kingdom: PhD thesis. 5. Elzeky, M. and Attia, Y.A. (1995). “Effect of Bacterial Adaptation on Kinetics and Mechanisms of Bioleaching Ferrous Sulphides.” The Chemical Engineering Journal. 56. B115-B124. 6. Rawlings, D.E., Tributsch, H. and Hansford G.S. (1999). “Reasons Why Leptospirillum like Species Rather than Thiobacillus ferrooxidans are the Dominant Iron-Oxidizing Bacteria in Many Commercial Processes for the Biooxidation of Pyrite and Related Ore.” Microbiology. 145. 5-13. 7. Battaglia-Brunet, F., d’Hugues, P., Cabral, T., Cezac, P., Garcia, J.L. and Morin, D. (1998). “The Mutual Effect of Mixed Thiobacilli and Lepstospirilli Populations on Pyrite Bioleaching.” Minerals Engineering. 11. 195-205. 8. Sampson, M.I. and Philips, C.V. (2001). “Influence of Base Metals on the Oxidizing Ability of Acidophilic Bacteria During the Oxidation of Ferrous Sulphate and Mineral Sulphide Concentrates, Using Mesophiles and Moderate Thermophiles.” Minerals Engineering. 14, 317-340. 9. Sandstrom, A. and Petersson, S. (1997). “Bioleaching of a Complex Sulphide Ore with Moderate Thermophilic and Extreme Thermophilic Microorganisms.” Hydrometallurgy. 46. 181-190. 10. Brierley, J.A. (1997).“ Heap Leaching of Gold-Bearing Deposits: Theory and Operational Description” in Rawling D.E “Biomining: Theory, Microbes and Industrial Processes” USA: Springer-Verlag and Landes Bioscience. pp. 104-114 11. Fogler, H.S. (1992). “Elements of Chemical Reaction Engineering.” 2nd Edition, New Jersey USA: Prentice-Hall International Inc. 12. Blanch, H.W. and Clark, D.S. (1997). “Biochemical Engineering.” New York: Marcel Dekker Inc. 13. Standbury, P.F., Whitaker, A. and Hall, S.J. (1995). “Principle of Fermentation Technology”. 2nd Edition, Great Britain: Pergamon Press Ltd. 14. Dew, D.W. and Godfrey, M.W. (1991). “Sao Bento BIOX® Reactor.” SAIMM Colloqium, Johannesburg. 15. Nemati, M., Harrison, S.T.L., Hansford, G.S. and Webb, C. (1998). “Biological Oxidation of Ferrous Sulfate by Thiobacillus ferrooxidans: A Review on the Kinetic Aspects.” Biochemical Engineering Journal. 1. 171-190 16. Savic, D.S., Veljkovic, V.B., Lazic, M.L., Vrvic, M.M. and Vucetic, J.I. (1998). “Effects of Oxygen Transfer Rate on Ferrous Iron Oxidation by Thiobacillus ferrooxidans.” Enzyme and Microbial Technology. 23. 427-431. 17. Guay, R., Silver, M., and Torma, A.E. (1977). “Ferrous Iron Oxidation and Uranium Extraction by Thiobacillus ferrooxidans.” Biotechnol. Bioeng., 29. 727-740. 522
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18. Boon, M. (1996). “Theoretical and Experimental Methods in the Modelling of Biooxidation Kinetics of Sulphide Minerals.” Delft University of Technology, Netherlands: PhD thesis. 19. Ahonen, L. and Tuovinen, O.H. (1995). “Bacterial Leaching of Complex Sulfide Ore Samples in Bench-Scale Column Reactors.” Hydrometallurgy. 37. 1-20. 20. Shuey, S.A. (1999). “Bioleaching: The Next Era in Refractory Mineral Processing.” Engineering and Mining Journal, New York. pp 16A-20A. 21. Yu, J.Y., McGenity, T.J. and Coleman, M.L. (2001). “Solution Chemistry During the Lag Phase and Exponential Phase of Pyrite Oxidation by Thiobacillus ferrooxidans.” Chemical Geology. 175. 307-317. 22. Dutrizac, J.E. (1982). “Jarosite-type Compounds and Their Application in the Metallurgical Industry.” in: Hydrometallurgy Research, Development and Plant Practice, TMS-AIME, New York. pp 531-551. 23. Brierley, J. A. and Kulpa, C.F. (1992). “Microbial Consortium Treatment of Refractory Precious Metal Ores.” (US Patent 5,127,942).
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Chromate reduction by immobilized cells of Desulfovibrio vulgaris using biologically produced hydrogen A.C. Humphriesa, D.W. Penfolda, C.F. Forsterb, L.E. Macaskiea* a
b
School of Biosciences, The University of Birmingham, Edgbaston, B15 2TT, UK School of Civil Engineering, The University of Birmingham, Edgbaston, B15 2TT, UK
Abstract Many industrial wastes contain Cr(VI), a carcinogen and mutagen, the toxicity of which can be ameliorated by reduction to Cr(III). Agar-immobilized cells of Desulfovibrio vulgaris NCIMB 8303 can reduce toxic Cr(VI) to less-toxic Cr(III) at the expense of metabolically produced hydrogen as the electron donor in a batch reactor under anaerobic conditions using H2 produced by Escherichia coli HD701. Complete reduction of chromate (500 µM) was achieved within 24 h with an initial reduction rate of 302 nmol/ h/ mg dry cell wt using a coupled dual reactor system. 1.
INTRODUCTION Biological hydrogen can be economically produced by a process in which oxidisable sugar waste can be used to provide electron donors for hydrogen producing bacteria (1). Additional benefits of biological hydrogen production are that it can be coupled to waste bioremediation, for example conversion of Cr(VI) to Cr(III) using bacteria that can effect this reduction via hydrogenase activity. An E. coli strain HD701, which cannot synthesise the formate hydrogen lyase (FHL) repressor (Hyc A) and is, therefore, up-regulated with respect to FHL expression, has been constructed previously (2). The formate hydrogen lyase (FHL) complex comprises formate dehydrogenase and hydrogenase 3. The function of this complex is to convert formate, a metabolic end product of anaerobic fermentation, into CO2 and H2: a homeostatic response to reduce the acidity of the medium. In this respect hydrogenase 3 is acting as a synthetic hydrogenase and thus provides an oxidisable substrate for a second organism that is proposed to oxidise H2 as a source of reducing power. Environmental pollution by hexavalent chromium (CrO42-) in marine and urban sediments and wastewaters originates predominantly from industrial sources, for example, leather tanning, electroplating and textile manufacture (4). Detoxification of Cr(VI), a mutagen and carcinogen, to less toxic Cr(III) can be achieved chemically or biologically. Chemical reduction is expensive and requires large amounts of chemicals, whereas biological reduction by Cr(VI) reducing bacteria is potentially economically favourable because the methods involved are cheap and simple (5). Additionally the use of toxic reagents is not required, and native, non-hazardous microorganisms can be used (6). Many Cr(VI)-reducing microorganisms have been identified, with examples including * Author for correspondence (Fax (44)-121-414-6557; E-mail:
[email protected])
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Pseudomonas fluorescens LB300 (7, 8), Agrobacterium radiobacter EPS-916 (9), Enterobacter cloacae strain HO1 (10, 11), and Desulfovibrio vulgaris NCIMB 8303. The latter is a member of the sulphate-reducing bacteria (SRB) with a well-established ability to reduce a variety of metals, including Cr(VI) (12-17). The SRB use an organic compound (e.g. formate) or molecular H2 as the electron donor and SO42- as the electron acceptor in dissimilatory sulfate reduction. Certain high valence metal ions, such as Tc(VII) and Cr(VI), can be used as electron sinks in the absence of sulfate, with consequent metal reduction (16,18). Previous studies have demonstrated that a complexing agent that is able to chelate Cr(III) must be present to permit Cr(VI) reduction by D. vulgaris (for example bicarbonate or citrate: (16)). This study investigates the potential for use of H2, produced by E.coli HD701 from industrial sugar wastes, as the electron donor for Cr(VI) reduction by D.vulgaris NCIMB 8303. For industrial purposes Cr(VI) reduction using freely suspended cells is disadvantageous because of downstream difficulties in biomass/ effluent separation (19). Consequently chromate reduction by agar-immobilized D. vulgaris NCIMB 8303 in batch mode was investigated, and compared to that of free cells. Agar was identified previously as the most suitable immobilization matrix for this purpose, imposing minimal mass transfer limitations and good structural integrity (A.C. Humphries, unpublished). The possibility of using H2 produced by E. coli HD701 using nougat and caramel wastewater as feedstocks for a dual coupled reactor system was also investigated. 2.
MATERIALS AND METHODS
2.1 Organisms and growth conditions E.coli strain HD701 was maintained as previously described (1). Desulfovibrio vulgaris NCIMB 8303 was obtained from Prof. D. R. Lovley (University of Massachusetts, Amherst, MA, USA). Cells were maintained routinely in butyl-rubber sealed 100 ml serum bottles in Postgate’s medium C (20) under O2-free N2 (OFN). For experiments, growing cultures (OD600 0.8 ± 0.2) were produced by withdrawing anaerobically 10 ml of an actively growing culture and adding it to 100 ml of fresh medium C. The culture was harvested after 24 h (OD600 0.8 ± 0.2) for hydrogen production tests. 2.2 Preparation of cell suspensions Resting cells of D. vulgaris were prepared by transferring cells anaerobically in a MACS-MG anaerobic workstation (Don Whitley Scientific Ltd) to closed centrifuge tubes and harvesting by centrifugation (8000 rpm, 10 min, 4ºC). Cells were washed three times and resuspended with anaerobic (degassed) stock solutions of 20 mM MOPS/ NaOH buffer (pH 7). Cell concentrate was stored at 4 ºC under N2 until use (within 24 h of harvesting). The biomass concentration of the cell concentrate was estimated by measuring the OD600, and conversion to dry cell weight via a previously determined calibration. 2.3 Immobilization of cells 200 mg Agar-Agar (FSB Laboratory supplies) was mixed with 9 ml deionised water, degassed with OFN, and autoclaved (121ºC, 20 min). The gel was allowed to cool to 40ºC before 1 ml cell concentrate (D. vulgaris, as above) was added to give a cell concentration of 2.5 mg dry cell wt/ ml gel. Gel bead formation was carried out in the MACS-MG workstation. The aqueous phase (degassed deionised water) and the oily phase (degassed 526
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olive oil) (21) were placed in a 100 ml vessel in a ratio 10:1 (v/v), respectively. The gelcell mixture was extruded into the oily phase using a 2.5 ml syringe fitted with a 21G needle. The gel-cell beads passed into the aqueous phase and were collected using a nylon net. Beads were washed once with Triton-X100 (Sigma) (0.01% (v/v)) and three times with degassed deionised water (21). 2.4 Hydrogen evolution Hydrogen production experiments used phosphate buffered saline (PBS) in 320 ml reactors (working volume of 310 ml) at 30ºC with 10% (v/v) inocula of E.coli HD701 cells pre-grown in nutrient broth to an OD600 of 1.0 ± 0.01. Carbon substrates were used at 10% of the total volume, and consisted of 100 mM glucose, nougat waste, or caramel waste (see below). Suspensions were gassed with argon for 1 h after inoculation and maintained at 30ºC, with stirring. Evolved hydrogen was passed through 1 M sodium hydroxide to remove carbon dioxide and into 30ºC column reactors containing the immobilized D. vulgaris and chromate reaction solution (Fig. 1A, B) at an approx. rate of 10 ml/ h. The exit gas was bubbled through olive oil to exclude air from the system. (A)
(B)
Tubing Gas Outlet needle
Olive Oil Water Out Cr (VI) reaction solution Water In
1M NaOH
Argon Inlet (for degassing)
30 °C Water Bath
Magnetic flea Magnetic Stirrer
Figure 1. The Cr(VI) reducing dual reactor system 2.5 Chromate reduction by D. vulgaris NCIMB using biologically produced H2 in batch mode For batch reduction, the reaction mixtures were set up in the anaerobic workstation (30ºC) in 14 ml jacketed glass columns. Cell beads (2 ml: equivalent to 5 mg dry cell wt) or free cells (equivalent to 5 mg dry cell wt) were suspended in 10 ml of a degassed solution of 20 mM MOPS-NaOH, 25 mM sodium citrate, and 500 µM sodium chromate, pH 7. Columns were removed from the workstation and degassed with OFN for 15 min. Controls for all experiments comprised cells or cell beads in the absence of hydrogen, or reaction solution in the absence of cells. Columns were maintained at 30ºC. Chromate reduction was initiated by the provision of hydrogen to the cells, either biologically produced (Fig. 1) or supplied from a commercially available hydrogen-containing gas 527
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cylinder. Samples were withdrawn periodically, via syringe, and supernatants (100 µl) were analysed for residual Cr(VI) using diphenylcarbazide (specific for Cr(VI) as described in 5, 6). 3.
RESULTS AND DISCUSSION
3.1 Nougat and Caramel Wastewater Nougat waste composition was as previously described (1). Caramel waste composition is shown in Table 1. Both wastes were stored at 4 ˚C. Table 1. Caramel waste composition Component Total soluble sugars Total reducing sugars Total protein content Ammonium ion Nitrate Phosphate Sulphate Sulphide
Concentration 53.0% (w/w) 25.5% (w/w) 1.9% (w/w) 2.2 mg.l-1 ND* AR** ND* ND*
* ND – Not detected ** AR – Above detectable range
3.2 Cr(VI) reduction by D. vulgaris using glucose as the carbon source for hydrogen production by E. coli Preliminary investigations showed that gas entry into the reaction column was not continuous (due to hydrostatic pressure), resulting in loss of anaerobiosis and subsequent loss of Cr(VI) reducing ability after 2 h. A trap was incorporated in the form of a 12 ml serum bottle containing 5 ml olive oil (Fig 1 A, B). Immobilized and free cells of D. vulgaris were incubated with Cr(VI) solution over 6.5 h, with reduction proceeding only in the presence of hydrogen. No Cr(VI) reduction occurred in the controls (i.e. electron donor and cells were both required). Free and immobilized cells of D. vulgaris reduced Cr(VI) at similar initial rates over the first 2 h at ca. 200 nmol/ h/ mg dry cell wt.). After 5 h, 120 ± 14.5 µM Cr(VI) remained in free cell suspensions, and 190 ± 19.9 µM Cr(VI) remained in immobilized cell reaction solutions (Fig. 2). The initial Cr(VI) reduction rates and final Cr(VI) concentrations are not significantly different (as determined by Students t-test analysis) for free and immobilized cells using biologically produced hydrogen as electron donor, suggesting that there are no mass or heat transfer limitations imposed by immobilisation of D. vulgaris in agar gel. 3.3 Cr(VI) reduction by D. vulgaris using chemical hydrogen All Cr(VI) reduction tests were repeated using commercially available hydrogen supplied from a cylinder at the minimal flow rate that could be attained (ca. 1ml/ min (Fig. 3)). Reduction of Cr(VI) occurred at an initial rate of 150 nmol/ h/ mg dry cell wt, with approx. 300 µM Cr(VI) remaining after 5 h for both free and immobilized cells. Student’s t-test analysis showed that Cr(VI) reduction by D. vulgaris using biologically produced hydrogen was significantly higher than when using chemical hydrogen at the higher gas
528
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flow rate. This diminished reduction capacity may be attributable to a temperature effect, as chemical hydrogen was not pre-warmed prior to entry into the reaction solution. 600
500
[Cr (VI)] (uM)
400
300
200
100
0 0
1
2
3
4
5
6
7
Time (h)
Figure 2. Reduction of 500 µM Cr(VI) by free and agar-immobilized resting cells of D. vulgaris NCIMB 8303 using biologically produced hydrogen as electron donor Legend: o : Free cells plus bio-hydrogen; ● : Free cells in the absence of hydrogen; ∆ : Agarimmobilized cells plus bio-hydrogen; ▲ : Agar-immobilized cells in the absence of hydrogen; ◊ : Agar-immobilized cells plus bio-hydrogen produced using nougat wastewater; ♦ : Agarimmobilized cells plus bio-hydrogen produced using caramel wastewater 700 600
[Cr (VI)] (uM)
500 400 300 200 100 0 0
1
2
3
4
5
6
7
Time (h)
Figure 3. Reduction of 500 µM Cr(VI) by free and agar-immobilized resting cells of D. vulgaris NCIMB 8303 using chemical hydrogen as electron donor Legend: o : Free cells plus bottled hydrogen; ● : Free cells in the absence of hydrogen; ∆ : Agar-immobilized cells plus bottled hydrogen; ▲ : Agar-immobilized cells in the absence of hydrogen
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3.4 Cr(VI) reduction by D. vulgaris using nougat and caramel wastes as oxidisable substrates for E. coli Immobilized cells of D. vulgaris were reacted with Cr(VI) reaction solution over 6.5 h in nougat or caramel waste as a carbon source for the E.coli reactor (Fig. 2). This gave a Cr(VI) removal rate of 234 ± 18.3 and 302 ± 14.5 nmol/ h/ mg dry cell wt., respectively. After 5 h, 110 ± 4.3 µM and 2 ± 0.8 µM Cr(VI) remained in the respective immobilized cell reaction solutions. Reduction of Cr(VI) using hydrogen produced by E.coli fed with nougat waste gave statistically similar (as determined by Students t-test analysis) reduction rates as when provided with 100 mM glucose, whilst those obtained for E.coli fed with caramel waste were significantly higher. The latter was the only substrate to promote complete loss of Cr(VI) within 5 h (Fig. 2). 4.
CONCLUSIONS Hydrogen generated biologically by E. coli HD701, using glucose, nougat or caramel wastewater as an oxidisable substrate, can be used by D. vulgaris NCIMB 8303 as the electron donor for Cr(VI) reduction. The total amounts, and rates, of Cr(VI) reduction by D. vulgaris were significantly higher with biologically evolved hydrogen compared to chemical hydrogen. Chromate reduction using hydrogen generated by E. coli supplied with 100 mM glucose or nougat waste occurred at similar rates with similar concentrations remaining after 6.5 h. In comparison, Cr(VI) reduction using hydrogen generated by E.coli using caramel waste occurred ca. 1.5 times faster, with Cr(VI) totally reduced after 6.5 h. The high rates of Cr(VI) removal using industrial sugar wastes as feedstock for E.coli hydrogen production demonstrate that this is an economically favourable and environmentally friendly system for the use of one waste as a starting material for the remediation of another. ACKNOWLEDGMENTS We wish to acknowledge the financial support of the BBSRC (Studentship Number: 00/B1/E/0698 to D. P.) and the EPSRC (Studentship Number: 00318650 to A. H.). The authors would like to thank Professor A. Böck (Lehrstuhl fur Mikrobiologie der Universitat, Munich, Germany) for providing the strain of E.coli and Cadbury Trebor Bassett Ltd (Bournville, Birmingham, UK) for provision of the nougat and caramel wastewaters. REFERENCES
1. 2. 3. 4. 5. 6.
D. Penfold, C. Forster, L. E. Macaskie. Enzyme Microbiol. Technol. (in submission) M. Sauter, R. Bohm, A. Bock. Mol. Microbiol. 6 (1992) 1523. J. Lloyd, G. Thomas, J. Finlay, J. Cole, L. Macaskie. Biotech. Bioeng. 66 (1999) 122. W. Smith, G. Gadd. Journal of Appl. Microbiol. 88 (2000) 983. P. Pattanapipitpaisal, N. Brown, L. Macaskie. Biotechnol. Letters. 23 (2001) 61. C. Cervantes, J. Campos-Garcia, S. Devars, F. Gutierrez-Corona, H. Loza-Tavera, J. Torres-Guzman, R. Moreno-Sanchez. FEMS Microbiology Reviews. 25 (2001) 335. 7. E. Chirwa, Y. Wang. J. Environ. Eng.-ASCE. 123 (1997) 760. 8. P, DeLeo, H. Ehrlich. Appl. Microbiol. Biotechnol. 40 (1994) 756. 9. A. Llovera, R. Bonet, M. Simonpujol, F. Congregado. Appl. Environ. Microbiol. 59 (1993) 3156. 10. K. Komori, A. Rivas, K. Toda, H. Ohtake. Biotechnol. Bioeng. 35 (1990) 951. 530
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11. P. Wang, T. Mori, K. Toda, H. Ohtake. J. Bacteriol. 172 (1990) 1670. 12. D. Lovley. Annu. Rev. Microbiol. 47 (1993) 263. 13. D. Lovley, E. Phillips. Appl. Environ. Microbiol. (1994) 726. 14. J. Lloyd, P. Yong, L. Macaskie. Appl. Environ. Microbiol. 64 (1998) 4607. 15. J. Lloyd, V. Nolting, V. Sole, K. Bosecker, L. Macaskie. Geomicrobiology Journal. 15 (1998) 45. 16. A. N. Mabbett, J. Lloyd, L. E. Macaskie. Biotech. Bioeng. 79 (2002) 389. A. C. Humphries, L. E. Macaskie. Biotechnol. Letters. 24 (2002) 1261. 17. J. Lloyd, A. N. Mabbett, D. Williams, L. E. Macaskie. Hydrometallurgy. 59 (2001) 327. 18. B. White, S. Wilkinson, G. Gadd. Appl. Environ. Microbiol. (1994) 726. 19. J. Postgate. The sulphate-reducing bacteria. (1979). Cambridge University Press. 20. A. Lopez, N. Lazaro, A. Marques. Journal of Microbiological Methods. 30 (1997) 231.
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Clean-up of mine waters from a uranium deposit by means of a constructed wetland S.N. Groudeva, K. Komnitsasb, I.I. Spasovaa and I. Paspaliarisb a
Department of Engineering Geoecology, University of Mining and Geology, Studentski grad – Durvenitza, Sofia 1700, Bulgaria b Laboratory of Metallurgy, National Technical University of Athens, Zografos 15780, Greece
Abstract Mine drainage waters from a uranium deposit located in Western Bulgaria contained uranium, radium, manganese and sulphates as main pollutants. pH of these waters was in the range of 6.0-7.1 and the concentrations of the above-mentioned pollutants usually were about 5-10 times higher than the relevant permissible levels for waters intended for use in the agriculture and industry. The main objective of this study was to test the possibility to clean-up these waters by means of a passive system, the constructed wetland. The wetland was a pond with a rectangular shape and was 8 m long, 4 m wide and 0.7 m deep. The pond was isolated from the ground by an impermeable plastic sheet. The bottom of the pond was covered by a 0.35 m layer consisting of a mixture of soil with a high organic content, plant and spent mushroom compost, cow manure, silt and sand. Several permeable soil barriers were constructed perpendicularly to the direction of the water flow. The wetland was characterized by an abundant water and emergent vegetation and a diverse microflora. Phragmites australis, Juncus bulbosus and different algae were the prevalent plant species in the wetland. The water flow rate varied in the range of approximately 3-9 m3/24 h. The water clean-up in the wetland was very efficient during the different climatic seasons, even during the cold winter months at water temperatures approximately 2-4°C. The dissolved hexavalent uranium was removed mainly as a result of its reduction to the tetravalent state by the hydrogen sulphide produced by the anaerobic sulphate-reducing bacteria. The removal of manganese was connected with the bacterial oxidation of the bivalent manganese to the tetravalent state. The Mn4+ was then precipitated mainly as MnO2. However, a portion of Mn2+ was precipitated as MnS and Mn(OH)2 as a result of the microbial dissimilatory sulphate reduction. Portions of uranium and manganese as well as most of the radium were removed by their sorption on the plant and microbial biomass and on the clay minerals in the wetland. Negative effects of the pollutants on the plant and microbial communities in the wetland were not observed. 1.
INTRODUCTION The uranium deposit B-1 located in Western Bulgaria for a long period of time was a site of intensive mining activities connected with underground recovery of uranium ores. 533
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The mining operations in the deposit were completed in 1990 due to a complex of political, economical and environmental reasons. However, some years after the closure of the underground mine the level of the groundwater reached the ground surface. These waters had a pH close to neutral and contained uranium, radium, manganese and sulphates as main pollutants. The concentrations of these pollutants usually were approximately 510 times higher than the relevant permissible levels for waters intended for use in the agriculture and industry. Pumping out of these waters as well as other measures to prevent their discharge into the environment were high-cost operations. Apart from the abovementioned waters, polluted waters with similar composition arose after rainfall as effluents from some dumps consisting of mining wastes and low-grade ores. The clean up of waters containing the above-mentioned pollutants is a complex and costly operation due to the entirely different biogeochemical cycles of the uranium, radium and manganese, reflecting the specific properties of these elements. The most widely used technologies are connected with an initial removal of uranium by ion exchange, followed by precipitation of radium as radium-barium sulphate and oxidative treatment to precipitate manganese as MnO2. Since the uranium mine was closed and there was no commercial interest in the abandoned deposit, the only alternative of the abovementioned high-cost schemes are passive treatment systems. These systems have been developed on the basis of naturally occurring biological and geochemical processes in order to improve the quality of the influent waters with minimal operation and maintenance costs (Cambridge, 1995; Gusek, 1995; Younger, 1997). Such systems, mainly natural and constructed wetlands, anoxic cells and rock filters, have been efficiently used to treat waters polluted with radioactive and heavy metals (Groudev et al., 1999; Groudev at al., 2001a; Groudev et al., 2001b; Groudev et al., 2001c). The main objective of this study was to test the possibility to clean-up the polluted waters from the B-1 uranium deposit by means of a passive system, the constructed wetland. Some data with regards to this study are presented in this paper. 2.
MATERIALS AND METHODS The wetland was a pond with a rectangular shape and was 8 m long, 4 m wide and 0.7 m deep. The pond was isolated from the ground by an impermeable plastic sheet. The bottom of the pond was covered by a 0.35 m layer consisting of a mixture of soil with a high organic content, plant and spent mushroom compost, cow manure, silt and sand. Several permeable barriers consisting of the above-mentioned materials were constructed perpendicularly to the direction of the water flow. The wetland was characterized by abundant water and emergent vegetation and a diverse microflora (Table 1). Phragmites australis, Juncus bulbosus and different algae were the prevalent plant species in the wetland but species of the genera Typha, Eleocharis, Carex and Poa were also present. The water flow rate varied in the range of 3-9 m3/24 h which reflected residence times of approximately 40-120 h. The water treatment was carried out during the different climatic seasons. The parameters measured in situ included pH, Eh, dissolved oxygen, total dissolved solids and temperature. Elemental analysis was done by atomic absorbtion spectrophotometry and induced plasma spectrophotometry in the laboratory. The radioactivity of the samples was measured, using the solid residues remaining after their evaporation, by means of a low background gamma-spectrophotometer ORTEC (HpGe-detector with a high distinguishing ability). The specific activity of Ra226 was measured using a 10 l ionization 534
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chamber. The total β-activity was measured by a low background instrument UMF1500M. Table 1. Microflora of the mine waters and the constructed wetland Samples Microorganisms
Aerobic heterotrophic bacteria Cellulose – degrading aerobes Fe2+ - oxidizing chemolithotrophs (at pH 2) Fe2+ - oxidizing heterotrophs (at pH 7) Mn2+- oxidizing heterotrophs (at pH 7) S2O32- - oxidizing chemolithotrophs (at pH 7) Anaerobic heterotrophic bacteria Cellulose-degrading anaerobes Sulphate-reducing bacteria Fe3+ - reducing bacteria Mn4+ - reducing bacteria
Mine waters before treatment
Waters from the wetland Cells/ml (g)
Sediments from the wetland
101 – 104
104 – 108
102 – 105
0 – 102
102 – 106
0 – 102
0 – 102
0 – 102
0 – 101
1 – 102
103 – 106
0 – 103
0 - 102
102 - 106
0 - 103
104 – 107
104 – 107
101 – 104
0 – 102
102 – 106
103 –107
0–1
101 - 103
102 - 105
0 – 103 0 - 102 0 – 101
102 – 106 1 - 102 0 - 102
103 – 107 102 - 104 102 - 104
Elemental analysis of solid samples from the bottom sediments and the plant biomass in the wetland was performed by digestion and measurement of the ion concentration in solution by atomic absorption spectrophotometry and induced coupled plasma spectrophotometry. Mineralogical analysis was carried out by X-ray diffraction techniques. The mobility of the pollutants was determined by the sequential extraction procedure (Tessier et al., 1979). Sediment samples were subjected to microbial leaching, using indigenous microorganisms from the wetland, under conditions previously described (Groudev et al., 2001a). The isolation, identification and enumeration of microorganisms were carried out by methods described elsewhere (Karavaiko et al., 1988; Groudeva et al., 1993). 3.
RESULTS AND DISCUSSION The water clean up in the wetland was very efficient during the different climatic seasons, even during the cold winter months at water temperatures lower than 5 °C (Tables 2 and 3). The significant decrease of the concentration of sulphate ions, the generation of hydrogen sulphide, the low levels of the redox potentials in the bottom zone of the wetland (usually lower than -150 mV), the presence of dissolved biodegradable organic compounds and the increase of the other metabolically interdependent 535
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microorganisms were indications that the microbial dissimilatory sulphate reduction was an essential process involved in the water clean up. Table 2. Data about the mine waters before and after their treatment by the constructed wetland Parameters Temperature, °C pH Eh, mV Dissolved O2, mg/l
Before treatment
After treatment
Permissible levels for waters used in agriculture and industry
(+0.7) – (+17.2)
(+0.5) – (+21.5)
-
6.0 - 7.1
7.1 - 7.7
6-9
(+224) – (+352)
(+178) – (+268)
-
1.7 – 5.1
2.5 – 5.5
2
Total dissolved solids, mg/l
1070 - 2015
808 - 1402
1500
Solids, mg/l
44 - 125
28 - 77
100
Sulphates, mg/l
352 - 1630
181 - 590
400
Uranium, mg/l
1.9 – 5.9
< 0.1
0.6
Radium, Bq/l
0.45 – 1.40
< 0.1
0.15
Total β-activity, Bq/l
1.05 – 2.50
< 0.5
0.75
Manganese, mg/l
2.8 – 17
0.1 – 0.7
0.8
Iron, mg/l
28 - 122
0.2 – 1.4
5
464 - 1252
215 – 440
400
0.9 – 3.2
10 – 25
20
Sulphates, mg/l Organic carbon, mg/l
Table 3. Removal of pollutants from the mine waters in the constructed wetland Pollutant Uranium Manganese Iron Sulphates
Pollutant removed, g/24 h During the cold winter months During the warmer months (at 0-5°C) 9 – 32 6 - 15 19 – 95 10 - 32 204 – 680 104 - 305 1250 – 4240 840 - 1285
The alkalinity produced during the microbial sulphate reduction in the form of hydrocarbonate ions as well as the alkalinity produced by the solubilization of some alkalizing minerals (mainly carbonates) present in the wetland gradually increased the pH of the waters. Furthermore the pH was subjected to fluctuations (usually as high as 0.5-0.7 units) during the twenty-four-hour period connected with the activity of the plants and phototrophic bacteria inhabiting the wetland. The consumption of CO2 for photosynthesis during the day increased the pH, while its secretion during the night as a result of the respiration had the reverse effect. The content of dissolved oxygen in the waters also was subject to similar fluctuations. The highest values of these two parameters were measured in the afternoon (usually at 4-6 pm). The chemical and mineralogical analysis of the sediments precipitated on the bottom of the wetland and on the submerged parts of the water plants revealed that in the sediments formed during the warmer months of the year a significant portion of the iron 536
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was present as the relevant sulphide (FeS), and uranium was present mainly in its tetravalent form (mainly but not only as the mineral uraninite). The easily soluble fractions (exchangeable and carbonate) of iron and uranium in these sediments were smaller than the refractory fractions (Table 4). Table 4. Mobility fractions of pollutants in sediments from the constructed wetland Pollutant Uranium I Uranium II Radium I Radium II Manganese I Manganese II Iron I Iron II
Exchangeable 17 – 44 35 – 79 71 – 88 73 - 91 8 – 32 46 – 77 7 – 25 44 – 80
Mobility fractions, % Carbonate Oxidisable Reducible 3 –10 53 – 82 1–7 8 – 16 8 – 23 5 – 14 5 - 12 3-7 7 – 18 5 - 15 2-4 8 - 21 6 – 12 1-4 59 - 84 9 - 21 1–2 14 – 32 4 – 14 1–5 53 – 86 8 – 21 1–4 16 – 41
Inert 0-5 3-7 1- 4 5-9 1-5
Notes: I - Samples taken during the warmer months; II – Samples taken during the cold winter months (at 0-5°C).
The sediments were refractory to leaching by mixed cultures of indigenous heterotrophic microorganisms at about neutral pH and under anaerobic conditions which were typical for the bottom zone of the wetland. The resistance of uranium to leaching under such conditions was connected with the inability of the microorganisms to produce sufficient quantities of peroxide compounds for the oxidation of the tetravalent uranium to the soluble hexavalent state. This, in turn, was due to the low rate of degradation of the solid organic substrates by the anaerobic cellulose-degrading microorganisms. This slow biodegradation limited the content of dissolved organic monomers, which could be used as substrates for the microbial production of peroxide compounds in the leach system. The inefficient solubilization of iron from FeS was due to the same reason. The chemical leaching by diluted sulphuric acid solutions (with pH in the range of 2-4) was also not efficient. However, the leaching by means of mixed cultures of acidophilic chemolithotrophic bacteria (Acidithiobacillus ferrooxidans, A. thiooxidans and Leptospirillum ferrooxidans) in well-aerated mineral suspensions and at pH approx. 2.02.5 resulted in efficient solubilization of iron and uranium. The above-mentioned results from the leaching experiments confirmed the results from the mineralogical analysis about the mineral forms of iron and uranium in the sediments. It must be noted that the environmental conditions in the bottom zone of the wetland are not favorable for the growth and activity of the acidophilic chemolithotrophs. In freshly precipitated sediments during the warmer months the relative portions of the easily soluble fractions of the pollutants were much higher than those in the older sediments. This was due to the fact that portions of the uranium, iron and manganese as well as most of the radium were removed by their sorption on the plant and microbial biomass (Table 5) and on the inorganic components (mainly on the clays) in the wetland. Different specimens of the same plant species grown in the wetland differed considerably from each other with respect to their contents of pollutants, regardless of the fact that they had been grown under similar conditions. The highest concentrations of pollutants were detected in the root systems of the plants. The concentrations in the leaves and stems were much lower. The pollutants were taken up as the relevant ions but also as fine solid products (such as sulphides, hydroxides and oxides) from different processes, mainly from 537
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the microbial sulphate reduction and the bacterial oxidation of Fe2+ and Mn2+. Some algae (mainly such related to the genera Scenedesmus, Microcystis, Anabaena and Nitela) and microorganisms (mainly related to the genera Aspergillus, Penicillium and Pseudomonas) also adsorbed pollutants. The content of uranium in some specimens of algae was as high as 320 mg/kg dry biomass, and that of the radium exceeded 500 Bq/kg. Table 5. Content of pollutants in living and dead plant biomass from the constructed wetland Pollutant Uranium, mg/kg dry Radium, Bq/kg dry Manganese, mg/kg dry
Phragmites australis I II 107 – 375 125 – 640 120 – 855 135 – 1070 120 – 493 95 – 710
Juncus bulbosus I II 70 – 244 91 - 315 62 – 510 105 - 700 84 – 357 125 - 482
Notes: I - Living biomass; II – Dead biomass taken during the winter months (at 0-5°C).
Negative effects of the pollutants on the indigenous plant community were not observed during the 3-year experimental period (from July 1998 to October 2001). The removal of manganese from the polluted waters was connected mainly with the microbial oxidation of the bivalent manganese to the tetravalent state. The Mn4+ was then precipitated mainly as MnO2. The oxidation of Mn2+ was carried out by heterotrophic bacteria producing hydrogen peroxide and the enzyme catalase. These bacteria were found mainly in the upper aerobic zone of the wetland. The most active isolates were related to the genera of Metallogenium, Leptothrix and Arthrobacter. A significant portion of the MnO2 accumulated on the cell surface of these bacteria. It must be noted, however, that portions of the bivalent manganese were precipitated as MnS and, in some niches with a relatively higher pH (>10) within the bottom anaerobic zone, in the form of MnCO3 and Mn(OH)2, i.e. these precipitates were formed as a result of the microbial sulphate reduction and the microbial and chemical alkalization of the waters. A portion of the iron was also removed as a result of the prior oxidation (mainly biological) of the ferrous ion to the ferric state. The Fe3+ was then precipitated as Fe(OH)3. Different heterotrophic bacteria able to oxidize Fe2+ at slightly acidic, neutral and slightly alkaline pH (from 4.5 to 7.5) were found in the wetland, including the isolates related to Metallogenium, Leptothrix and Arthrobacter, which were able to oxidize also the bivalent manganese. It must be noted that the chemical oxidation of Fe2+ by O2 was possible at the above-mentioned values but proceeded at lower rates than the bacterial oxidation. Chemolithotrophic bacteria able to oxidize Fe2+ at low pH values (less than 4) were found in the wetland but in very low numbers (Table 1). They were related to the species Acidithiobacillus ferrooxidans and Leptospirillum ferrooxidans. The growth and activity of the indigenous microflora during the cold winter months (December – February) of the year, when the temperature of the waters in the wetland was usually less than 5°C and often about 0°C, were markedly inhibited. However, the removal of pollutants was efficient (Table 3) although the residence times were increased. The character of the sediments precipitated during this time of the year in most cases was similar to that of the freshly precipitated sediments during the warmer months. The pollutants were present mainly as the easily soluble exchangeable and carbonate fractions (Table 4). The contents of pollutants in the dead plant biomass were much higher then these in the living plants during the warmer months. The contents of pollutants in the clays present in the wetland steadily increased during the winter months. Some clay specimens contained more than 1.5 g uranium per kg dry clay and the content of manganese exceeded 538
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12 g per kg (the initial content of manganese in this clay was less than 5 g/kg). These data revealed that the role played by some sorbents, mainly by dead plant biomass and clays, in the removal of pollutants was essential during the cold winter months. However, the decrease of the concentrations of sulphates and the relatively high number of viable sulphate-reducing bacteria in the soil layer at the bottom of the wetland were indications that the microbial sulphate reduction proceeded, although at lower rates, in this anaerobic zone where the temperature was higher than that of the upper water layer. The results obtained during this study showed that the treatment of waters polluted with radioactive elements and heavy metals can be efficiently carried out by constructed wetlands with a proper size and design, with suitable and well-developed plant and microbial communities. ACKNOWLEDGEMENTS Parts of this work were financially supported by the European Commission under COPERNICUS project No ICA2-CT-2000-10010. REFERENCES
1. Cambridge, M., (1995). Use of passive systems for the treatment of mine outflows and seepage, Minerals Industry International, May: 35-42. 2. Groudev, S.N., Georgiev, P.S., Komnitsas, K., Spasova, I.I. and Angelov, A.T., (1999). Treatment of waters contaminated with radioactive elements and toxic heavy metals by a natural wetland, Paper presented at the International Conference on Wetlands Remediation, Salt Lake City, Utah, USA, November 16-17, (1999). 3. Groudev, S.N., Georgiev, P.S., Spasova, I.I, Angelov, A.T. and Mitrov, T., (2001a). Biotechnological treatment of mine waters contaminated with radioactive elements, In: New Developments in Mineral Processing, G. Onal, S. Atak, A. Guney, M.S. Celik and A.E. Yűce, Eds., pp. 571-574, Beril Ofset, Istanbul. 4. Groudev, S.N., Komnitsas, K., Spasova, I.I., Georgiev, P.S. and Paspaliaris, I., (2001b). Contaminated sediments in a natural wetland in a uranium deposit, Paper presented at the Sediments Conference 2001, Battelle Geneva Research Centre, Venice, October 10-12, (2001). 5. Groudev, S.N., Nicolova, M.V., Spasova, I.I., Komnitsas, K. and Paspaliaris, I., (2001c). Treatment of acid mine drainage from an uranium deposit by means of a natural wetland, Paper presented at the ISEB 2001 Meeting on Phythoremediation, Leipzig, Germany, May 15-17, (2001). 6. Groudeva, V.I., Ivanova, I.A., Groudev, S.N. and Uzunov, G.C., (1993). Enhanced oil recovery by stimulating the activity of the indigenous microflora of oil reservoirs. In: Biohydrometallurgical Technologies, vol. II, A.E, Torma, M.L. Apel and C.L. Brierley, Eds., pp. 349-356, TMS, The Minerals, Metals & Materials Society, Warrendale, PA. 7. Gusek, J.J. (1995). Passive-treatment of acid rock drainage: What is the potential bottom line?, Mining Engineering, March: 250-253. 8. Karavaiko, G.I., Rossi, G., Agate, A.D., Groudev, S.N. and Avakyan, Z.A., (1988). Biogeotechnology of Metals. Manual, GKNT International Projects, Moscow. 9. Tessier, A., Campbell, P.G.C. and Bisson, M., (1979), Sequential extraction procedure for speciation of particulate trace metals, Analytical Chemistry, 51(7): 844-851. 10. Younger, P.L., Ed., (1997). Mine Water Treatment Using Wetlands, University of Newcastle, Newcastle, UK. 539
15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Degradation of tetracyanonickelate (II) by Cryptococcus humicolus in biofilm reactors Hyouk Kee Kwona, Seung Han Woob, Joon Yong Sunga and Jong Moon Parkb a
LG Institute of Environment Safety & Health, 134 Shinchon-Dong, Seodaemoon-Gu, Seoul 120-749, Korea b Department of Chemical Engineering, Division of Molecular and Life Sciences, Pohang University of Science and Technology (POSTECH), San 31, Hyoja-dong, Nam-gu, Pohang 790-784, Korea Abstract A new yeast capable of degrading high concentrations of a metallocyanide, TCN (tetracyanonickelate (II), K2[Ni(CN)4]), was isolated from coke-plant wastewater and identified as Cryptococcus humicolus. The strain was able to degrade cyanide compounds in both aerobic and anaerobic conditions. The highest degradation activity was achieved at pH 7.5 in aerobic condition and at pH 5.0 in anaerobic condition. The feasibility of biological detoxification of TCN-containing synthetic wastewater by the strain was investigated using two fixed-bed biofilm reactors operated in the aerobic and anaerobic conditions. At 24 h of hydraulic retention time, influent TCN of 50 mg CN dm-3 was completely degraded in an aerobic reactor and about 90% of the TCN was degraded in an anaerobic reactor. When the hydraulic retention time decreased to 9 h, the removal efficiency decreased to 83% and 61% for the aerobic and anaerobic reactors, respectively. Keywords: biodegradation, biofilm, bioreactor, cyanide, metallocyanide 1.
INTRODUCTION Cyanide compounds are mainly generated from various chemical industries, including metal extraction, electroplating, coal gasification, ore leaching, and production of synthetic fibers [1]. These compounds are of major environmental concern due to their toxicity to living organisms [2]. In metal-bearing wastewater, most of cyanides are present in the form of metal (such as Fe, Cu, Ni, and Zn) complexes rather than as free cyanide. Such metal-cyano complexes are highly stable and more resistant to biological attack compared with free cyanide. Biodegradation of cyanides has been studied with a number of microorganisms, including bacteria [3-5] and fungi [6-10]. Most of the microorganisms have been investigated in aerobic conditions to apply for conventional activated sludge processes. On the contrary, anaerobic treatment of cyanides has not been intensively studied except for the cases of some methanogenic mixed cultures [11, 12]. Anaerobic wastewater treatment could be more advantageous due to its low energy requirement compared with aerobic treatment. 541
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In our previous report, for the first time, a new yeast strain (Cryptococcus humicolus) capable of degrading high concentrations of metallocyanide compounds was isolated from wastewater [13]. The strain was able to utilize free cyanide as a single nitrogen source by converting CN to ammonia and carbon dioxide in aerobic conditions. It was also elucidated that molecular oxygen from O2 was not directly incorporated during the conversion of the free cyanide. The ability of facultative degradation is greatly beneficial for practical applications since gradients of oxygen concentrations in microenvironment can seriously inhibit the biodegradability in the system requiring the strict aerobic or anaerobic condition. The aim of the present work was to investigate the treatability of metallocyanides in both aerobic and anaerobic conditions using the strain Cryptococcus humicolus. A ceramic "honey-comb" structure was used as support matrix for the biofilm in fixed-bed reactors. 2.
MATERIALS AND METHODS
2.1 Microorganism and growth conditions Cryptococcus humicolus was grown in an enrichment medium containing 3 g glucose, 5 g KH2PO4, 4.2 g K2HPO4.3H2O, 0.5 g MgSO4, 0.04 g FeSO4.7H2O, 0.0015 g MnSO4.H2O per liter. The pH was adjusted to 7.5. The strain was maintained on mineral agar medium containing TCN of 260 mg CN dm-3 by subculturing at 7-day interval. Colonies on the agar medium were cultivated in a 500 ml Erlenmeyer flask containing 0.1 dm3 mineral medium and TCN of 260 mg CN dm-3 for 3 days in a rotary shaker at 250 rpm. Ten ml of culture was harvested by centrifugation (6,000 x g) and washed twice with mineral medium. This was resuspended in 0.01 dm3 of mineral medium and used as inoculum for biodegradation experiments. 2.2 Effect of pH The effect of initial pH ranging from 4.0 to 8.5 on TCN degradation was examined in flask cultures under aerobic and anaerobic conditions. Initial concentration of TCN was 260 mg CN dm-3. Anaerobic condition was accomplished by bubbling high purity N2 gas. During culture period, pH and TCN concentration were measured with time. 2.3 Fixed-bed biofilm reactor The schematic diagrams of honey-comb typed bioreactors and their dimensions are shown in Fig. 1 and Table 1, respectively. The influent was introduced into the bottom of the reactors by a peristaltic pump. For the aerobic bioreactor, air was sparged at the bottom. A ceramic structure (chemical composition: 2MgO.2Al2O3.5SiO2) was used as a support matrix for biofilm formation. One unit of structure contains 625 pixels and each pixel is a regular hexahedron void with a side length of 0.5 cm. 2.4 Reactor operation Inoculum was prepared by cultivation of C. humicolus in a fermenter using 4 dm3 enrichment medium. Each reactor was filled with the microbial inoculum prepared in the fermenter and synthetic wastewater. The synthetic wastewater contained 1 g of glucose and TCN (52 mg CN dm-3 tap water). The reactors were operated in a continuous mode at HRT (hydraulic retention time) of 24 h during 3 months before reaching steady state. For the aerobic reactor, the aeration rate was controlled at 3 dm3.min-1. The TCN biodegradation was measured at various HRT values and aeration rates. 542
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Figure 1. Schematic diagrams of fixed-bed biofilm reactors Table 1. Summary of the dimensions of the fixed-bed biofilm reactor Items
Values
Inner diameter of the reactor Total volume of the reactor Working volume Height of the unit matrix Diameter of the unit matrix Void fraction of the matrix Hydrous water content Number of the matrices per reactor
15 cm 8.3 dm3 6.2 dm3 10 cm 15 cm 0.66 0.16 cm3 water / g-matrix 4
2.5 Liquid circulation Liquid circulation time (tc, min) was measured using hydrogen ion as a tracer at various aeration rates ranging from 3 to 8 dm3.min-1. Using the liquid circulation time and the liquid volume (VL, dm3), the liquid circulation rate (Qc, dm3.min-1) can be calculated as follows: V − (1 − ε )VB V (1) Qc = L = w tc tc
where ε is the void fraction of the bed, Vw (dm3) is the working volume, and VB (dm3) is the bed volume. Pressure drop between the top and the bottom of the bed was also measured by a manometer at various aeration rates. 2.6 Analytical methods The samples were filtered through a 0.2 µm pore size filter before analyses. TCN was analyzed by the silver nitrate titration method after distillation according to APHA standard methods [14]. Ammonia was determined by Nesslerization method [14]. Glucose concentration was determined using o-toluidine reagent kit (Sigma, USA). Dissolved 543
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oxygen and pH value were measured by electrodes (Ingold, USA). Cell growth was determined by measuring the optical density of culture broth samples at 600 nm (OD600). Microbial cell dry weight was determined by the value obtained after centrifugation (6,000 x g for 10 min) and drying at 80°C for 24 h. 3.
RESULTS AND DISCUSSION
3.1 Effect of pH on TCN degradation Figs. 2 and 3 show the effect of the initial pH on the degradation of TCN under both aerobic and anaerobic conditions. The optimal pH values for TCN degradation in aerobic and anaerobic conditions were quite different. While the highest degradation activity was achieved at pH 7.5 in the aerobic condition, pH 5.0 was the optimum in the case of anaerobic condition. In the aerobic condition, the TCN degradation rate did not show any significant difference in the range of 6.5 to 7.5. However, in the anaerobic condition, strong inhibition was observed above pH 7.0 as shown in Fig. 3. For all the tests, pH values did not change significantly with time (data not shown). As expected, the biodegradation rate was generally higher in aerobic condition than in anaerobic one. However, below pH 5.0, the aerobic growth of the strain C. humicolus was significantly inhibited (data not shown), while high degradation activity was observed in anaerobic conditions. It suggests that anaerobic treatment of cyanides using this strain could be beneficial for acidic wastewater compared to aerobic treatment. 300 300
200 150
TCN (mg dm-3)
-3
TCN (mg dm )
250
pH 5.5 pH 6.0 pH 6.5 pH 7.0 pH 7.5 pH 8.5
250
100
200 150 pH 4.0 pH 4.5 pH 5.0 pH 6.0 pH 6.5 pH 7.0
100 50
50
0
0 0
20
40
60
80
Time (h)
Figure 2. Effect of initial pH on TCN degradation under aerobic condition
0
20
40
60
80
Time (h)
Figure 3. Effect of initial pH on TCN degradation under anaerobic condition
3.2 Liquid circulation The aerobic reactor was operated as an air-lift with a riser part formed by aeration in the center of the reactor. This brought about repeated sinusoidal detection of pH in the trace test for measurement of liquid circulation time. The liquid circulation time in the aerobic reactor was measured at various aeration rates as shown in Fig. 4. The liquid circulation rate calculated by equation (1) was proportional to the aeration rate. Initially, no pressure drop was detectable. After 3-months operation, the pressure drop started to build up around 2.5 cm H2O due to the biofilm formation within the bed. The overgrown biomass was removed from the reactor when the pressure drop increased over 5.0 cm H2O. 544
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Figure 4. Liquid circulation characteristics of the bioreactor at various aeration rates 3.3 Effect of HRT on TCN degradation Figs. 5 and 7 show the TCN degradation by C. humicolus in aerobic and anaerobic reactors at various HRTs. In the aerobic operation, TCN was completely degraded at 24 h of HRT. Decrease of HRT resulted in the gradual decrease of TCN removal efficiency. For example, at 9 h of HRT, the removal efficiency dropped to 83.2%. Ammonia generated from TCN degradation remained in the effluent after being used as a nitrogen source for microorganisms. The ammonia production rate was almost constant at 0.4 mg NH3-N dm-3 h-1 regardless of HRT. The pH increased from 6.7 to 7.4 with increasing HRT from 9 h to 24 h, respectively. This is possibly due to the higher production of ammonia by TCN degradation at higher HRT. The effect of aeration rate on TCN degradation in the aerobic operation was investigated at various HRTs (9, 11, 14, 18 and 24 h). The efficiency of TCN degradation increased proportionally with increasing aeration rate as expected (Fig. 6). However, above 7 dm3.min-1, the degradation rate was not further enhanced. Moreover, TCN was not completely degraded below 18 h of HRT despite high aeration rate.
Figure 5. Effect of HRT on TCN degradation in aerobic operation at an aeration rate of 7 dm3.min-1
545
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Figure 6. Effect of aeration rate on effluent TCN at various HRTs in the aerobic reactor
Figure 7. Effect of HRT on TCN degradation in anaerobic operation
In the anaerobic operation, the removal efficiency of TCN was generally lower than in the aerobic operation. In addition, the decrease of removal efficiency by decreasing HRT in anaerobic condition was greater than under aerobic conditions. The removal efficiency dropped from 90% to 61.1% with decreasing HRT from 24 h to 9 h, respectively. In the anaerobic operation, ammonia was constantly generated at a rate of about 0.1 mg NH3-N dm-3.h-1 regardless of HRT, which was lower than that of the aerobic operation. Unlike the aerobic operation, the pH value decreased slightly from 5 to 4.4 with increasing HRT from 9 h to 24 h, respectively. This is thought to be due to the acid formation by fermentation of glucose in the anaerobic conditions. 3.4 TCN profile with bed depth Fig. 8 shows the TCN profile along the bed depth for the aerobic and the anaerobic operation. The dissolved oxygen values were zero at all the depths in the anaerobic reactor, and 7.6, 7.5, 7.4 and 7.2 at D1, D2, D3 and D4 in the aerobic reactor, respectively. In the anaerobic operations, TCN concentration gradually decreased as wastewater passed through the bed. However, in the aerobic operation, most of influent TCN was degraded within the first unit matrix and thereafter TCN was maintained almost constant. This is likely due to rapid liquid mixing in the aerobic reactor by the rigorous aeration. 546
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Figure 8. Concentration profile of TCN along the depth of bed with HRT of 24 h in the anaerobic or aerobic reactor at air flow rate of 7 dm3.min-1 4.
CONCLUSIONS A new yeast strain, Cryptococcus humicolus, showed facultative property on TCN (tetracyanonickelate (II)) degradation, which is able to grow in both aerobic and anaerobic conditions. This facultative ability can be beneficial for practical applications since oxygen limitation in aerobic operation or the presence of oxygen in anaerobic operation is not detrimental to cell activity. Furthermore, in the acidic condition, TCN degradation rate in the anaerobic condition was comparable to that in the aerobic condition. Using fixedbed biofilm reactors, at 24 h of hydraulic retention time, influent TCN of 50 mg CN dm-3 was completely degraded in the aerobic operation and about 90% of the TCN was degraded in the anaerobic operation. The removal efficiency was not seriously decreased even when the hydraulic retention time decreased to 9 h in both aerobic and anaerobic operation. If the performance of anaerobic reactor is further optimized, it can be a competitive technology for the treatment of cyanide compounds in wastewater. REFERENCES
1. S.A. Raybuck, Biodegradation, 3 (1992) 3. 2. C.J. Knowles and A.W. Bunch, Adv. Microb. Physiol., 27 (1986) 73. 3. G. Rollinson, R. Jones, M.P. Meadows, R.E. Harris and C.J. Knowles, FEMS Microbiol. Lett., 40 (1987) 199. 4. J. Silva-Avalos, M.G. Richmond, O. Nagappan and D.A. Kunz, Appl. Environ. Microbiol., 56 (1990) 3664. 5. R. Harris and C.J. Knowles, J. Gen. Microbiol., 129 (1983)1005. 6. M.J. Cluness, P.D. Turner, E. Clements, D.T. Brown and C. O’Reilly, J. Gen. Microbiol., 139 (1993) 1807. 7. P. Wang, D.E. Matthews and H.D. VanEtten, Arch. Biochem. Biophys., 287 (1992) 569. 8. Dumestre, T. Chone, J. Portal, D. Gerard and J. Berthelin, Appl. Environ. Microbiol., 63 (1997) 2729. 9. M. Barclay, V. Tett and C.J. Knowles, Enzyme. Microb. Tech., 23 (1998) 321. 10. H. Yanase, A. Sakamoto, K. Okamoto, K. Kita and S. Sato, Appl. Microbiol. Biotechnol., 53 (2000) 328. 547
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11. P.M. Fedorak and S.E. Hrudey, Wat. Sci. Tech., 21 (1987) 67. 12. R.D. Fallon, D.A. Cooper and M. Henson, Appl. Environ. Microbiol., 57 (1991) 1656. 13. H.K. Kwon, S.H. Woo and J.M. Park, FEMS Microbiol. Lett., 214 (2002) 211. 14. A.E. Greenberg, R.R. Trussell and L.S. Clesceri (eds.), Standard Methods for the Examination of Water and Wastewater, American Public Health Association, Washington, 1995.
548
15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Development of a bio-process using sulfate-reducing bacteria to remove metals from surface treatment effluents F. Battaglia-Bruneta, S. Fouchera, A. Denamura, S. Chevardb, D. Morina, I. Ignatiadisa a
BRGM, Environment and Process Division, Biotechnology Unit, Av. Claude Guillemin, 45060 Orléans Cedex 02, France b SETS S. A., ZI Les Vigneaux, B.P. 3, 36210 Chabris, France
Abstract The surface treatment activities are essential and associated to any industrial sector. However, they produce large quantities of effluents. The majority of these pollutants are currently removed through physical and chemical treatments, and metals are generally precipitated as hydroxides. The present project proposes to replace some steps of the classical treatment process by biological reactions. Sulfate-reducing bacteria (SRB) would perform the bio-reducing operations. The objective of this work was to reduce sludge volume and toxicity, and if possible to recover some metals selectively. Three processes involving sulfate-reducing bacteria were considered: (1) the sulfate produced by oxidation of sodium bisulfite, may be converted into H2S in a bioreactor, in which the acidic/alkaline metals-containing effluents would be directly injected; (2) the sulfate produced by oxidation of bisulfite may be injected in a SRB-bioreactor producing H2S that would be transferred in a separate precipitation reactor, in order to selectively recover metals; (3) the Cr(VI)-containing effluents may be directly treated in the bio-reactor, that would also produce some H2S to precipitate the other metals in a separate reactor. This last configuration was tested in a 20-L column bioreactor, fed in continuous mode with a real effluent, whose Cr(VI) concentration was adjusted at 30 mg/l. 1.
INTRODUCTION Surface treatment is a widespread industrial activity, necessary to many important economical sectors such as transports, building-trade. The expenses of environmental protection in surface treatment plants represent 25% of the investment and 10-15% of the operating costs [1]. In spite of significant improvement of the practices, the contamination of the environment due to surface treatment industries reaches 30% of the global industrial toxic discharge [1]. Metals-containing effluents are generally treated by physical and chemical processes generating voluminous amounts of metal-hydroxide sludge. The charge for sludge removal may reach 30% of the effluent treatment operating costs. Consequently, any improvement of the flowsheet that could reduce the volume and toxicity of the solid residues should be considered. The present study proposes to introduce a biological step in the process that could replace a fraction of the hydroxide sludge by metal sulfide precipitates, less voluminous and easier to dry. 549
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2.
DATA ON THE INDUSTRIAL PLANT
2.1 Flowsheet of the effluent treatment plant at S.E.T.S (Fig. 1) The real industrial plant chosen for the present study is equipped with a classical and common effluent treatment plant. The three main types of effluents are cyanide rinse solution (4 m3.h-1), acid/alkaline rinse solution (10 m3.h-1) and chromate rinse solution (4 m3.h-1). Cyanides are oxidized with hypochloride (ClO-) in alkaline conditions. Chromate is reduced with sodium bisulfite (NaHSO3) in acidic conditions. The effluents from cyanide oxidation and chromate reduction are conducted to the neutralization tank, which also receives acid/alkaline bathes. The precipitation of metal hydroxides is performed in this reactor. cyanide rinse
acid/alkaline rinse
Cr(VI) rinse
lime HClO-
cyanide oxidation
neutralisation
Cr(VI) reduction
NaHSO3 H2SO4
NaOH flocculation decantation
solids
filter press
lime or NaOH
sand filter
treated effluent
Figure 1. Simplified flowsheet of S.E.T.S. effluent treatment plant 2.2 Composition of the effluents Samples of S.E.T.S. effluents were analyzed. Examples of effluent compositions are given in Table 1. The information given by effluent analyses was used to choose representative data. The further design of new flowsheets and comparison with the conventional process are based on these data (Table 2). 3.
MATERIAL AND METHODS
3.1 Pilot plant A glass column, 0.10 m inner diameter and 2.75 m total height was filled with pozzolana (8-10 mm) kindly supplied by Carrière de la Denise, le Puy, France. The temperature was kept at 32°C by water circulation in an outer double-jacket. The column was up-flow fed in continuous mode. Gas flow-rates were 20 l.h-1 for H2 and 1.2 to 1.6 l.h1 for CO2. The synthetic medium was made up of concentrated Industrial Urea Medium and concentrated Cr(VI) solution separately pumped into the column. The bioreactor was connected to a glass precipitation column, 0.04 m inner diameter and 1.75 m total height (Fig. 2). Metals-containing acid/alkaline effluent and alkaline effluent from the cyanide destruction tank fed the precipitation column.
550
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Table 1. Composition of S.E.T.S. effluents (n.d: not determined) Compounds [mg.l-1]
Cr(VI) rinse effluent
Acid/alkaline rinse effluent
Cr(VI) reduction tank
Cyanides rinse effluent
Ca2+
50
49
56
6
9
8
9
6
32
153
52
251
15
60
8
5
1
3
0.4
8
Mg
2+
Na+ +
K
NH4
+
-
54
570
55
86
-
39
74
20
7
3-
0
20
1
0
Cl
NO3 PO4 -
F
30
3
4
0.4
2-
6
26
480
21
-
n.d.
n.d.
n.d.
42
3+
2
13.6
0.8
0.25
2+
0
0.06
0.08
1.2
Zn2+
SO4
CN Fe
Cu
4
34
21
5.5
Ni
2+
0.04
1.75
0.08
12.25
Al
3+
3
2.5
1.4
0
10
0
0.02
n.d.
Cr(VI)
Table 2. Data chosen for process design Effluents Cr(VI) rinse effluents Cr(VI) rinse effluents Cr(VI) reduction tank Acid-alkaline effluents Acid-alkaline effluents Acid-alkaline effluents Acid-alkaline effluents
Flow-rate [m3.h-1] 4 4 4 10 10 10 10
Compound Cr(VI) SO42SO42Zn2+ Fe3+ Ni2+ Cu2+
Concentration [mg.l-1] 25 85 925 60 50 2.5 0.1
3.2 Experimental programme The experimental programme was the following: (phase 1) the column was inoculated with a D. norvegicum-containing bacterial population (2 litres at 2x108 bact.ml-1), and the biofilm was allowed to develop by feeding the reactor with a 5 g.l-1 SO4-containing Industrial Urea Medium (IUM); (phase 2) the sulfate concentration in the feeding was decreased to 0.5 g.l-1; (phase 3) Cr(VI) concentration in the feed was increased from 0 to 30 mg.l-1; (phase 4) the residence time was reduced to 7 h by increasing the feed flow-rate; (phase 5) the column was fed with a real Cr(VI) containing electroplating effluent. Nutrients were added to the effluent, and a concentrated Cr(VI) solution was separately pumped in order to adjust the Cr(VI) concentration to 30 mg.l-1. The composition of the Industrial Medium with Urea was the same as in [4]. The liquid effluents from the bioreactor were conducted into the precipitation column, also fed with the outlet-gas from the biological step. A metals-containing acid/alkaline real effluent was injected into this 551
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system. An alkaline effluent is used to increase the pH and favour sulfide precipitation (Fig. 2). Sulfide sludge settled down at the bottom of the column, and the gas phase was bubbled into a zinc sulfate solution (50 g.l-1) in order to remove H2S through ZnS precipitation. H2
CO2
gas
liquid
gas
effluent Cr(VI)
nutrients and sulfate
acid/alk. alkaline effluent
sludge
effluent
Figure 2. Schematic representation of the pilot plant 3.3 Analysis The bioreactor was equipped with pH and Eh (redox potential of Pt/Ag-AgCl) probes. Samples were daily taken in the outlet stream in order to analyse dissolved sulphide (potentiometric method), sulfate (kit MERCK spectroquant® 1.1458.0001) and Cr(VI) (kit MERCK spectroquant® 1.14758.0001). 4.
RESULTS
4.1 Design of potential process flowsheets including a biological step Three process flowsheets including a biological step are proposed. In these alternative flowsheets, metals (except chromium) are precipitated as sulphides. In the first option, the chemical neutralisation tank is replaced by a bioreactor (Fig. 3). The effluent from the Cr(VI)-reduction tank contains sulfate that will be converted into HS- by sulfate-reducing bacteria. Metals from other effluents will be precipitated into the bioreactor with biologically produced HS-. The second option proposes to perform the biological production of HS- and the precipitation of metal sulphides in separate reactors (Fig. 4). In this process, the chemical Cr(VI) reduction is maintained, and this chemical step still provides sulfate for the biological step. Some alkaline effluent from the cyanidesoxidation tank may be used to increase the pH in the bioreactor. In the third option (Fig. 5), the chemical Cr(VI)-reduction tank is replaced by a bioreactor in which SO42- and 552
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Cr(VI) are biologically reduced. Some sulfate has to be added into the bioreactor, because chromate rinse effluents cannot provide enough SO42-. This last option was tested at the laboratory-pilot scale. cyanide rinse
acid/alkaline rinse
Cr(VI) rinse
lime cyanide oxidation
HClO-
Bio-reactor HS- production and precipitation of metals
NaHSO3
Cr(VI) reduction
H2SO4
NaOH Nutrients, H2
solids
flocculation decantation
filter press
lime or NaOH
sand filter
treated effluent
Figure 3. Alternative flowsheet, option 1 – neutralisation tank replaced by a bioreactor
Nutrients, H2
cyanide rinse
Bio-reactor
lime or NaOH
HS- production
Cr(VI) rinse
Precipitation of metals as sulfides
Cr(VI) reduction
lime HClONaOH
solids
cyanide oxidation acid/alkaline rinse
filter press
flocculation decantation
NaHSO3 H2SO4
lime or NaOH
sand filter
treated effluent
Figure 4. Alternative flowsheet, option 2 – neutralisation tank replaced by a bioreactor and a precipitation reactor
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cyanide rinse
acid/alkaline rinse
Cr(VI) rinse
lime -
HClO
cyanide oxidation
Precipitation of metals as sulfides
Bioreactor Cr(VI) reduction HS- production
Nutrients, H2 Na2SO4
NaOH flocculation decantation
solids
filter press
lime or NaOH
sand filter
treated effluent
Figure 5. Alternative flowsheet, option 3 – Cr(VI) chemical treatment and neutralisation tanks replaced by a bioreactor and a precipitation reactor 4.2 Biological treatment of a Cr(VI)-containing effluent With 500 mg.l-1 SO42- in the feeding medium, Cr(VI) concentration was progressively increased to 30 mg.l-1 (Fig. 6, phase III), and the residence time was reduced to 7 h while maintaining the efficiency of the bioreactor: in these conditions, Cr(VI) concentration in the outlet was lower than 0.1 mg.l-1 (Fig. 6, phase IV), and sulfate was nearly entirely reduced. The accumulation of blue-green colored Cr(III) hydroxyde on the pozzolana was visible. When the synthetic feeding medium was replaced by the real Cr(VI) rinse effluent from S.E.T.S. (Fig. 6, phase V), the system was not disturbed as long as the residence time was 7 h or higher. This result was positive, and proved that the bacteria were able to resist to pollutants such as nitrate or cyanide, that were present in the effluent at low concentrations. When the residence time was decreased from 7 to 5.5 h, sulfate and Cr(VI) reduction processes were inhibited. The Cr(VI) concentration in the outlet reached 11 mg.l-1. A batch phase without chromate was necessary to restore the bacterial activity.
Figure 6. Evolution of Cr(VI) concentration in the feed and outlet of the bioreactor, and residence time vs. operating time 554
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In the outlet stream, pH was in the range of 8 to 9, and redox potential fluctuated between –400 and –500 mV (ref. Ag/AgCl) (Fig. 7). The inhibition of the biofilm, when the flow-rate was too much high, was accompanied by a strong increase in redox potential, from –400 to –100 mV. This sensitive parameter will be useful as an indicator for the operating status of the bioreactor.
Figure 7. Evolution of pH and redox potential of the outlet liquid from the bioreactor vs. operating time
Elimination of metals from the acid/alkaline rinse was performed in the precipitation column. The liquid effluent from the bioreactor brought dissolved sulfide in the mixture. However, the bubbling bioreactor outlet gas, which contains H2S, greatly improved the precipitation efficiency. Only Zn2+ and Cu2+ were present in relatively high concentrations in the real effluent. These two metals were entirely precipitated by applying a 20-min residence time in the precipitation column. However, a residence time of 60 min was applied, because the effluent may contain metals with slower precipitation kinetics than Cu2+ and Zn2+, and it was thought to be more realistic for the cost estimate. 4.3 Preliminary technical and economical evaluation A preliminary evaluation of the process was performed based on the following configuration: Cr(VI) is reduced to Cr(III) hydroxide in a column fixed-film bioreactor filled with pozzolana, working at 30°C and pH in the range 7-8.5. The chromium hydroxides are recovered by settling at the bottom of the column. Gas and liquid from the bioreactor are mixed in a reactor with acid/alkaline effluents to precipitate metal sulfides, which are recovered by settling and filtration. The gas phase from the precipitation reactor is re-injected into the bioreactor. Consumption and production fluxes for this process are given in Table 3. For the evaluation of the classical treatment (bisulfite for Cr reduction and lime for metals precipitation), data from Rigaud and Girards [1] were used. The investment and operating costs of this process will largely depend on the way to obtain H2 (electron source for the anaerobic bacteria) and CO2 (for pH regulation). In-situ coal reforming would correspond to the highest investment and lowest operating cost. Buying the two gases separately as stock tanks would correspond to the lowest investment and highest operating cost. In this last configuration, the investment cost was estimated 555
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closely to 281,000 €, and the operating cost reaches 114,000 €.year-1. As a comparison, the operating cost of the classical physico-chemical process is close to 75,000 €.year-1. The cost of waste solids elimination would be lower with the "biological" process (9,600 €.year-1), than with the classical process (17,300 €.year-1). Table 3. Process material balance Products urea (46% N) H2SO4 KOH MgSO4 Na acetate DAP pozzolana H2 CO2 Products liquids wet sulfides dry sulfides Wet Cr hydroxide Dry Cr hydroxide
5.
Consumption (tonnes.year-1) 1.5 10 12.6 3.4 17.5 1.6 37 1.6 28.3 Production 3. -1 m year tonnes.year-1 142,252 5.4 15.8 3.5 13.9 0.6 3.1 0.36 1.9
DISCUSSION AND CONCLUSION The study of a real effluent treatment plant resulted in proposing several process flowsheets including a biological step for the reduction and/or precipitation of metals using sulfate-reducing bacteria. These micro-organisms are able to reduce Cr(VI) into Cr(III) by two combined mechanisms: direct enzymatic reduction [2, 3, 4], and indirect chemical reduction by hydrogen sulfide [5]. The bacterial population was able to maintain its ability to reduce sulfate and chromate in the pilot plant fed with a real surface-treatment chromic effluent. The minimum residence time that could be applied was 7 h for a feed Cr(VI) concentration of 30 mg.l-1. In 2-litre column bioreactors, the sulfate-reducing biofilm was able to withstand a feeding Cr(VI) concentration of 100 mg.l-1 when the residence time was 19 h [5]. The hydrogen sulfide produced in the bioreactor was used to treat the other metals-containing effluents of the surface-treatment plant. A preliminary economic evaluation revealed that a new process including a biological step would be more expensive than the classical physical and chemical process, particularly in terms of operating costs. However, further optimisation of the biological step may lead to a cheaper configuration. The nutrients consumption can probably be reduced. Devices to produce H2 and CO2 at lower cost may be developed. Lowering the pH of the feed solutions could reduce carbon dioxide consumption. The biological step allows to reduce the amount of solid wastes. The cost of waste management, particularly for this type of solids containing toxic metals, will probably increase, and surface treatment plants will be encouraged to reduce their sludge production. In this context, the possibility to precipitate metals as sulfides rather than as hydroxides could become interesting. 556
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AKNOWLEDGEMENTS This research was supported by ADEME (Contract No. 0002016), and by the Research Division of BRGM (Contribution N°01214). REFERENCES
1. J. Rigaud and L.M. Girard (eds.) Traitements de Surfaces – Epuration des eaux. Publication de l’Agence de l’Eau Rhône-Méditerranée-Corse et du Syndicat Général des Industries de Matériels et procédés pour les Traitements de Surface, (2002). 2. D.R. Lovley and E.J.P. Phillips, Appl. Environ. Microbiol., 60 (1994) 726. 3. C. Michel, M. Brugna, C. Aubert, A. Bernadac and M.Bruschi, Appl. Microbiol. Biotechnol., 55 (2001) 95. 4. F. Battaglia-Brunet, S. Foucher, A. Denamur, I. Ignatiadis, C. Michel and D. Morin, J. Ind. Microbiol. Biotechnol., 28 (2002) 154. 5. F. Battaglia-Brunet, S. Foucher, D. Morin, I. Ignatiadis Water, Air & Soil Pollution: Focus, in press
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Effects of total-solids concentration on metal bioleaching from sewage sludge L.D. Villar* and O. Garcia Jr†. Department of Biochemistry and Chemical Technology, Institute of Chemistry, São Paulo State University, P.O. Box 355, Araraquara, SP 14.801-970, Brazil Abstract The possibility to remove metals from sewage sludge at high solids concentration by bacterial leaching may reduce the inherent cost of the process by lowering the sludge volume to be treated and increasing the metal concentration in the leachate. The effect of total-solids on the removal efficiency of chromium, copper, lead, nickel, and zinc from anaerobically digested sewage sludge was investigated. Indigenous sulphur-oxidizing acidithiobacilli were enriched from the sludge with 1% (w/v) S°. After 3 consecutive transfers, the enriched sludge was retained as the inoculum. The bioleaching assays were conducted in shake flasks containing fresh sludge at pH 7.0, which was inoculated with the enriched sludge (5% v/v), and supplemented with 0.5% (w/v) S°. Flasks were incubated at 30°C and 200 rpm. Total-solids concentration investigated varied from 10 to 40 g L-1 (10, 25, 32.5, and 40 g.L-1). The variation of pH and oxidation-reduction potential (ORP) was monitored during the bioleaching time course. Decrease in pH and increase in ORP values were faster at lower solids concentration. In spite of this, final solubilization efficiency was independent of solids variation for each metal investigated, resulting in the average final solubilization: Cr, 50.6%; Cu, 87%; Ni, 94.4%; Pb, 41.2%; and Zn, 99.5%. On the opposite, maximum solubilization rates, expressed as mg L-1.day-1, were greatly influenced by the solids concentration for each metal ion tested, resulting in either increase or decrease of solubilization rate depending on the metal considered. Thus, the effect of solids concentration on metal bioleaching was more pronounced on the kinetics of bioleaching than on the final removal achieved, thus encouraging the application of the process even at high solids concentration. Keywords: bioleaching, indigenous thiobacilli, metals, sewage sludge, total-solids
* Present address: Centro Técnico Aeroespacial, Instituto de Aeronáutica e Espaço, Divisão de Química 12228-904, São José dos Campos-SP, Brazil. (
[email protected]) † Corresponding author (
[email protected]) Thanks are due to FAPESP for supporting this research, and to the municipal sewage sludge works (SABESP) of Franca for providing the sludge samples. OGJ acknowledges CNPq for Researcher Fellowship.
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1.
INTRODUCTION Disposal of sewage sludge generated during the treatment of municipal wastewater is becoming a growing problem. As an illustration of this issue, the metropolitan region of the city of São Paulo, Brazil, generates 295 ton.day-1 (dry basis) of sewage sludge, with an estimation of 750 ton.day-1 to be generated in year 2015 [1]. To accommodate this increase, sludge disposal in agriculture is being encouraged, since it appears to be the most attractive option for both economical and environmental reasons. Sewage sludge can be considered as a low-grade N-P fertilizer with a typical composition of 4% (w/w) nitrogen and 2% (w/w) phosphorous. Additionally, because of its organic composition, usually near to 40% (w/w), sludge can also be utilized as a soil ameliorant [2]. However, the accumulation of metals present in wastewaters into the solid phase of the sludge may pose some constraints to the sludge application into agricultural fields. Even when metal concentration in sludge is within the limits established through national regulations, there is still a risk they could reach toxic levels in soils [3]. Some recent studies [4, 5] have demonstrated that there is an accumulation of metals in soils after sludge application for long periods, although this accumulation was not showed to increase metal bioavailability in soil. In order to offer some alternative to the problems associated with the presence of metals in sewage sludge, chemical and biological processes have been investigated for the leaching of metals from sludge. It has been demonstrated that the chemical leaching is often more expensive than the bioleaching process due to the consumption of large amounts of inorganic acids [6]. Bioleaching of metals from sewage sludge employs sulphur- and iron-oxidizing bacteria from the Thiobacillus genus, which had some species recently reclassified into Acidithiobacillus, Halothiobacillus and Thermiothiobacillus [7]. Thiobacillus and the other new genera promote the acidification of sewage sludge through the oxidation of reduced sulphur compounds to sulphate by either direct or indirect mechanisms [8]. Most of the studies on bioleaching of metals from sludges have been performed in Canada. Some of these works have employed cultures of A. ferrooxidans and/or A. thiooxidans at the beginning of the studies on metal bioleaching from sludge. Afterwards, indigenous thiobacilli, enriched from sludge amended with sulphur, were successfully used, thus eliminating the requirement of chemical acidification of sludge to pH 4.0. Although several parameters can affect the bioleaching process, some of them are of greater importance when establishing the conditions for field application. Those parameters are initial pH, temperature and total-solids concentration. In previous works, we have addressed the effects of temperature on the kinetics and efficiency of metal bioleaching from sludge at two different initial pH values: pH 4.0 [9] and pH 7.0 (unpublished data). Higher efficiencies and solubilization rates were obtained at initial pH of 7.0 and temperature of 30°C for chromium, copper, and lead. Nickel and zinc showed higher solubilization rates for initial pH of 4.0 at the same temperature (unpublished data). In the operation of sewage treatment plants, it is less expensive to manage higher solids concentration sludges, thus implementation of bioleaching process would be more feasible if higher solids in sludge would not interfere in the efficiency of this process. In this context, we have investigated the effect of solids concentration, in the range of 10 to 40 g.L-1, on the kinetics and efficiency of metal solubilization from anaerobically treated sludge. 560
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2.
MATERIALS AND METHODS
2.1 Sampling and characterization of sludge A sample of anaerobically digested secondary sludge was obtained from the municipal wastewater treatment plant in the city of Franca, state of São Paulo, Brazil. Sludge was collected in sterilized polyethylene bottles, shipped cold, and kept at 4°C before utilization. Total solids concentration (TS) was determined as described in Standards Methods [10] in a number of 5 replicates. Total concentration for the metals chromium, copper, lead, nickel, and zinc in the sludge were determined after acid digestion with HNO3/H2O2 [11]. Dissolved metals in the sludge were determined by centrifugation of 20-mL samples at 3,300 g for 30 min, followed by membrane filtration. Determination of dissolved- and total-metal concentrations was carried out in triplicates, and analysis was conducted using plasma emission spectroscopy (ICP-AES). 2.2 Metal bioleaching assays Indigenous sulphur-oxidizing bacteria were enriched from the sludge with 1% (w/v) S°. After 3 consecutive transfers, the enriched sludge was retained as the inoculum. To obtain different total solids concentrations, the sludge (25 g total solids L-1) was either diluted with deionized water or concentrated by centrifugation. The solids concentrations obtained were 10, 32.5, and 40 g L-1. For the experiments, 250 mL of sludge (initial pH of 7.0) was inoculated with 5% (v/v) of the enriched sludge, and supplemented with 0.5% (w/v) S°. For each solids concentration investigated, duplicate flasks were used. Two controls were included for TS of 25 g L-1: (i) biological control: inoculated sludge without S° amendment, and (ii) chemical control: autoclaved sludge (at 121°C for 20 min) amended with S°, but not inoculated. All flasks were incubated in a gyratory shaker at 30°C and 200 rpm. 2.3 Chemical analysis Samples, 20 mL, were periodically withdrawn for pH (combined electrode, Orion 520A) and ORP (platinum electrode, Cole-Palmer) measurement and subsequent centrifugation at 3,300 g for 30 min, with the supernatant being analysed for solubilized metals by plasma emission spectroscopy (ICP-AES). Metal solubilization efficiency, defined as r, was calculated as the ratio between the solubilized metal due to bioleaching, and total metal present in the sludge (Equation 1). t [ Me sol ] − [ Me diss ] r(%) = × 100 (1) [ MeT ] × TS × 0.001
where, t [ Me sol ] , solubilized metal present in sludge at time t of the bioleaching experiment, in mg -1 L ; [ Mediss ] , dissolved metal in sludge before the bioleaching experiment, in mg L-1;
[ MeT ] , total-metal concentration in sludge, in mg kg-1 dry sludge; TS , total-solids concentration for sludge, in g dry sludge L-1.
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3.
RESULTS AND DISCUSSION The enrichment of indigenous sulphur-oxidizing thiobacilli for inoculum production resulted in a reduction of the time required to decrease the sludge pH down to 1.3 from 8 to 6 days after three consecutive transfers (data not shown). Both controls performed for the initial acidification did not show any significant changes in pH (data not shown). The enriched sludge was kept at 4°C and used as inoculum. Total-solids concentration for the sludge samples collected was determined in five replicates and resulted in the concentration of 25 g L-1. The average total-metal concentration for this sludge is presented in Table 1 along with the U.S. EPA [12] recommended limits for land application. Table 1. Total-metal concentration (mg.kg-1 dry sludge) and dissolved-metal concentration (mg.L-1) for the anaerobically treated sludge U.S. EPA [12] recommended limits for total-metal concentration for land application. Values in parentheses are dissolved/total-metal concentration ratio (% w/w)
Total-Metal Concentration Anaerobic Sludge
Recommended Limits
Dissolved-Metal Concentration
Chromium
225
-*
0.08 (1)
Copper
253
1,500
0.04 (1)
Lead
129
300
0.10 (3)
Nickel
53
420
0.13 (9)
Zinc
929
2,800
0.39 (2)
Metal Ions
* No recommended limit.
Dissolved-metal concentrations were significant only for nickel (9% of total metal concentration present as dissolved metal), indicating that metal ions in sludge are mainly in the solid phase. Despite of the low values obtained for dissolved-metal concentration, they were taken into account when calculating the solubilization efficiency by using Equation 1. Although the metal contents for this sludge are below the EPA recommended limits, for all the metal ions analyzed, it is still interesting to look for processes for metal removal from sludge, since decontaminated sludge can be applied to soil at higher loads and for a longer period of time, with a lower risk of environmental contamination. Figure 1 presents the pH and ORP profiles obtained during the bioleaching assays. Both the decrease in pH and the increase in ORP were faster at lower solids concentrations due to changes in slope, as it can be verified from Figure 1. For solids-concentration values of 32.5 and 40 g L-1, a short lag phase of about 1 day was also observed for pH decrease (Figure 1). Some studies [13, 14], which had also determined the sulphate production during bioleaching, had attributed this observation to the increase in buffering capacity of the sludge with higher solids contents, thus preventing the pH to fast decrease. The controls showed in Figure 1 were run for the TS of 25 g L-1. Both controls did not show a significant variation in pH (decrease from 7.0 to 6.0), although some increase in ORP was observed, especially for the biological control. During bioleaching, the anaerobic sludge in the controls was submitted to aeration, which could explain the increase in ORP observed. 562
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Figure 1. Variation in sludge pH (A) and oxidation-reduction potential (B) during bioleaching assays at various sludge solids concentration Symbols: (■) TS = 10g L-1; ( ) TS = 25 g L-1; (▲) TS = 32.5 g L-1; (○) TS = 40 g L-1; (+) biological control for TS = 25 g L-1; (x) chemical control for TS = 25g L-1
The solubilization efficiency (r) is presented in Figure 2 for the metals investigated.
Figure 2. Metal solubilization during bioleaching assays for the metals chromium, copper, lead, nickel, and zinc at various total-solids concentration Symbols: (■) TS = 10g L-1; (△) TS = 25 g L-1; (▲) TS = 32.5 g L-1; (○) TS = 40 g L-1; (+) biological control for TS = 25 g L-1; (x) chemical control for TS = 25g L-1
Fitting of the data obtained was conducted following the sigmoidal model described in Boltzmann equation (Equation 2). Both controls did not show any significant solubilization. A1 − A2 + A2 (2) y= 1 + e ( x − x0 ) / dx 563
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where, xο is the centre of the linear range of the function, dx is the width of the linear range, A1 is the initial y value, i.e., y (-∞) and A2 is the final y value, i.e., y (+∞). The values of A2 obtained from Boltzmann’s fitting were compared by analysis of variance (ANOVA) for each metal. When the null hypothesis was rejected, Tukey’s test was applied to compare the means. The results obtained are presented in Table 2, along with the maximum solubilization rates (vmax) calculated from the slope of the linear part of the curves showed in Figure 2. The same statistical treatment mentioned above was applied to the maximum solubilization rates. Final solubilization values higher than 100% observed for nickel and zinc (Table 2) can be attributed to the metal income from the enriched sludge used as inoculum, which was not considered by Equation 1. Table 2. Final solubilization (in %) and maximum solubilization rate (in mg.L-1.day-1) for the bioleaching assays at various total-solids concentration (in mg.L-1) Metal Chromium
Copper
Nickel
Lead
Zinc
Total-Solids Concentration
Final Solubilization
Maximum Solubilization Rate (vmax)
10
53.3a*
1.8a
25
52.8a
1.2b
32.5
45.8a
1.2b
40
50.5a
1.2b
10
80.9a
1.5a
25
78.8a
1.7a
32.5
96.4a
4.1b
40
91.9a
4.1b
10
92.2a
0.23a
25
81.6a
0.58b
32.5
103a
0.82c
40
101a
0.89c
10
49.3a
0.34a
25
38.4a
0.74b
32.5
39.0a
0.72b
40
38.2a
0.53c
10
102a
3.6a
25
99.1a
10b
32.5
98.5a
24c
40
98.3a
18d
* For each metal, final solubilization and maximum solubilization rates were submitted to analysis of variance (ANOVA). No statistical differences were obtained for final solubilization at the solids concentrations investigated. For maximum solubilization rates, means followed by the same letter are not different by the Tukey’s test at 5%.
Although final solubilization for the metals investigated was not influenced by the solids variation, as showed by the results obtained by analysis of variance, at 5% of 564
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significance level (Table 2), solids markedly modified maximum solubilization rates. For chromium, at the lower solids concentration it was obtained the higher solubilization rate, whereas for the other metals, there was a tendency of higher rates to be obtained at higher solids contents, especially at 25 and 32.5 g.L-1. In fact, higher solids concentration implies higher insoluble metal concentration in the medium, if expressed in mg metal.L-1. The increase in metal concentration accelerates the process of metal bioleaching through the chemical equilibrium. However, higher solids concentration also implies lower decrease in pH, and thus, lower hydrogen ions available for bioleaching, which decreases the rate of solubilization. For this reason, depending on the metal being leached, such equilibrium may be affected more by hydrogen ions availability than by insoluble metal concentration, or vice-versa. 4.
CONCLUSIONS Variations of the solids concentration in the range of 10 to 40 g.L-1 do not interfere in the final metal solubilization yield obtained during metal bioleaching from sewage sludge. However, the maximum solubilization rate is markedly influenced by solids variation, with higher rates being obtained at TS above 25 g.L-1 for copper, lead, nickel and zinc. An immediate advantage resulted from this study is the possibility to conduct metal bioleaching process in wastewater treatment plants at higher solids concentration, thus lowering the costs of this sludge treatment. REFERENCES
1. M.T. Tsutiya (ed.), Biossólidos na Agricultura (Biosolids in Agriculture), SABESP, São Paulo, 1997. (in Portuguese) 2. L. Korentajer, Water SA, 17 (1991) 189. 3. R. Renner, Environ. Sci. Technol., 34 (2000) 430A. 4. J.J. Kelly, M. Häggblom and R.L. Tate III, Soil Biol. Biochem., 31 (1999) 1467. 5. B.P. Knight, A.M. Chaudri, S.P. McGrath and K.E. Giller, Environ. Pollut., 99 (1998) 293. 6. D. Couillard and G. Mercier, Environ. Pollut., 66 (1990) 237. 7. D.P. Kelly and A.P. Wood, Int. J. System. Evolut. Microbiol., 50 (2000) 511. 8. A.M. Martin (ed.), Biological Degradation of Wastes, Elsevier, Amsterdam, 1991. 9. L.D. Villar and O. Garcia Jr., Int. Biohydrometallurgy Symp., Ouro Preto, Brazil (2001) 21B. 10. L.S. Clesceri, A.E. Greenberg and R.R. Trussell (eds.), Standard Methods for the Examination of Water and Wastewater, 17th ed., Am. Public Health Assoc./Am. Water Works Assoc./Water Pollut. Control Fed., Washington, DC, 1989. 11. J.J. Delfino and R. E. Enderson, Water & Sewage Works (1978) R32. 12. U.S. EPA., Standards for the Use and Disposal of Sewage Sludge (Code of Federal Regulations 40 CFR Part 503), Washington, DC, 1996. 13. T.R. Sreekrishnan, R.D. Tyagi, J.F. Blais and P.G.C.Campbell, Water Res., 27 (1993) 1641. 14. R.D. Tyagi, J.F.Blais, J.C. Auclair and N. Meunier, Water Environ. Res., 65 (1993) 196.
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Enhancement of electrodialytic soil remediation through biosorption Pernille E. Jensena*, Lisbeth M. Ottosena and Birgitte K. Ahringb a
BYG.DTU, Building 204, Technical University of Denmark, DK-2800 Kgs. Lyngby, Denmark b BiC-BioCentrum, Building 227, Technical University of Denmark, DK-2800 Kgs. Lyngby, Denmark 1.
INTRODUCTION Toxic metals are being introduced to the environment through numerous different processes including different industrial production-processes, incineration emissions and waste disposal. The final sink for most metals in the environment is soil or sediment, where the metals adsorb strongly. Therefore metal-polluted soils are found in vast amounts in populated areas, where they possess a hazard to both environment and humans. Through the last decades a rising concern of metal pollution has developed, and different methods for decontamination of metal-polluted soils have been investigated. Since the metals cannot be degraded like organic pollutants, the only three ways to treat such metal-polluted land are i) to remove the polluted soil and deposit it in a place, where it does less harm or ii) stabilize the metals in the soil to make them less bioavailable and thereby less toxic or finally to iii) mobilize the metals in order to remove them from the soil followed by either deposition of the metals or possibly by reuse of these. At the moment deposition of polluted soil is the dominating procedure, when dealing with contamination problems. However the ideal solution seen in an environmental perspective would be the recovery of both metal and soil through separation of the two. This research aims at development of a method for such a separation. The separation process used is the electrodialytic remediation method described below, which has shown efficient in remediation of copper polluted soil [Ottosen et al., 1997]. The method has also been investigated for remediation of soils polluted with other metals, e.g. lead. It was shown that mobilization of lead only takes place at very low soil pH (< 3). At such low pH many other soil components dissolve, leaving the soil less usable after treatment [Ottosen et al., 2001]. In order to enhance the electrodialytic mobility of lead at higher pH values, addition of microbial biosorbents to the soil is suggested. The properties and possibilities of such biosorbents are further discussed below.
* Corresponding author:
[email protected]
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2.
ELECTRODIALYTIC SOIL REMEDIATION The electrodialytic soil remediation method was developed at the Technical University of Denmark for decontamination of soil polluted with heavy metals. The method takes advantage of electric current, which has the ability to desorb and move ions in soil [Hansen et al., 1997]. The electric current results in hydrolysis reactions at the electrodes, producing acid at the anode and base at the cathode. This results in development of an acidic front, which moves from the anode to the cathode and an alkaline front moving in the opposite direction. The development of an alkaline front is problematic to the remediation because lead and most other metals precipitate in alkaline environments. Therefore, in addition to electric current, the electrodialytic method involves a cation-exchange membrane, separating the catholyte from the soil to prevent hydroxide ions from entering the soil [Ottosen et al., 1997]. Similarly at the anode end an anion-exchange membrane is separating anolyte and soil. This makes it possible to control the chemistry of the system liquids. The anion exchange membrane does however not prevent development of an acidic front because water splitting is taking place at the membrane surface, followed by transport of the hydroxide-ion across the anion-exchange membrane [Ottosen et al. 2000]. In figure 1 a schematic illustration of a laboratory cell for investigation of the electrodialytic soil remediation method is given.
Figure 1. Schematic illustration of the electrodialytic remediation method. The soil is placed in compartment ΙΙ. Electrolytes are placed in compartment Ι and ΙΙΙ. Illustrated is the rejection of cations by the anion-exchange membrane, the rejection of anions by the cation-exchange membrane and the water splitting at the surface of the anion-exchange membrane. AN = Anion-exchange membrane. CAT =Cationexchangemembrane 3.
BIOSORPTION Biosorption has been extensively investigated for decontamination of heavy-metal polluted solutions [Volesky and Holan, 1995]. The process has many advantages which can also be of value for soil remediation: Cheap biomass can be obtained from different industrial waste streams [Volesky, 2001], a fast biosorption process shortens the remediation time [Gupta et al., 2000], possible regeneration and reuse of both biomass and heavy metals makes the method both economically and practically feasible [Volesky, 2001]. Also biosorption works well at low heavy metal concentrations (1-100 mg/L) [Gupta et al., 2000], which are the concentrations most often found in soil pore-water during electrodialytic remediation [Ribiero, 1998].
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Biosorption of Pb from solution has been investigated for several different organisms including bacteria, algae, yeasts and filamentous fungi. Examples are the bacteria Phormidium laminosum [Blanco et el., 1999], Pseudomonas aeruginosa [Chang et al.,1998], several different marine algae [Holan and Volesky, 1994] and mixed species of green, reed and brown algae [Diniz et al., 2001], the filamentous fungi Aspergillus niger, Mucor rouxii and Rhizopus oryzae [Baik et al. 2002], Rhizopus oligosporus [Ariff et al., 1999] and Rhizopus arrhizus [Fourest et al., 1994; Brady and Tobin, 1995], and the yeasts Saccharomyces uvarum [Ashkenazy et al., 1997] and Saccharomyces cerevisiae [Engl and Kunz, 1995; Ashkenazy et al., 1999]. Also experiments have been performed with mixed aerobic wastewater culture [Ghosh and Bupp, 1992] as well as microbial products such as chitin and chitosan [Eiden et al., 1980] and cell walls [Baik et al., 2002]. Results are reported for both living [Ghosh and Bupp, 1992; Engl and Kunz, 1995] and dead biomass [Holan and Volesky, 1994; Diniz et al., 2001]. Also biomass that has been pretreated in different ways can be mentioned such as NaOH treated cell walls [Ashkenazy et al., 1997; Baik et al., 2002] or dried and powderized cells [Fourest et al., 1994; Ariff et al., 1999] and acetone washed cells [Ashkenazy, 1997]. Finally biomass which has been immobilized in variable matrixes has been tested for possible use in treatment of lead polluted wastewater. Examples are Blanco et al. (1999) who tested biomass immobilized in polysulfone and epoxy resin beads and Chang et al. (1998), who tested biomass entrapped in calcium alginate beads and polyacrylamide gel. It is shown that biomass can be as efficient an adsorbant as commercial ion exchangers. Holan and Volesky (1994) showed this for lead at solution equilibrium concentrations of 200 mg/g. However, when lowering the equilibrium concentration to 10 mg/g, the commercial ion exchangers showed a higher efficiency than the biomass. Maximum biomass-equilibrium capacities in the order of 300 mg/g lead are commonly seen [Aikin et al., 1979; Holan and Volesky, 1994; Diniz et al., 2001]. A maximum possible metal loading is often seen for solution-equilibrium concentrations above 200 mg/g lead [Ariff et al., 1999; Diniz et al., 2001]. For many biosorbents it is seen that lead is the metal sorbing most efficiently to biomass after uranium when looking at the mass adsorbed per gram dry biomass [Brady and Tobin, 1995]. Important parameters for biosorption are pH and growth state of the microorganisms [Ledin 2000]. At low pHs the sorption decreases because of protonation of the negatively charged groups on the biomass. At high pH sorption may also decrease. This is probably due to heavy metal precipitation or formation of negatively charged metal-complexes in alkaline environment. 4.
MICROBIAL ENHANCEMENT OF LEAD MOBILITY IN SOIL It is known, that microorganisms accumulate heavy metals in polluted soil systems [Ledin, 2000]. It has also been shown that heavy metals can be mobilized in aquifer materials by leaching with bacterial extracellular polymers [Chen et al., 1995], and in soil by stimulation of microbial growth [Chanmughathas and Bollag, 1987; Bender et al., 1989]. Further more it is shown that microorganisms can be transported in soil by addition of electric current [DeFlaun and Condee, 1997]. The purpose of this study is to add leadadsorbing microorganisms to lead polluted soil, making them work as “complexing agents” with mobilization of the metal as a consequence. The mobile and electrically charged microorganisms are then to be transported out of the soil by means of electrodialysis, carrying the lead with them.
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5.
EXPERIMENTAL WORK An experimental setup for addition of microbial matter to soil with succeeding electrodialytic remediation is developed. The setup needs to be different than the one shown in figure 1, because the microorganisms are unable to pass an ion-exchange membrane. Generally the size of soil-microorganisms is around 1 µm. Consequently a barrier with pores of this size or larger must be used to separate soil and process-water.
Figure 2. Schematic illustration of the setup for experimental evaluation of the bioelectrodialytic soil remediation method. AN = Anion-exchange membrane. CAT = Cation-exchange membrane. NET = fiberglass armed insect-net with pore size of 1 mm. PM = Passive Membrane. Soil and biomass is mixed in compartment ΙΙΙ
A barrier consisting of fiberglass-armed insect-net with a pore-size of 1 mm is tested, since this represents a low-cost solution. With this setup it is possible to collect the heavymetal loaded biomass in the concentration-chamber ΙΙ in front of the anion-exchange membrane. Circulation of the process liquid at the soil side of the cation-exchange membrane is necessary in order to avoid fouling of the membrane and high electrical resistance. A passive membrane is placed between the insect-net and the cation-exchange membrane in order to avoid soil particles in the pump and tubes for this circulation. The setup is seen in figure 2. 5.1 Soil characteristics Laboratory experiments with two different lead-polluted soils are performed. Soils with medium carbonate content are chosen, because it was earlier shown that leadpolluted, carbonaceous soils are not easily decontaminated by electrodialysis [Pedersen, 1999]. Such soils are therefore candidates for combined bio-electrokinetic treatment. Also both soils contain lead at levels exceeding the value where Danish authorities prohibit human contact with the soil (400 mg/kg). Both soils are Danish. One soil is from a former army-site in Copenhagen: Holmen (soil 1). The other soil (soil 2) comes from the town Kalundborg situated in western Zeeland. The source of pollution at this site is not known. Parameters important for the remediation are shown in table 1. As is common for originally polluted soils, the lead is inhomogeneously distributed within the soil. This is illustrated by the high standard deviation on the analytical result. When experiments are made, the soil used is mixed as well as possible, and three samples are taken from this mixture before starting the experiments to give a more precise estimate of the exact amount of soil in the individual experiment. Apart from lead, soil 2 also contains copper at 764 +/- 424 mg/kg and Zn at 2028 +/- 544 mg/kg.
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Table 1. Characteristics of experimental soils Pb [mg/kg] CaCO3 % (w/w) CEC meq/100g pH Clay % (w/w) Org. matter % (w/w)
Soil 1 637+/-242 7,7 8,2 7,4 0,7 7,0
Soil 2 765+/-760 9,1 4,9 7,6 2,8 3,6
5.2 Sequential extraction Sequential extraction of Lead has been performed in order to get an impression of how mobile the lead is in the soil before treatment. Results are shown in figure 3. The method used is the one recommended by the "Standards, Measurements and Testing Programme of the European Union" [Mester, 1998]. Here it is recognized that it is not possible to find a completely selective method for extraction of metals from certain soil fractions, but it is possible to use this method to get a qualitative impression of how the metals are bound and their mobility. Step 1 gives an estimate of how much metal is found as exchangeable ions and carbonates. Step 2 extracts metals that are reduceable. Step 3 extracts oxidizable metals and step 4 reveals the residual and least mobile fraction. The results of sequential extraction are shown in figure 3. It is seen that lead is binding tightly to the soil, as no lead is found as exchangeable ions and carbonates. Furthermore it is shown that the lead in soil 1 is less mobile than the lead in soil 2.
Figure 3. Results from sequential extraction of lead from the two soils 6.
CONCLUSIONS Lead polluted soil is not easily remediated, and improved remediation technology is a prerequisite for successful remediation. It is rendered that combination of biosorption and electrodialytic remediation can improve remediation efficiency in an economically feasible manner.
REFERENCES
1. Ariff, A.B., Mel, M., Hasan, M.A., Karim, M.I.A. (1999), The kinetics and mechanism of lead (II) Biosorption by powderized Rhizopus oligosporous, Worls Journal of Microbiology and Biotechnology, 15, 291-298. 571
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2. Ashkenazy, R., Gottlieb, L., Yannai, S. (1997), Characterization of Acetone-Washed Yeast Biomass Functional Groups Involved in Lead Biosorption, Biotechnol. Bioeng., 55, 1-10. 3. Ashkenazy, R., Yannai, S., Rahman, R., Rabinovitz, E., Gottlieb, L., Fixation of spent Saccharomyces cerevisiae biomass for lead sorption, Appl. Microbiol. Biotechnol., 52, 608-611. 4. Baik, W.Y., Bae, J.H., Cho, K.M., Hartmeier, W. (2002), Biosorption of heavy metals using whole mold of mycelia and parts thereof, Bioresource Technology, 81, 167-170. 5. Bender, J.A., Archibold, E.R., Ibeanusi, V., Gould, J.P., (1989), Lead removal from contaminated water by a mixed microbial ecosystem, Wat. Sci. Tech., 21,1661-1664. 6. Blanco, A., Sanz, B., Llama, M.J., Serra, J.L. (1999), Biosorption of heavy metals to immobilized Phormidium laminosum biomass, Journal of Biotechnology, 69, 227-240. 7. Brady, J.M., Tobin, J.M. (1995), Binding of hard and soft metal ions to Rhizopus arrhizus biomass, Enzyme and Microbial Technology, 17, 791-796. 8. Chang, J.-S., Huang, J.-C., Chang, C.-C., Tarn, T.-J. (1998), Removal and recovery of lead fixed-bed biosorption with immobilized bacterial biomass, Wat. Sci. Tech., 38, 171-178. 9. Chanmugathas P., Bollag, J.-M. (1987), Microbial mobilization of cadmium in soil under aerobic and anaerobic conditions, J. Environ. Qual., 16, 161-167. 10. Chen, J.-H., Lion, L.W., Ghiorse, W.C., Shuler, M.L. (1995), Mobilization of adsorbed cadmium and lead in aquifer material by bacterial extracellular polymers, Wat. Res., 29(2), 421-430. 11. DeFlaun, M.F., Condee, C.W. (1997), Electrokinetic transport of bacteria, Journal of Hazardous Materials, 55, 263-277. 12. Diniz, V.G.S., Silva, V.L., Lima, E.S., Abreu, C.M.A. (2001), Lead Biosorption in “Arribadas” Algal Biomass, Biohydrometallurgy: Fundamentals, Technology and Sustainable Development, Part B 13. Eiden, C.A., Jewell, C.A., Wightman, J.P. (1980), Interaction of lead and chromium with chitin and chitosan, Journal of Applied Polymer Science, 25, 1587-1599. 14. Engl, A., Kunz, B. (1995), Biosorption of Heavy Metals by Saccharomyces cerevisiae: Effects of Nutrient Conditions, J. Chem. Tech. Biotechnol., 63, 257-261. 15. Fourest, E., Canal, C., Roux, J.-C. (1994), Improvement of heavy metal biosorption by mycelial dead biomasses (Rhizopus arrhizus, Mucor miehei and Penicillium chrysogenum): pH control and cationic activation, FEMS Microbiology Reviews, 14, 325-332. 16. Ghosh, S., Bupp, S. (1992), Stimulation of biological uptake of heavy metals, Wat. Sci. Tech., 26, 227-236. 17. Gupta, R., Ahuja, P., Khan, S., Saxena, R.K., Mohapatra, H. (2000), Microbial Biosorbents: Meeting challenges of heavy metal pollution in aqueous solutions, Current Science, 78(8), 967-973. 18. Hansen H.K., Ottosen L.M., Villumsen A. (1997), J. Chem. Tech. and Biotechnol. 70: 67-73. 19. Holan, Z.R., Volesky, B. (1994), Biosorption of lead and nickel by biomass of marine algae, Biotechnology and Bioengineering, 43, 1001-1009. 20. Ledin, M. (2000), Accumulation of metals by microorganisms – processes and importance for soil systems, Earth-Science Reviews, 51, 1-31. 21. Mester, Z., Cremisini, C., Ghiara, E., Morabito, R. (1998), Comparison of two sequential extraction procedures for metal fractionation in sediment samples, Analytica Chimica Acta, 359, 133-142.
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22. Ottosen L. M., Hansen H. K., Laursen S., Villumsen, A. (1997), Electrodialytic Remediation of Soil Polluted with Copper from Wood Preservation Industry, Environ. Sci. Technol. 31(6): 1711-1715. 23. Ottosen, L.M., Hansen, H.K., Hansen, C.B. (2000), Water-splitting at ion-exchange membranes and potential differences in soil during electrodialytic soil remediation, Journal of Applied Electrochemistry, 30, 1199-1207. 24. Ottosen, L.M., Hansen, H.K., Ribeiro, A.B., Villumsen, A. (2001), Removal of Cu, Pb and Zn in an applied electric field in calcareous and non-calcareous soils, Journal of Hazardous Materials B85, 291-299. 25. Pedersen, A.J., Jensen, P.E.J. (1999), Electrodialytic remediation of Pb contaminated soil – effect of soil properties and Pb distribution, in: Proceedings of the 2.nd symposim, Heavy Metals in the Environment and Electromigration Applied to Soil Remediation, Technical University of Denmark, Lyngby, Denmark. 26. Ribeiro, A.B. (1998), Use of Electrodialytic remediation technique for removal of selected heavy metals and metals from soils, PhD.-thesis, Department of Geology and Geotechnical Engineering (now BYG-DTU), Technical University of Denmark, Denmark, 1998. 27. Volesky, B., Holan, Z.R. (1995), Biosorption of Heavy Metals, Biotechnol. Prog., 11, 235-250. 28. Volesky, B. (2001), Detoxification of metal-bearing effluents: biosorption for the next century, Hydrometallurgy 59, 203-216.
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Fundamentals of the uranium separation in constructed wetlands F. Glombitza, F. Karnatz, H. Fischer, J. Pinka and E. Janneck Department of Biotechnology, G.E.O.S. Freiberg Ingenieurgesellschaft mbH, Gewerbepark Schwarze Kiefern, D 09633 Halsbrücke, Germany Abstract The microbial process of the uranium separation from mine flooding and drainage water was investigated and the results will be demonstrated. Uranium separation takes place in constructed wetlands under anaerobic conditions due to microbial reduction of the U6+ into insoluble U4+. The concentration of uranium can be reduced to values lower than 0.3 mg/l. Competitive reactions take also part and influence the degree of separation efficiency as well as the solubility and mobility of the precipitated uranium. The influences of the microbial community composition as well as the mine water composition on the separation were demonstrated. Sulphate-reducing microorganisms are able and responsible to reduce uranium in contrast with denitrifying bacteria. The formed HCO3- as final product from the transfer of the organic carbon source into biomass is the reason for the formation of uranylcarbonate complexes and the release of the stored uranium. Highly uranium concentrated fraction can therefore be detected sometimes in the case of missing SRB and of unfavourable HCO3- concentrations. Recommendations for the risk assessment and the evaluation of the stability of a constructed wetland will be derived after the treatment of a pilot plant. 1.
INTRODUCTION Uranium mine drainage and process waters contain different hazardous substances in most cases. Depending on the pH of the water, uranium and radium are the most important contaminants among different heavy metals. Iron, manganese, zinc, and some other heavy metals exist as cations, arsenic, sulphate and carbonate as anions. The classical water treatment technology passes three steps, a) separation of Uranium by precipitation or ion exchange, b) separation of Radium by BaCl2 precipitation and c) separation of Arsenic as well as some heavy metals by addition of FeCl3 and precipitation of Fe-As-compounds by addition of NaOH [1-2]. The sludge formed is separated, mixed with cement to form stable compounds as concrete and stored in special disposals [3]. The cost of this kind of water treatment is very high and lies in a range from 5-15 €/m³ water, depending on the treated water volume. Thus, cheaper methods for the treatment are necessary. Investigations for the separation of uranium by natural reactions in wetlands are a precondition for the development and use of cheaper and safer techniques in future. 575
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2.
SUMMARY / COMPOSITION OF BASIC REACTIONS OF THE URANIUM SEPARATION The possibility to separate uranium in a wetland at anaerobic conditions is a known fact [4-6]. Knowledge of the different reactions, which take place in such a system, is a precondition to control the process and guarantee a long-time process stability. Summaries of some important possible reactions, which can take part in a wetland, are shown in table 1. Table 1. Some possible important uranium separation processes in an anaerobic wetland system
Microbiological reactions: SO42- + microorganisms + C org. NO3- + microorganisms + C org. U6+ + microorganisms (SRB) U6+ + microorganisms (DENI) Chemical reactions: U6+ + HS U6+ + S2Biosorption: U6+ + biomass
Æ Æ Æ Æ
HS- + biomass + HCO3- + OHN2 + biomass + HCO3- + OHU4+ (UO2) ? U4+ (UO2) ?
Æ Æ
U4+ + S° + H+ U4+ S2, U 6+ S3, UOS (?)
Æ
U - biomass
Microorganisms reduce sulphate and nitrate ions. Sulphide is one product of the reaction. It can be assumed, that the U6+-ion, which exists as uraniumsulphonyl or uraniumcarbonyl, is reduced and precipitated by means of the formed S2- or reduced and precipitated by the existing microorganisms [7-8]. Separation of Uranium by means of biosorption can also be taken into consideration as described by Tzesos [9]. 3.
INVESTIGATIONS OF THE REACTION
3.1 Uranium separation with the aid of sulphide and dithionite Uranium containing model solutions and mine waters were treated by the addition of increasing concentration of S2- as well as dithionite. Uranium precipitation could not be observed even after a long reaction time. Figures 1 and 2 show the behaviour of the uranium concentration by increasing concentration of sulphide and during a long reaction time. A decrease of the uranium concentration could not be observed in both cases.
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Figure 1. Influence of the dithionite concentration on the uranium concentration
Figure 2. Demonstration of the influence of the reaction time on the uranium concentration
3.2 Separation by means of sulphate-reducing microorganisms The influence of the sulphate-reducing microorganisms on the uranium concentration was investigated by experiments with living SRB in uranium containing sulphate free mediums. The uranium containing solution was inoculated by living SRB and the uranium concentration was determined. The same experiment was carried out with dead "autoclaved" cells to determine and take into consideration the possible influence of the biosorption on the decrease of the uranium concentration. The results of these experiments are demonstrated in figure 3.
Figure 3. Separation of uranium by SRB
The results in figure 3 demonstrate the ability of the SRB to separate uranium. U6+ is reduced to U4+, which precipitates as black compound, and was specified by X–Ray as Uraninite. 4.
REACTION IN A CONTINUOUS FLOWED WETLAND
4.1 Continuous column tests The use of SRB for the reduction of uranium in a wetland process requires anaerobic conditions in the reaction basin. For this purpose, column tests were carried out for a period of some years. Conditions and results are summarised in Tables 2, 3 and 4. 577
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Table 2. Process conditions of a long time column process Parameter Volume of the column Graves Volume of water in the column Water flow rate Residence time DOC concentration in the water (CH3OH) Uranium concentration in the drainage water Uranium concentration in the treated water
Number and Dimension 4,590 ml 5,277 g 1,720 ml 36 ml/h 48 h 80 – 100 mg/l 1 – 2 mg/l < 0. 3mg/l
Table 3. Consumption coefficients Component
Dimension
mg/l C - limited
mg/l NH4 - limited
Consumption of P
mg/l
0.65
0.13
Consumption of DOC
mg/l
104.4
89.6
-
mg/l
4.62
2.57
-
mg/l
143
216.87
+
mg/l
1.12
-
mg/l
1.99
0.61
mg/l
884
388.6
Formation of C
mg/l
173.9
76.45
Reduction of U
mg/l
0.84
1.24
Consumption of NO3 Consumption of SO4
Consumption of NH4
Consumption of total N Formation of HCO3
-
Table 4. Specific coefficients Kind of coefficients
Dimension
Amounts C – limited
Amounts without NH4 NH4 – limited
Related on DOC
g SO4/g DOC mg N/g DOC mg P/g DOC mg U/g DOC mg DOC/g SO4 mg N/g SO4 mg P/g SO4 mg U/g SO4 mg DOC/mg U mg N /mg U mg P/mg U mg SO4/mg U
1.37 19.06 6.22 8.04 730.07 13.92 4.55 5.87 124.29 2.37 0.77 170.24
2.93 6.9 1.49 13.93 340.99 2.33 0.51 4.75 71.77 0.49 0.107 210.48
Related on SO4
Related on U
Table 3 and 4 contain the obtained consumption and specific consumption coefficients for the separation and the reduction per unit sulphate and the consumed carbon. 340-730 mg of DOC are able to reduce 1 g sulphate. The formed biomass can separate between 4.8-5.8 mg of uranium simultaneously depending on the reaction 578
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conditions. 71-124 mg of DOC are required for the separation of 1 mg uranium. The process conditions were changed for the simulation of possible disturbances and for checking different situations concerning the stability. The question turned up, what happens after a break or during the phase of restarting the process. The continuous process in a column was stopped and the column was flooded only with drinking water for some month. Drainage water was sprinkled again after this time and the concentration in the effluent as well as the composition of the microbial community and the CFU values were determined. Figure 4 shows the alteration of the U, DOC and HCO3- concentrations.
Figure 4. Representation of the start situation of a continuous process after a longtime break
Figure 5 shows the alteration of the sulphate concentration due to the reduction to sulphide. The results demonstrate a decrease of the DOC in the treated water due to its consumption and consequently an increase of the HCO3- concentration.
Figure 5. Alteration of the sulphate concentration
The uranium concentration increases in the treated water during the initial step. A release of uranium as [UO(CO3)3]2- uranylcarbonate by the formed HCO3 can be assumed and is possible [8]. A decrease of the Uranium concentration can be observed in the next steps although the HCO3- concentrations reach the same values. The analysis of the 579
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microbial population and the behaviour of the CFU of the denitrifying and sulphatereducing bacteria are therefore investigated. Results are demonstrated in Figure 6. A rapidly increasing number of denitrifying bacteria can bee seen followed by an increase of the number of sulphate-reducing bacteria some days later.
Figure 6. Representation of the development of the denitrifying and sulphatereducing bacteria
Figure 7 shows the CFU-values and the alteration of the uranium concentration. This figure reveals a decrease of the uranium concentration in that moment where the sulphatereducing bacteria turned up in the system. That means: only sulphate-reducing bacteria are responsible for the uranium separation and the decrease of the concentration. Denitrifying bacteria do not reduce and separate uranium.
Figure 7. Connection between uranium separation and the alteration of the CFU values of denitrifying and sulphate-reducing bacteria
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Figure 8. Separation of the uranium by means of the sulphate-reducing bacteria 4.2 Practical results and practical experiences from a real wetland These results are used for the controlled removing of uranium from drainage water in a constructed wetland. Main part of the wetland is an anaerobic reaction chamber with a total volume from 150 m³ and a volume of free water around 50-70 m³. The water residence times lie in a range from 50 to 100 h depending on the water flow rate. The plant was started in the summer of 2001. The separation process is based on the cultivation of sulphate-reducing bacteria for uranium reduction. Figure 9 shows the separation of uranium up to the point of the total consumption of the usable carbon sources. Uranium separation could not be observed and the process collapsed after this point. DOC values are lower than 10 mg/l. Figure 10 shows the situation after starting a carbon dosing followed by increasing number of SRB. Uranium concentration in the treated water was decreased to a value lower than 0.3 mg/l as a result of the cultivation of the sulphatereducing microorganisms.
Figure 9. Uranium concentration at different places in the reduction chamber and collapsing of the uranium separation due to a lacking of a carbon source (PB: Mine drainage water, S2o, S2m, S2u different sampling points)
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Figure 10. Representation of the situation of uranium separation after dosing of carbon sources (PB mine drainage water, RK/S3 sampling point at the end of the reduction chamber) 5.
CONCLUSION Uranium can be separated in constructed wetlands at anaerobic conditions. The separation takes place by reduction of the U6+ by means of sulphate-reducing bacteria. Denitrifying bacteria do not separate uranium by the same mechanism. In the case of missing of SRB Uranium can be delivered by the microbial formed HCO3- and an increase of the concentration can be observed. Constant conditions concerning the growth of SRB are a precondition of stable process conditions and stable storage of the precipitated and stored uranium therefore. ACKNOWLEDGEMENTS The German Ministry for Education, Science, Research and Technology (BMBF FKZ 02 WB 0104), and the EU in the 5th frame programme (Project number: EVK1 - 1999 00168P), which are gratefully acknowledged, support this work. REFERENCES
1. G. Kießig, Ch. Kunze, Wasserbehandlung und Rückstandsentsorgung Geowissenschaften 14 (1996) 11 481-485. 2. M. Hüttl, H. Weinl, D. Laubrich Neue Wasserbehandlungsanlage für untertägige Sanierung des Grubenfeldes Ronneburg, WLB Wasser, Luft, Boden Zeitschrift für Umwelttechnik 11-12, 2002 pp. 35-37. 3. G. Kießig Erfahrungen aus realisierten Wasserbehandlungsvorhaben bei Wismut Tagungsband: 7. Wismut – Workshop of Water and sludge treatment, Konventionelle und innovative Lösungen 24-26.09.1997, Chemnitz. 4. H. Nisbet Wetland filtration research at ERA Ranger mine Wetland research in the wet dry tropics of Australia Workshop Jabiru NT 22-24 March 1995 p 165-72 Ed. by CM Finlayson.
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5. St. Shinners An Overview of the application of constructed wetland Filtration at ERA Ranger mine, National engineering conference Darwin, 21-24 April 12996, Reprints of Papers, pp.367-378. 6. F. Glombitza, E. Janneck, F. Karnatz, G. Kießig, A. Küchler, M. Hüttl, Die Anwendung naturnaher Wasserbehandlung am Beispiel der Pilotanlage Pöhla. Natural Attenuation, neue Erkenntnisse, Konflikte, Anwendungen, Resümee und Beiträge zum 2. Symposium Natural Attenuation 07-08.12.2000 Ed. DECHEMA Gesellschaft für chem. Technik e.V. 2001 pp. 220-21, ISBN 3-89764-021-1. 7. W. Heymel Prinzipien und Methoden der technischen Uranfällung Ed. Wismut 1963. 8. Lovley D.R., Phillips E. J. P., Gorby Y.A., Landa E.R., (1991) Microbial reduction of uranium, Letters To Nature, Vol. 350, 4 April, 413-416. 9. M. Tsezos Engineering Aspects of Metal binding by biomasses in: Microbila minerla Recovery Eds.: H.L. Ehrlich, C. L. Brierly, Mc Graw Hill Inc., 1990, pp. 325-339, ISBN 0-07-007781-9.
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Geomicrobiological risk assessment of abandoned mining sites K. Bosecker, G. Mengel-Jung and A. Schippers Federal Institute for Geosciences and Natural Resources (BGR), Section Geomicrobiology, Stilleweg 2, D 30655 Hannover, Germany e-mail:
[email protected], A.
[email protected] Abstract Some microorganisms colonize mineral surfaces, leading to acidification of the water draining from the mine or waste heap (AMD/ARD) and to contamination of the water with heavy metals. Depending on the mineral composition of the mine tailings and waste rock and on the environmental conditions (temperature, humidity, oxygen availability) the microbial activity can cause major environmental problems. Acid mine drainage from mine wastes may be predicted by investigating the site for the presence of metal-mobilizing bacteria (e.g. iron-oxidizing acidophiles, sulfur-oxidizing acidophiles, sulfur-oxidizing neutrophiles, acidophilic heterotrophs), quantifying the various indigenous populations, and measuring their leaching activity. We have studied mine waste dumps in a number of countries – with different metal contents and in different climate zones. Because bacterial leaching occurs mostly in sulfide ores, research has been concentrated on this kind of sulfide waste dump in Zimbabwe, Namibia, Bolivia, Peru, Brazil, Chile, Cuba and Kazakhstan. Metal-mobilizing bacteria were identified in all of the samples. The efficiency of the isolated bacteria and their significance for the mobilization of toxic substances from the waste dumps were determined. Possibilities for permanently minimizing environmental contamination of mining areas are shown. Keywords: acid mine drainage, mining waste, environmental risk potential, quantitative ecology, iron-oxidizing acidophiles, sulfur-oxidizing acidophiles, sulfur-oxidizing neutrophiles, acidophilic heterotrophs 1.
INTRODUCTION The mining and processing of metals creates spoil and waste heaps everywhere in the world. Some of the material in these heaps, which often have a volume of several million cubic meters, contains low concentrations of metallic ore. Sometimes the metal concentration may be remarkably high in the ore. Remediation measures are seldom carried out on these waste heaps to integrate them into the landscape. As a rule – at least in former times – no measures are taken when the mine is shut down and the waste heaps are left to be affected by weathering. Rainwater and residual water in the heaps dissolves the metals and sulfuric acid is produced. The metals and acid are transported into the surface water and groundwater, a process known as acid-mine drainage (AMD) or acid-rock drainage (ARD). Mobilization of the metals and generation of acid in the waste result 585
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from chemical and microbiological oxidation. The conditions under which bacteria mobilize metals and other toxic substances are fairly well understood [1-4], but the extent to which this occurs at abandoned mining sites is not very well known. The bacterial activity depends on the mineral composition of the waste material, the sulfide, oxide or carbonate content and on the environmental conditions, such as pH, temperature, moisture content, and the availability of oxygen. In principle, the microbial leaching processes used at an industrial scale to recover metals from low-grade ores, refractory ores or concentrates, also occur in old mine dumps and at abandoned mining sites [5]. AMD is the most severe environmental problem the mining industry is faced with. It leads to contamination of the surface water and groundwater with toxic metals and degrades the groundwater quality downstream from the heaps and dumps. In the most extreme case reported, the Richmond Mine of the Iron Mountain, California, AMD/ARD contained metal concentrations as high as 200 g/L, sulfate concentrations as high as 760 g/L and had a negative pH as low as -3.6 [6]. For this reason, the Federal Ministry for Economic Cooperation and Development (BMZ) commissioned BGR to study bacterial mobilization of metals in mine waste heaps and to find ways to minimize groundwater contamination in order to improve the water supplies of the local population. During the past three years BGR has studied mine waste dumps in a number of countries – dumps with different metal contents and in different climate zones. Because bacterial leaching occurs mostly in sulfide ores, research has been concentrated on this kind of waste dump. Mining wastes in Bolivia, Brazil, Chile, Cuba, Kazakhstan, Namibia, Peru and Zimbabwe have been studied. The objective was to obtain enough information to predict the contamination load on the basis of the type of ore and the climatic conditions and to determine the conditions needed to permanently minimize contamination of surface water and groundwater. The solutions to the problems will be adapted to the economic and social conditions in the developing countries. The developing countries are to be given the basis for treatment of their waste heaps. 2.
MATERIALS AND METHODS
2.1 Sampling The sampling sites, the number of samples and the respective climate conditions are given in Table 1. Table 1. Sampling sites, number of samples and climate conditions Country Bolivia Brazil Chile Cuba Kazakhstan
Mines 8 2 3 1 1 2 2
Samples 43 5 12 4 6 4 20
Namibia Peru Zimbabwe (Total)
1 7 4 (31)
16 31 24 (165)
586
tropical tropical tropical subtropical subtropical tropical temperate warm tropical tropical tropical
Climate cold warm warm continental warm warm continental
semi-humid semi-humid arid semi-humid semi-humid semi-arid semi-arid
warm cold warm
semi-arid humid semi-arid
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After consultation with the mining geologists, samples were taken preferentially from sulfidic waste heaps more than five years old. Depending on the local conditions and the accessibility of the waste heaps, mostly the lower slopes of heaps were chosen for sampling. In some cases, additional sampling was carried out at flotation tailings ponds. After removing the upper 10-20 cm of the surface layer, approximately 250-300 g of waste material with a particle size less than 25 mm was collected in 125 mL sterile plastic jars. Before transport to the laboratory in Hannover, the jars were kept at ambient temperature, and from time to time the screw lids were carefully loosened for sufficient aeration. For transport by air, the jars were kept in the hand luggage and taken into the cabin. Special permits were needed to pass the security checks without the samples being x-rayed. In the extreme case, it was three weeks until processing was started in the laboratory. 2.2 Analytical measurements Five grams of each sample was suspended in 13.5 mL of 1 M KCl. The fluid was shaken for 5 min at 130 rpm/min at room temperature. The pH was measured in the supernatant after 2 hours and after 24 hours. Portions of the samples were dried to a constant weight using a Sartorius MA50 moisture analyzer. After grinding of the samples, major and trace elements were analyzed by x-ray fluorescence spectrometry (XRF). On-site, pH-value and concentrations of iron(III) and copper in seepage water from waste rock heaps were measured using analytical test strips (Merck). 2.3 Detection and quantification of microbial groups related to metal mobilisation To detach bacteria from the surface of the solid material, 10 g of fresh sample was placed in 100 mL of Leathen medium [7] without iron and shaken at 130 pm for 2h at room temperature. After settling of the solid phase, the supernatant was used for total- and viable-cell counting. The total number of cells was determined by acridine orange direct count using epifluorescence microscopy. The number of the viable cells of acidophilic sulfur-oxidizing, acidophilic iron-oxidizing and moderately acidophilic autotrophic bacteria was quantified by the "most probable number technique" (MPN), using serial 10fold dilutions in three tubes containing the suitable growth medium. The tubes were incubated on a shaking table at 120 rpm for 4-6 weeks in the dark at 30°C. Growth of acidophilic sulfur–oxidizing bacteria was indicated by high acid production in Starkeymedium [8]. Acidophilic iron-oxidizing bacteria were considered to be present if the Leathen-medium [7] became reddish-brown. Growth of moderately acidophilic autotrophic bacteria was indicated by a decrease in pH of more than one pH unit using the growth medium described by Matin and Rittenberg [9]. The presence of bacteria was demonstrated by microscopy. Acidophilic and neutrophilic heterotrophic microorganisms were counted by serial 10-fold dilution and spread-plating on agar as described by Harrison [10] and on R2Aagar [11], respectively. The agar plates were incubated in the dark at 30°C for 4-6 weeks, depending on when the number of colonies remained constant. 2.4 Calorimetric measurements The bioleaching activity of the microorganisms was determined by a microcalorimetric technique [12-17]. Heat is produced in mine waste heaps due to the oxidation of pyrite. Oxidation of pyrite to Fe(III) and sulfate has a reaction energy of – 1546 kJ/mol. The measured heat output correlates with the number and activity of the 587
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leaching bacteria in the sample and the rate of pyrite oxidation can be calculated from these values. Heat output due to chemical oxidation was measured in control experiments. The calorimetric measurements were carried out in the laboratory of W. Sand, Department of Microbiology, University of Hamburg, Germany. 3.
RESULTS AND DISCUSSION The results for waste rock and tailings material from the different countries are presented separately in Tables 2-7. Large differences between waste rock and tailings material were observed with respect to the concentrations of the main metals. The metal content in waste rock material is generally higher than that in tailings material. Iron is the dominant metal (Tables 2 and 3). The waste rock from Bolivia and Peru contained very high concentrations of the listed main metals, and high contents of Cu and Zn were also detected in material from other countries (Table 2). Table 2. Main metals [g/kg] in waste rock Bolivia Brazil Chile Cuba Kazakhstan Namibia Peru Zimbabwe
As 6.2 4.1 <1 1.6 <1 < 0.1 27.7 18.4
Cu 4.2 <1 7.4 3.0 18.6 42.1 97.6 4.4
Fe 170 354 83 224 68 215 437 173
Pb 20.5 < 0.1 < 0.1 22.7 2.4 2.0 37.7 < 0.1
Sb 93.7 < 0.1 < 0.01 <1 < 0.01 < 0.01 8.5 < 0.1
Sn 30.5 < 0.1 < 0.1 < 0.01 < 0.01 < 0.01 2.3 < 0.01
Zn 26.5 <1 1.3 1.3 9.0 11.3 29.3 <1
Table 3. Main metals [g/kg] in tailings Bolivia Brazil Chile Cuba Kazakhstan Namibia Peru Zimbabwe
As 1.3
Cu < 0.1
Fe 137
Pb 1.4
Sb <1
Sn 3.5
Zn 2.2
<1 <1 1.5 < 0.1 3.5 <1
6.5 <1 2.0 1.6 <1 1.9
45 101 127 232 337 115
< 0.1 < 0.01 1.6 <1 7.4 <1
< 0.1 < 0.1 < 0.1 < 0.01 <1 < 0.01
< 0.1 < 0.01 < 0.01 < 0.01 <1 < 0.01
< 0.1 <1 5.7 7.4 27.4 < 0.1
The maximum number of total bacteria ranged between 106 and 108 bacteria per gram waste rock or tailings material from all countries (Table 4 and 5). A significant proportion of the total bacteria were the acidophilic Fe(II)- and sulfur-oxidizing bacteria, which oxidize metal sulfides and their degradation products (sulfur compounds) at low pH. High numbers of these bacteria were detected in waste rock material from all countries and in the tailings material from almost all countries. The low number of acidophilic bacteria in the rock samples from Brazil may be explained by the slightly alkaline condition in the sample (Table 4 and 6). Neutrophilic sulfur-oxidizing bacteria oxidize the sulfur compounds formed by chemical sulfide oxidation at near neutral pH. These bacteria were detected in waste rock material from all countries, but they were not detected in tailings material from Bolivia and Zimbabwe. The lowest numbers of neutrophilic S-oxidizers were found in the acidic waste rock material from Bolivia and Cuba with pH-values of 1.9 588
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and 2.0, respectively (Table 4 and 6). Acidophilic heterotrophic bacteria which degrade organic compounds, thus permitting the continued growth of bioleaching bacteria, were found in some cases. In waste rock material, maximum numbers of these cells were in the range of 103-106 cells/g dry sample, tailings material contained less acidophilic heterotrophic bacteria. Table 4. Maximum number of bacteria (N/g) in waste rock
Bolivia Brazil Chile Cuba Kazakhstan Namibia Peru Zimbabwe n.d.: not detected
Total bacteria 1.3 x 108 2.5 x 108 4.6 x 107 5.1 x 106 5.2 x 107 9.8 x 107 2.2 x 107 2.3 x 108
S-oxidizers acidophile 1.2 x 107 5.2 x 103 2.6 x 106 2.8 x 106 4.9 x 106 > 1.0 x 107 1.2 x 107 1.1 x 107
Fe(II)-oxidizers acidophile 5.1 x 106 2.8 x 102 1.2 x 105 2.9 x 106 1.2 x 107 2.2 x 107 3.1 x 108 1.1 x 107
S-oxidizers neutrophile < 102 7.9 x 106 2.1 x 104 7.1 x 102 2.8 x 106 1.2 x 107 1.2 x 107 1.1 x 107
Heterotrophs acidophile 7.5 x 103 n.d. n.d. 2.7 x 105 1.9 x 104 9.4 x 104 n.d. 2.6 x 106
Table 5. Maximum number of bacteria (N/g) in tailings
Bolivia Chile Cuba Kazakhstan Namibia Peru Zimbabwe n.d.: not detected
Total bacteria 7.8 x 106 1.5 x 107 1.4 x 106 5.9 x 107 1.4 x 108 3.0 x 107 7.8 x 107
S-oxidizers acidophile n.d. 3.7 x 105 2.8 x 104 3.7 x 106 1.3 x 107 1.4 x 107 4.5 x 106
Fe(II)-oxidizers acidophile 2.7 x 102 2.8 x 106 n.d. 1.3 x 108 > 1.3 x 107 3.1 x 106 1.1 x 107
S-oxidizers neutrophile n.d. 1.3 x 106 2.8 x 103 1.2 x 106 1.3 x 107 1.4 x 107 n.d.
Heterotrophs acidophile 5.2 x 102 n.d. n.d. n.d. 2.7 x 102 n.d. 2.8 x 103
Table 6. Maximum leaching activity in waste rock, corresponding pH and number of acidophilic S- and Fe(II) oxidizers P total [µW/g] Bolivia 17 Brazil <5 Chile <5 Cuba 22 Kazakhstan 23 Namibia 17 <5 Peru 89 0 Zimbabwe n.i. n.i.: not investigated
P biotic [%] 88 n.i. 79 77 83 0 89 0
pH 1.9 8.3 4.3 2.0 2.3 2.9 6.7 2.5 6.9
Acid. S-oxidizers [N/g] 8.2 x 103 n.d. 2.2 x 106 2.8 x 106 2.7 x 104 4.9 x 106 1.0 x 104 3.1 x 106 n.d.
Acid. Fe(II)-oxidizers [N/g] 2.7 x 103 n.d. 1.2 x 105 2.8 x 106 1.2 x 107 2.2 x 107 4.7 x 104 7.0 x 105 2.6 x 102
n.d.: not detected
Waste rock material from Peru and tailings material from Kazakhstan and Namibia showed very high total leaching activity ( >50 µW/g sample), and more than 75% of the sulfide oxidation was due to bacterial activity (Table 6 and 7). In agreement with the high 589
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bioleaching activity, a pH below 3 and high numbers of acidophilic Fe(II)- and sulfuroxidizing bacteria were detected for these samples. Significant microbiological sulfide oxidation was also detected in waste rock material from Bolivia, Cuba, Kazakhstan, and Namibia, and in tailings material from Peru. Table 7. Maximum leaching activity in tailings, corresponding pH and number of acidophilic S- and Fe(II) oxidizers P total [µW/g] Bolivia 7 Brazil n.i. Chile <5 Cuba <5 Kazakhstan 57 <5 Namibia 70 56 Peru 70 38 Zimbabwe n.i. n.i.: not investigated
P biotic [%] 0 n.i. 0 75 0 93 19 5 21
2.4
Acid. S-oxidizers [N/g] n.d.
Acid. Fe(II)-oxidizers [N/g] 2.7 x 102
3.9 7.5 2.3 6.9 2.2 6.3 6.6 6.6
5.1 x 103 2.8 x 104 5.2 x 105 n.d. 4.9 x 106 n.d. 1.4 x 107 5.6 x 104
2.8 x 106 n.d. 1.3 x 108 1.0 x 102 > 1.2 x107 n.d. 3.1 x 106 n.d.
pH
n.d.: not detected
The results confirm that AMD/ARD in mine waste and tailings is caused by the bioleaching activity of microorganisms. High numbers of bioleaching bacteria and high bioleaching activity were detected worldwide. Surprisingly, high numbers of bioleaching microorganisms were detected in tightly packed tailings material. The influence of climate conditions on the contamination potential of abandonded mine sites still needs to be assessed. Mitigation of the environmental impact from mine waste must include measures to inhibit bioleaching bacteria. 4.
RECOMMENDATIONS FOR PREVENTION OF ENVIRONMENTAL DAMAGE Microbiological studies of mining heaps in different countries have provided information about the potential for damage to the environment from mine waste and tailings heaps. Countermeasures are to be developed on the basis of the results. Bioleaching, the main reason for AMD-generation on the one hand, has some potential for remediation of abandoned mines on the other hand. Heavy metalcontaminated mine sites can be remediated by using adapted, sophisticated heap and dump leaching technologies where optimum conditions for the growth of the leaching microorganisms are kept constant and any seepage of the leachate is prevented. In this way, both processes, remediation of mining sites and recovery of valuable metals, may be achieved simultaneously [46]. Measures to inhibit bioleaching include underwater storage of the waste, covering of the heaps, application of inhibitory substances to the waste, and planting of vegetation on the heaps [18-22]. The importance of microorganisms has been highlighted by Ledin and Pedersen [5]. During the last several years, extensive programmes to evaluate measures for mitigating the environmental impact of mine waste have been carried out in Canada [23] 590
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and in Sweden [24]. The mining industry initiated "The International Network for Acid Prevention" in 1998 [25]. Underwater storage of the waste is currently a frequently applied measure. Due to reduced oxygen diffusion, a significant reduction of the number of bioleaching bacteria and consequently of the sulfide oxidation rate can be achieved [15, 26-27]. Underwater storage of the waste may require expensive moving of the waste material to a new site, which is currently being done in Germany by the Wismut GmbH [28]. The most important consequence of moving the waste material is that the sulfide is oxidized by increased contact of the material with air. Thereby a significant portion of the pyrite may be oxidized to Fe(III) compounds. During the subsequent storage under water, heavy metals may be mobilized by Fe(III)-reducing bacteria since large amounts of heavy metals are adsorbed on Fe(III) oxide particles [13, 29]. Detailed studies of anaerobic mobilization of metals in mine waste are lacking. Dry cover, e.g. loamy soil or a synthetic material such as high density polyethylene (HDPE) reduces or prevents the penetration of air and water into the heap or tailings pond and decreases the oxidation of metals, which requires oxygen. In addition, a dry cover prevents the spreading of heavy-metal-bearing dust and facilitates the growth of vegetation. There are many requirements for dry cover making efficient dry cover expensive [28, 30]. A simple dry cover of loamy soil reduces the oxidation rate, but does not prevent microbial oxidation of sulfide [31, 32]. Backes et al. [33] and Bennett et al. [34] showed a significant reduction of the concentration of oxygen inside heaps after covering. A microbiological study of uranium mine waste heaps of the Wismut GmbH in Germany showed that the number of bioleaching bacteria and the oxidation rate were lower in a covered part of a heap than in an uncovered part [13]. Substances that raise the pH may inhibit bioleaching, degrading the conditions for growth of bioleaching bacteria [35]. Other substances, called biocides, kill bacteria, still others passivate the pyrite surface. Lime, which is quite expensive, is often used to raise the pH and to buffer infiltrating water. Application of the anionic detergent sodium dodecylsulfate (SDS) has shown to be efficient and inexpensive in a field study [36, 37]. A lysimeter study showed that after several SDS applications on an already heavily weathered mine waste heap, the number of bioleaching bacteria was reduced [35], but not the leaching of heavy metals [38]. SDS is easily washed out or degraded, thus application as pellets has been recommended [37]. Field studies with isothiazolinone showed a significant reduction of the amount of leached heavy metals [38]. Pilot experiments using sodium paraffin sulfonate (E30) or calcium fluoride resulted in a reduction of bioleaching activity [39]. In laboratory experiments, bioleaching was significantly inhibited after application of benzoic acid and sorbic acid [40], thymole [31], fatty acid amines [41], fulvic acids, tannic acid, and oxalic acid [42]. These substances cannot only inhibit bioleaching bacteria; they also support the growth of organotrophic microorganisms which compete with the bioleaching bacteria for oxygen. Organic substances may be inexpensive, e.g., sewage sludge, compost, chicken manure, conifer bark, oil shale waste could be used. In a field study, bioleaching activity was significantly reduced after addition of organic substances [20, 33, 38]. But the addition of organic substances may also enhance the formation of metal-organic complexes or the reduction of Fe(III) oxides, which may increase the mobilization of heavy metals (see above). Studies concerning this aspect are lacking.
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Other inhibitory substances inactivate or encapsulate the pyrite particles, but so far only laboratory studies are known [20, 43-44]. Planting inhibits bioleaching twice: Firstly, plant roots compete with the bioleaching bacteria for oxygen, and secondly, plant roots produce exudates which are inhibitory organic substances. Planting directly on the uncovered surface of a mine waste or tailings heap is problematic because plants usually do not tolerate low pH [5]. Bioleaching may kill the plants even several years after planting [45]. For these reasons planting should be combined with covering and/or application of lime. A combination of inhibitory measures has to be taken into consideration in each case. Efficiency and cost are the most important criteria for the selection of measures. An evaluation of the geomicrobial risk potential should be carried out to select the most efficient measure, as has been done in this study. ACKNOWLEDGEMENTS This project was financed by the Federal Ministry for Economic Cooperation and Development (BMZ) as part of technical cooperation. We gratefully acknowledge the technical assistance of D. Spier, F. Korte, J. Lodziak and D. Requard. The calorimetric measurements were carried out by Th. Rohwerder and K. Kinzler in the laboratory of W. Sand. REFERENCES
1. K. Bosecker, FEMS Microbiol. Rev., 20 (1997) 591. 2. H. Brandl, in: H.-J. Rehm, G. Reed, A. Pühler, and P. Stadler (eds.), Biotechnology Vol. 10, Wiley-VCH, Weinheim, 2001. 3. G. Rossi, Biohydrometallurgy, McGraw-Hill Book Comp., Hamburg, 1990. 4. D.E. Rawlings, Ann. Rev. Microbiol., 56 (2002) 65. 5. M. Ledin and K. Pedersen, Earth-Sci. Rev., 41 (1996) 67. 6. D.K. Nordstrom and C.N. Alpers, PNAS USA, 96 (1999) 3455. 7. W.W. Leathen, L.D. McIntyre and S.A. Braley, Sci., 114 (1951) 280. 8. R.L. Starkey, J. Bacteriol., 10 (1925) 135. 9. Matin and S.C. Rittenberg, J. Bacteriol., 107 (1971) 179. 10. A.P. Harrison Jr., Ann. Rev. Microbiol., 38 (1984) 265. 11. D.J. Reasoner and E.E. Geldreich, Appl. Environ. Microbiol., 49 (1985) 1. 12. W. Sand, R. Hallmann, K. Rohde, B. Sobotke, and S. Wentzien, Appl. Microbiol. Biotechnol., 40 (1993) 421. 13. Schippers, R. Hallmann, S. Wentzien, and W. Sand, Appl. Environ. Microbiol., 61 (1995) 2930. 14. T. Rohwerder, A. Schippers and W. Sand, Thermochimica Acta, 309 (1998) 79. 15. B. Elberling, A. Schippers and W. Sand, J. Cont. Hydrol., 41 (2000) 225. 16. Schippers, P.-G. Jozsa, W. Sand, Z.M. Kovacs and M. Jelea, Geomicrobiol. J., 17 (2000) 151. 17. W. Sand, T. Gehrke, P.-G. Jozsa and A. Schippers, Hydrometallurgy, 59 (2001) 159. 18. I.D. Pulford, in: M.C.R. Davies (ed.), Land reclamation, 269, Elsevier, Amsterdam, 1991. 19. Proceedings of the International Land Reclamation and Mine Drainage Conference and of the Third International Conference on the Abatement of Acidic Drainage, Pittsburgh, Pennsylvania, 1994.
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20. V.P.B. Evangelou, Pyrite oxidation and its control, CRC Press, Boca Raton, Florida, 1995. 21. Reviews in Economic Geology, Vol. 6A and 6B, Society of Economic Geologists, Inc., Littleton, Colorado, 1999. 22. Proceedings of the Fifth International Conference on Acid Rock Drainage, Society for Mining, Metallurgy and Exploration, Inv., Littleton, Colorado, 2000. 23. MEND 2000, www.nrcan.gc.ca/mets/mend/. 24. MiMi, www.mimi.kiruna.se. 25. www.inap.com.au/inap/homepage.nsf. 26. P. St-Germain, H. Larratt and R. Prairie, in: Proceedings of the Fourth International Conference on Acid Rock Drainage, Vol. I, 131, Vancouver, B.C. Canada, 1997. 27. P.H. Simms, E.K. Yanful, L. St-Arnaud and B. Aubé, Appl. Geochem., 15 (2000) 1245. 28. R. Gatzweiler, S. Jahn, G. Neubert and M. Paul, Waste Management, 21 (2001) 175. 29. D.E. Cummings, A.W. March, B. Bostick, S. Spring, F. Caccavo Jr., S. Fendorf and R.F. Rosenzweig, Appl. Environ. Microbiol., 66 (2000) 154. 30. B. Steinert, S. Melchior, K. Burger, K. Berger, M. Türk and G. Miehlich, Hamburger Bodenkundliche Arbeiten (in German), Band 32, Institute for Soil Science, University of Hamburg, Germany, 1997. 31. M. Silver and G.M. Ritcey, Hydrometallurgy, 15 (1985) 255. 32. D.K. Gibson and G. Pantelis, in: D.S. Brown and D.P. Hodel (eds.), Mine drainage and surface mine reclamation, 248, Bureau of Mines, US Department of the Interior, 1988. 33. C.A. Backes, I.D. Pulford and H.J. Duncan, in: D.S. Brown and D.P. Hodel (eds.), Mine drainage and surface mine reclamation, 91, Bureau of Mines, US Department of the Interior, 1988. 34. J.W. Bennet, J.R. Harries, G. Pantelis and A.I.M. Ritchie, in: J. Sally, R.G.L. McCready and P.L. Wichlacz (eds.), Biohydrometallurgy, 551, Canmet, Ottawa, 1989. 35. A. Schippers, P.-G. Jozsa, Z.M. Kovacs, M. Jelea and W. Sand, Waste Management, 21 (2001) 139. 36. B. Splittorf and V. Rastogi, in: Proceedings of the 1995 Annual Meeting of the American Society for Surface Mining and Reclamation, 471, Gillette, Wyoming, 1995. 37. R.L.P. Kleinmann, in: K.C. Brady, M.W. Smith and J. Schuck (eds.), Coal mine drainage prediction and pollution prevention in Pennsylvania, Chapter 15, The Pennsylvania Department of Environmental Protection, 1998. 38. P.-G. Jozsa, A. Schippers, W. Sand, Z.M. Kovacs, A.A. Nagy and M. Jelea, in: V.S.T. Ciminelli and O. Garcia Jr. (eds.), Biohydrometallurgy: Fundamentals, Technology and Sustainable Development, Part B, 297, Elsevier, Amsterdam, 2001. 39. F. Glombitza, U. Iske, M. Bullman and J. Ondruschka, Acta Biotechnol., 12 (1992) 79. 40. S.J. Onysko, R.L.P. Kleinmann and P.M. Erickson, Appl. Environ. Microbiol., 48 (1984) 229. 41. K. Nyavor, N.O. Egiebor and P.M. Fedorak, The Science of the Total Environment, 182 (1996) 75. 42. K. Sasaki, M. Tsunekawa, S. Tanaka and H. Konno, Colloids and Surfaces A: Physicochemical and Engineering Aspects, 119 (1996) 241. 43. V.P.B. Evangelou, Ecological Engineering, 17 (2001) 165. 44. K. Nyavor and N.O. Egiebor, The Science of the Total Environment, 162 (1995) 225. 45. A.Schippers, P.-G. Jozsa, W. Sand, Z.M. Kovacs and M. Jelea, Geomicrobiol. J., 17 (2000) 151. 46. K. Bosecker, Hydrometallurgy, 59 (2001) 245.
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Immobilisation and growth of Acidithiobacillus ferrooxidans on refractory clay tiles E. Donati, L. Martínez, G. Curutchet Centro de Investigación y Desarrollo de Fermentaciones Industriales (CINDEFICONICET), Facultad de Ciencias Exactas, Universidad Nacional de La Plata, 47 y 115 (1900) La Plata Abstract Refractory clay tiles are composed of kaolin, which is the commercial name of clay composed mainly of kaolinite mineral. Two of those tiles and caowool packed in glass columns were used for the immobilisation of Acidithiobacillus ferrooxidans cells in a ferrous iron medium, which was percolated through the supports. Colonisation was carried out by several medium replacements with no further inoculation until maximum ferric iron productivity was reached. One of those tiles was discarded due to the high iron precipitation during bacterial growth. The columns with the other supports were used for ferrous iron oxidation in batch and continuous flow modes of operation seeming to be promising supports for A. ferrooxidans. A ferrous iron oxidation rate of 14.5 mmol.l-1.h-1 was reached in one of the columns in continuous culture. After using for several cultures, pieces of tiles with immobilised cells were stored at 4°C. Samples at different times were incubated in ferrous medium showing high cell activity even after 6 months. Keywords: Acidithiobacillus ferrooxidans, immobilization, storage, clay tiles 1.
INTRODUCTION Acidithiobacillus ferrooxidans is a chemolithoautotrophic acidophilic bacterium with great importance in biohydrometallurgy. This microorganism is involved in the bioleaching of metal sulphide ores basically due to its iron (II) oxidising capacity. Metal sulphides are dissolved by the oxidising action of iron(III), which is continually regenerated by the cells. Iron(III) production by bacterial action can also be employed in bioremediation processes [1,2]. The natural tendency of A. ferrooxidans to grow on surfaces makes it an adequate microorganism for cell immobilisation and useful to increase iron(III) productivity. Different immobilisation methods and several supports have been employed, including the use of glass beads, activated carbon particles, sand, polystyrene, polyurethane, PVC and diatomaceous earth [3-8]. A. ferrooxidans cells immobilised on support can grow at higher dilution rates reaching high iron(III) productivity. Reactors with attached A. ferrooxidans producing iron(III) continually have been employed to enhance metal recovery from ores or in other applications [9-11].
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In most cases, under the conditions in which the immobilisation process was performed, abundant jarosite (basic iron(III) sulphate) precipitation was produced. Although this precipitation is described an unwanted phenomenon, jarosite deposits participate in the process of biofilm formation [12-13]. Support matrices with high porosity can be used to enhance jarosite precipitation and biofilm formation. In this report, we present the immobilisation of A. ferrooxidans cells in refractory clay tiles (composed by kaolin) highly porous. Iron(II) oxidation rates during the biofilm formation and the oxidative efficiency of this biofilm during storage at 4°C were evaluated. 2.
MATERIALS AND METHODS
2.1 Microorganism and cultivation A strain of A. ferrooxidans from Santa Rosa de Arequipa (DSM11477) was used through out these experiments. The microorganism was grown and maintained on 9 K medium at initial pH 1.80 [14]. 2.2 Support matrices The supports used in this study were a ceramic fibre, whose commercial name is caowool (S1), and two refractory clay tiles (S2 and S3). The composition of the last supports is the following: S2: SiO2: 27.7%, Al2O3: 50.8%, Fe2O3: 0.38%, TiO2: 0.25%, CaO: 17.1%, MgO: 0.14%, Na2O: 1.94% and K2O: 1.34%. S3: SiO2: 54.7%, Al2O3: 42.8%, Fe2O3: 0.7%, TiO2: 1.7% and CaO: 0.6%. 2.3 Attachment to the supports 20 mg of each support were added to 5.0 ml of bacterial suspension previously filtered to remove jarosite. The test tubes were incubated at 30ºC and 180 rpm for 10 min. The mixtures were then filtered through black ribbon filter paper and the number of the cells remaining in suspension was determined by direct microscopic counting. Cell removal by the glass walls of the test tubes was determined in control tests in the absence of the supports. That value was taken into account when calculating the percentage of attachment to solids. 2.4 Bacterial growth in the presence of the support Cubic sections of each support (3 cm x 4 cm x 3 cm) corresponding to 7.5 g (S1) or 13.4 g (S2 or S3) were added to 250 ml Erlenmeyer flasks containing 100 ml of 9 K medium inoculated with a A. ferrooxidans inoculum. The flasks were incubated at 30°C and 180 rpm. Sterile controls were done with the inocula replaced by sterile medium. 2.5 Process of biofilm formation The reactor was a glass column (30 cm x 5 cm). A piece of the support (18 cm x 4 cm x 2 cm) was suspended into the column. 250 ml of 9 K medium (pH 1.80) previously inoculated with A. ferrooxidans were percolated through the column until the complete oxidation of iron. An air current of 120 liter.h-1 supplied the gaseous nutrients. When iron(II) was completely oxidised, medium was replaced by fresh medium without any intermediate inoculation. This procedure of consecutive batches allows reaching the 596
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formation of a biofilm. In continuous cultures the flow rates for fresh media were regulated with peristaltic pumps. 2.6 Support storage After the process of biofilm formation, pieces of the support with immobilised biomass were stored in the open air or in iron-lacking medium for 24 months. After 2, 12 and 24 weeks, samples of the support were placed in 250 ml Erlenmeyer flasks with 100 ml of sterile 9 K medium and maintained at 30°C in an orbital shaker at 180 rpm. 2.7 Analytical procedures Iron(II) concentration was measured by titration with potassium permanganate. Total soluble iron was determined by atomic absorption spectrophotometry. Bacterial population in suspension was monitored using a Petroff-Hausser camera in a microscope with a phase contrast attachment. Solid residues obtained during the experiments were analysed through X-ray diffraction and Mossbauer spectroscopy. 3.
RESULTS AND DISCUSSION The percentages of attachment of A. ferrooxidans to the supports were 1.8%, 23.2% and 38.2% for S1, S2 and S3 respectively. This adhesion is dependent of the physicalchemical properties of the bacteria and the solid surfaces and although it is responsible for the first and reversible phase of biofilm formation, it is not enough to decide if the support is ideal to generate an adequate process for biofilm formation. That is why, after the former phase, an irreversible phase should occur. This last phenomenon is associated strongly to the jarosite precipitation [13] and could also be in relationship with the initial bacterial adherence. A high bacterial attachment to the support would produce a fast iron(II) oxidation very close to the surface increasing pH and iron(III) concentration and the subsequent iron(III) precipitation. These deposits on the surface would allow the later adsorption of cells on the pores and the formation of the biofilm. However, although the initial attachment could be important there are other factors to be taken into account. Thus for example, some supports could increase the adhesion of jarosite on their surface independently to the presence of attached bacteria. Once precipitation occurred cells could adsorb to the surface of highly porous jarosite. On the other hand, the supports could inhibit the bacterial activity. For these reasons, an experiment of bacterial growth in the presence of each support was carried out. Figure 1 shows the results of the bacterial growth in the presence of each support. Although the supports did not affect the growth significantly, in the presence of the support 2 bacterial growth was slower than in the cultures with the other supports. Moreover, in this case final pH value was the highest, which is in agreement with the more basic characteristics of this support. Since the formation of iron(III) precipitates is highly dependent on pH, in this case the amount of iron(III) precipitation was abundant and consequently a little amount of total iron remains soluble at the end of the bacterial growth. Due to this behaviour, support 2 was discarded for the following experiments. In order to analyse the iron(III) deposits produced during the growth, the supports were kept in contact with the spent medium for 12 hours. That procedure allowed a higher iron(III) precipitation: 55%, 97% and 78% for the cultures with supports 1, 2 and 3 respectively. These solids were characterised as potassium and ammoniumjarosite by Xray diffraction. However, analysis through Mossbauer spectroscopy proved the existence of one or more phases of ferric oxohydroxides (whose exact composition was not 597
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determined) joined to jarosite. In the solid residues, the percentage of jarosite was 41% (for S1) and 21% (for S2 and S3). Figure 2 shows the results for the consecutive batches run using the immobilised biomass on supports 1 and 3 and a rate between medium volume and mass support (V/m) of 5 ml/g.
Figure 1. Oxidation of iron(II) by A. ferrooxidans cells in the presence of different supports. White bars: oxidation rates. Grey bars: final pH values. Black bars: final bacterial population in suspension
Figure 2. Iron(III) productivity in bioreactors with immobilised biomass on supports 1 (white bars) and 3 (black bars) at 5 ml/g (volume/support mass). The values correspond to first, third and sixth cycle for support 1 and first, fifth and eleventh for support 3 598
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In the reactor with support 1, productivity values show the trend to decrease. This behaviour indicates that there is only a small-immobilised biomass, which is even further removed by successive washing after each cycle. For this reason, this support was discarded after six cycles. On the other hand, when support 3 was employed a constant increase in iron(III) productivity until a maximum value (after 6-7 cycles) was observed. The constant value is reached probably because the new iron(III) deposits avoid the nutrient access. However, the iron(III) productivity and the final soluble iron were much lower than those using other supports [7,10-11]. When the reactor, operating in repeated batches mode, reached its highest iron(III) productivity, culture medium was replaced by fresh medium and the continuous flow of medium through a peristaltic pump was started. Figure 3 shows iron(III) productivity as a function of the dilution rate. The productivity values are higher than those obtained in the batch mode. It was also observed that the higher the dilution rate, the lower the iron(III) precipitation.
Figure 3. Iron(III) productivity in the bioreactor with immobilised biomass on support 3 at different dilution rates
In order to test that, a new experiment in "batch mode" was carried out modifying the rate V/m to 10 ml/g. Figure 4 shows that iron(III) productivity was significantly enhanced by increasing V/m. As it was expected, iron(III) precipitation was lower allowing higher iron(III) concentrations at the end of each cycle. This behaviour did not change even if the medium was maintained in contact with the support for more than 24 hours after iron(II) was exhausting. One factor to be taken in account in the industrial application of a reactor with immobilised biomass is its operational stability after stopping operation for long periods. The reactor should be stopped temporarily and re-started whenever required without further manipulation. This characteristic should be proved not only when the support is maintained wet with nutrient-lacking medium but also when the medium was totally drained from the column.
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Figure 4. Iron(III) productivity in the bioreactor with immobilised biomass on support 3 at 10 ml/g (volume/support mass). White bars: iron(III) productivity. Black bars: final bacterial population in the suspension. The values correspond to first, third and sixth cycle.
In our case, after 2 and 12 weeks of storage the support with immobilised biomass was placed in a medium with iron. The results of iron(II) oxidation are shown in Figure 5. It can be observed that the cells immobilised in the wet support were reactivated very rapidly after storage of 2 or 12 weeks. When the support was stored for 24 weeks (data not shown), cells were reactivated after a long lag phase (7-10 days). However the support stored in open air only shows adequate activity after 2 weeks.
Figure 5. Iron(II) oxidation by A. ferrooxidans cells immobilised on support 3 before storage (first graph) and after storage for 2 and 12 weeks (second and third graphs respectively) in open air (F) or in medium without iron (J) at 4°C 600
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4.
CONCLUSION Summarising, some refractory kaolin tiles can be employed as adequate support for the immobilisation of A. ferrooxidans cells generating a good biofilm. This biofilm achieved high iron(III) productivity in both batch and continuous modes of operation with lower iron(III) precipitation at higher dilution rates. The immobilised biomass on the support clearly retained their oxidative efficiency during storage at 4ºC under dried or wet conditions for 2 or 24 weeks respectively. ACKNOWLEDGEMENTS This work was funded by the Agencia Nacional de Promocion Cientifica y Tecnologica (PICT 99). REFERENCES
1. J. Barrett, M.N. Hughes, G.I. Karavaiko and P.A. Spencer (eds.), Metal Extraction by Bacterial Oxidation of Minerals, Ellis Horwood limited, Chichester, 1993. 2. D.E. Rawlings (ed.), Biomining: Theory, Microbes and Industrial Processes, SpringerVerlag, Berlin, 1997. 3. S. Grishin and O. Tuovinen, Appl. Environ. Microbiol. 54 (1988) 3092. 4. M.J. Garcia, I. Palencia and F. Carranza, Proc. Biochem. 24 (1989) 84. 5. H. Armentia and C. Webb, Appl. Microbiol. Biotechnol. 36 (1992) 697. 6. N. Wakao, K. Endo, K. Mino, Y. Sakurai and H. Shiota, J. Gen. Microbiol. 40 (1994) 349. 7. C. Webb and G. Dervakos (eds.), Studies in viable cell immobilization, Academic Press, London, 1996. 8. M.E.A.G. Oprime and O. Garcia Jr. In: V.S.T. Ciminelli and O. Garcia Jr (eds.), Biohydrometallurgy: Fundamentals, Technology and Sustainable Development, Part A, Elsevier, Amsterdam, 2001, pp. 369-375. 9. S. Porro, C. Pogliani, E. Donati and P.H. Tedesco, Biotechnol. Lett. 15 (1993) 207. 10. E. Donati, L. Lavalle, V. de la Fuente, P. Chiacchiarini, A. Giaveno and P. Tedesco. In: T. Vargas, C.A. Jerez, J.V. Wiertz and H. Toledo (eds.), Biohydrometallurgical Processing, Vol.1, Universtiy of Chile, Santiago, 1995, pp. 293-300. 11. G. Curutchet, E. Donati, C. Oliver, C. Pogliani and M.R. Viera. In: R.J. Doyle (ed.), Microbial Growth in Biofilms, Part B: Special Environments and physicochemical aspects, Methods in Enzymology, Vol. 337, Academic Press, San Diego, 2001, pp. 171-186. 12. D.G. Karamanev, J. Biotechnol., 20 (1991) 51. 13. C. Pogliani and E. Donati, Process Biochem. 35 (2000) 997. 14. M.P. Silverman and D.G. Lundgren, J. Bacteriol. 77 (1959) 642.
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Investigation of bioremediation techniques for cleaning-up arsenic contaminated soils K. Vaxevanidou, N. Papassiopi and I. Paspaliaris School of Mining and Metallurgical Engineering, National Technical University of Athens, GR 157 80 Zografos, Greece Abstract Several soil remediation technologies, which were developed for cleaning up heavy metal contaminated soils were found to have limited effectiveness on metalloid As, mainly due to its strong association with the Fe(III)-oxides of soil. The aim of this study was to investigate whether it is possible to enhance the mobilization of As through the biological reduction of Fe(III)-oxides. Two representative Fe(III) reducing microorganisms, i.e. Shewanella putrefaciens and Desulfuromonas palmitatis, were used in the experimental work. The effectiveness of this treatment was initially evaluated using pure compounds representing the association of As with soil Fe(III), i.e. a ferric arsenate (scorodite, FeAsO4.2H2O) and a ferric oxyhydroxide (FeOOH) with adsorbed As, and was then applied on a soil sample contaminated with As due to past mining activities. The results of the study indicate that a combined biological and chemical treatment, exploiting the Fe(III) reducing ability of D. palmitatis bacteria and the chelating strength of EDTA, could provide a promising alternative for the remediation of As contaminated soils. Keywords: Soil remediation, arsenic, Fe(III) reducing bacteria, EDTA 1.
INTRODUCTION Contamination of soils with As consists a major environmental problem and a great research effort has been devoted recently for the development of appropriate remediation technologies. The remedial methods which are currently under development for the remediation of As contaminated soils can be broadly classified in two main categories. The first category of technologies are based on the in-situ immobilization principle and aim at reducing the mobility of arsenic by treating the soil with appropriate chemical reagents, which can transform the mobile As species to low solubility compounds. Lime, cement, ferrous and ferric compounds, etc. have been evaluated as stabilization agents at laboratory and field scale (1,2,3). The main advantage of stabilization techniques is their low cost and their ease of applicability, but there is always concern about the long-term stability of pollutants, since changing environmental conditions may cause their remobilization. The second category of technologies is oriented towards the removal of arsenic from the soil matrix in order to eliminate permanently the environmental risks. Chemical leaching processes (4,5), electrokinetic techniques (6) and recently phytoremediation methods (7) are some of the alternatives, which have been investigated for cleaning-up arsenic contaminated soils. The application of bioremediation techniques, 603
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though reported in some review papers as potential option (8), has not been thoroughly tested up to now. A biological process, based on the use of Fe(III)-reducing microorganisms, was investigated in this study for obtaining the removal of arsenic from contaminated soils. It is well recognized that the iron oxide and hydroxide mineral phases play a key role in the retention of As in soils. Both As(V) and As(III) species are strongly adsorbed by the amorphous or crystalline ferric oxides of soil (9,10). As a consequence the reductive dissolution of Fe(III) oxides is a possible mechanism for obtaining the release of bound As. This is supported by recent studies, which provided evidence for the mobilization of arsenic under the action of Fe(III) reducing bacteria (11). The experimental work was carried out using two representative Fe(III)-reducing microorganisms, i.e. Shewanella putrefaciens and Desulfuromonas palmitatis. S. putrefaciens, is a facultative anaerobe, which was shown to be a dissimilatory Fe(III) reducer under anaerobic conditions (8). This microorganism can grow both in fresh and seawater. S. putrefaciens can reduce Fe(III) using H2, formate or lactate as electron donors, but lactate is incompletely oxidized to acetate. D. palmitatis is a strict anaerobic, marine microorganism, which can grow using Fe(III) as electron acceptor and a wide variety of short and long chain fatty acids as electron donors. In the case of D. palmitatis, all organic compounds serving as electron donors are completely oxidized to CO2 (12). The methodology that was followed to evaluate the effectiveness of this biological treatment comprised three main steps: •
The initial experiments, aiming at comparing the efficiency of the two bacteria species, were conducted using a ferric arsenate compound, i.e. FeAsO4.2H2O (scorodite), as model solid representing the association of As with Fe(III). D. palmitatis was found to be much more effective than S. putrefaciens in reducing the trivalent iron of crystalline scorodite and was selected for the subsequent experimental work. The initial tests with scorodite revealed also an undesirable side effect. Despite the high Fe(III)-reducing efficiency of D. palmitatis, arsenic was not released in solution due to the formation of low solubility Fe(II)-As(V) compounds.
•
In a second series of tests the effectiveness of D. palmitatis was evaluated using a ferric oxyhydroxide compound, FeOOH, containing or not adsorbed As. In this series of tests the biological treatment was combined with the use of EDTA (ethylenediamine tetracetic acid), a well known chelating reagent, which was added in the aqueous solution in order to keep Fe(II) in solution and avoid the precipitation of ferrous arsenates.
•
The combined biological and chemical treatment was finally tested on a soil sample contaminated with As due to past mining activities.
2.
MATERIALS AND METHODS
2.1 Microorganisms and media Shewanella putrefaciens (DSM 6067) and Desulfuromonas palmitatis (DSM 12391) were purchased from the German Collection of Microorganisms (DSM Branschweig, Germany). The basic medium in the case of S. putrefaciens contained (per litre): 2.5g of NaHCO3, 0.1 g of KCl, 1.5 g of NH4Cl, 0.6 g of NaH2PO4xH2O, 0.02 g of L-agrinine hydrochloride, 0.02 g of L-glutamine, 0.04 g of DL-serine, 10 ml of vitamin solution and 604
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10 ml of trace element solution (13). The initial growth was carried out aerobically using sodium lactate (20mM) as electron donor. Experiments with scorodite were conducted with 50mM of scorodite as electron acceptor and 20mM of lactate as electron donor. The basic medium of D. palmitatis contained the following constituents per litre: 20.8 g of NaCl, 2.5 of g NaHCO3, 0.77 g of KCl, 1.0 g of NH4Cl, 0.1 g of NaH2PO4⋅H2O, 0.2 g MgSO4⋅7H2O, 1mL of selenite-tungstate solution, 10 mL of vitamin solution, 10 mL of trace elements solution, 10.6 g of MgCl2⋅6H2O and 1.52 g of CaCl2⋅2H2O (DSM Medium 837). The initial growth was carried out using sodium acetate (10mM) as electron donor and sodium fumarate (50mM) as electron acceptor. Tests with scorodite and FeOOH were conducted using the same electron donor, i.e. 10mM acetate, and 50mM of scorodite or FeOOH as electron acceptors. 2.2 Solid materials The scorodite, FeAsO4.2H2O was prepared in the laboratory by combining a solution of 0.3M Fe(NO3)3 with 25 g/L As(V) (as As2O5) and heating to 150°C for 1 hour in an autoclave. The solids assayed 23.8% Fe and 33.8% As and analysis by X Ray Diffraction (XRD) identified only scorodite. The ferric oxyhydroxide, FeOOH, was purchased from Sigma-Aldrich (PN 546267) and assayed 62.1% Fe. X Ray Diffraction analysis indicated that the solid is amorphous. The soil material, used in the tests, originates from location Kyprianos at the mining area of Lavrion, Greece. The main characteristics of soil material are summarized in Table 1. As seen in the Table, the soil is alkaline, pH~8.5, and contains a high percentage of calcite, i.e. CaCO3 =39.5%. It is highly contaminated with Pb and Zn, i.e. up to 24000 mg/kg, and to a lesser extent with As, 950 mg/kg, and Cd, 360 mg/kg. Mineralogical analyses indicated that the main minerals are quartz, calcite, fluorite, muscovite and clinochlore. Contaminants are mainly found in the form of carbonates, e.g. PbCO3, Pb3(CO3)2OH2, ZnCO3 etc, and arsenates, e.g. Pb5(AsO4)3Cl, Zn2AsO4OH and FeAsO4.2H2O or associated with the Fe(III) oxy-hydroxides, mainly limonite, containing up to 2-5% PbO, 2-6% ZnO and 1-2% As2O5. Table 1. The main characteristics of soil material pH, alkalinity Soil pH CaCO3 equ. % Other Properties LOI % Insoluble %
8.65 39.5 17.2 15.8
Main metals Ca Mg Al Fe Mn
% 34.1 0.81 0.73 2.50 0.20
Contaminants As Pb Zn Cd
mg/kg 946 19640 23920 357
LOI: Lost of ignition at 900°C, Insoluble: following digestion in aqua regia (HNO3/HCl 1:3).
2.3 Experiments with scorodite All the experiments were carried out using 120 mL serum bottles and applying standard anaerobic techniques. For the preparation of experimental slurries, 0.924 g of scorodite were transferred in the serum bottles. The aqueous phase constituents were added from concentrated stock solutions and the volume was adjusted with deionised water to 72 and 80 ml for the biological and control tests respectively. The slurries were bubbled for approximately 20min with N2:CO2 (80:20) to remove dissolved oxygen and then sealed. The slurries corresponding to the biological tests were inoculated with 8 mL from the D. palmitatis or S. putrefacians cultures, which were grown at an optical density 605
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of 1.0 cm-1. The serum bottles were incubated horizontally on a shaking waterbath at 30°C. Two kinds of samples were analyzed during the tests with scorodite, i.e. aqueous phase and slurry samples. The procedure for taking the aqueous phase samples was the following: an aliquot of 1mL was withdrawn from the serum bottle using a syringe and the sample was filtered through a 0.2 µm RC-membrane filter. The filtrate was weighed to take account for the retention of solution in the filter (~0.20-0.25mL), diluted to 10mL with deionised water and analyzed for total As(T), As(III), total Fe(T) and Fe(II). A second aliquot of 1 mL was taken following the rigorous agitation of the pulp, in order to maintain the solids in suspension. The aliquot of slurry was transferred to a test tube containing 9mL of 6M HCl. In this highly acidic solution the solids were dissolved within ½ hour and the extract was analyzed for As(T), As(III), Fe(T) and Fe(II). 2.4 Experiments with ferric oxyhydroxide, FeOOH The bacteria species used in these experiments were D. palmitatis due to their greater efficiency in reducing the trivalent iron of scorodite. Growth medium was the same described above, with acetate serving as electron donor and FeOOH as electron acceptor. Tests were carried out by mixing the aqueous solutions with 50mM of FeOOH and 50mM of CaCO3, which was added in order to simulate the calcareous matrix of Lavrion soil. Contamination with As was simulated by adding 1.5 mM of As (as Na2HAsO4) in the aqueous solution. This As/Fe ratio corresponds approximately to the As and Fe proportion in Lavrion soils, i.e. 28 µmole As/mmole Fe. During these tests, the biological activity of D. palmitatis was combined with the chelating strength of EDTA, in order to hinder the precipitation of insoluble Fe(II)arsenates through formation of soluble Fe(II)-aqueous complexes. EDTA was added in the aqueous solution in the form of its calcium salt, i.e. Na2CaEDTA. Previous experiments carried out in similar calcareous soils, indicated that soil Ca binds a considerable part of EDTA and competes with the other divalent metals for complexation with EDTA (14,15). This effect was taken into consideration and Na2CaEDTA was used instead of the most common Na2H2EDTA salt. The biogenic Fe(II) was expected to replace Ca in the corresponding EDTA chelate due to its higher complexation constant, i.e. logK=16.2 for Fe(II) versus 12.1 for Ca. To reduce the competitive effect of Ca, tests were also carried out with the addition of both Na2CaEDTA and NaHCO3 in the aqueous solution. Thermodynamic calculations indicated that an excess of carbonate ions induces the removal of Ca from the aqueous solution as solid CaCO3 and thus promotes the complexation of Fe(II) by EDTA. The experimental conditions applied during the treatment of FeOOH are summarized in Table 2. Six experiments (1C, 2C, 3C, 1B, 2B, 3B) were conducted in the presence of As and three (1B’, 2B’, 3B’) without As, to evaluate eventual inhibition of the biological activity due to the presence of arsenic. Tests 1C, 2C and 3C were carried out under abiotic conditions (control tests) to estimate the contribution of purely chemical mechanisms. The aqueous solution in tests 1C, 1B and 1B’ contained only the components required for D. palmitatis growth, while in tests 2C, 2B and 2B’ 100mM of Na2CaEDTA was also added. Tests 3C, 3B and 3B’ were carried out by adding both Na2CaEDTA (100mM) and NaHCO3 (100mM) in the aqueous solution. After the preparation and sterilization of slurries, serum bottles were incubated under agitation at 30°C. Inoculation of the biological tests was carried out after 7 days of incubation. This period of time was estimated to be sufficient for the adsorption of As on 606
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FeOOH. After the completion of the tests, the solid residues were subjected to acidic treatment with 6M HCl and the extract was analyzed for Fe(II), in order to evaluate whether the biological treatment resulted in the formation of insoluble Fe(II) compounds, as well as for total Fe, As and Ca. The solids were also examined with XRD. Table 2. Summary of experimental conditions with FeOOH
FeOOH CaCO3 Medium Na2HAsO4 Na2CaEDTA NaHCO3 Bacteria
50 mM 50 mM 1.5 mM 100 mM 100 mM 10% v/v
1C + + + +
Tests with adsorbed As 2C 3C 1B 2B + + + + + + + + + + + + + + + + + + + + + +
3B + + + + + + +
Tests without As 1B’ 2B’ 3B’ + + + + + + + + + + +
+
+ + +
2.5 Experiments with soil An experimental protocol, similar to that applied for FeOOH, was also used for soil. The soil was subjected to three kinds of treatment: • Simple biological treatment with D. palmitatis • Biological treatment combined with chemical assistance by Na2CaEDTA • Biological treatment in the presence of both Na2CaEDTA and NaHCO3 All tests were carried out by mixing 4 g of soil with 80 ml of solution. After the completion of the tests, solid residues were subjected to the 6M HCl treatment and the extract was analyzed for all metal contaminants, as well as for total Fe and Fe(II). 2.6 Analytical techniques Total elements concentrations were analyzed with flame atomic adsorption spectrophotometry (AAS). Hydride generation coupled with AAS was used for samples containing low concentrations of As. Iron (II) was determined by the phenanthroline technique and As(III) by the molybdenum method as modified by Cummings et al. (11). Due to the strong interference of EDTA with the above colorimetric methods, speciation of Fe and As was carried out only for the samples which did not contain EDTA. 3.
RESULTS AND DISCUSSION
3.1 Scorodite experiments Representative results from the biological treatment of scorodite with S. puterfaciens and D. palmitatis are shown in figure 1. The release of As in the aqueous solutions is presented in figure 1a. The trend of arsenic mobilization was quite different between the two bacteria species. Treatment with S. puterfaciens resulted in the continuous release of As in solution, up to the final percentage of 7%, whereas treatment with D. palmitatis led to an initial rapid increase of aqueous As, followed by the gradual reprecipitation of the element. During the treatment of scorodite with D. palmitatis, a drastic change of the solid phase was observed. White scorodite was transformed into a dark green solid phase, between the 11th and 16th day of treatment. This effect was not observed with S. putrefaciens. To investigate solid phase transformations, slurry samples were 607
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systematically analyzed for total Fe and As, as well as for Fe(II) and As(III). Arsenic (III) was below detection limits in all samples, indicating that neither S. putrefaciens nor D. palmitatis were able to alter the oxidation state of As. Concentrations of Fe(II) in the slurry samples are shown in figure 1b. It is seen that the trivalent iron of scorodite was reduced to a percentage of 80% within 16 days, using the D. palmitatis bacteria. S. putrefaciens were found to be less efficient, with only 10% reduction achieved after 65 days of treatment.
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a. Release of As in the aqueous phase
b. Total reduced Fe(II) in both aqueous and solid phases
c. XRD pattern of S. putrefaciens bioresidue
d. XRD pattern of D. palmitatis bioresidue
Figure 1. Treatment of scorodite with S. puterfaciens and D. palmitatis.
Evaluating the chemical analyses of aqueous and slurry samples in the case of D. palmitatis, it was deduced that the green solid phase consists essentially of ferrous arsenates. The XRD analysis of D. palmitatis product (Figure 1d) confirmed the presence of symplesite, Fe3(AsO4)2.8H2O, in this solid. Symplesite however appears to be a secondary phase. The most intense peaks at d=2.345nm and d=2.031nm remained unidentified and may correspond to other ferrous arsenate compounds, which were not included in the available XRD database. In the S. putrefaciens residue only scorodite was detected (Fig. 1c) 3.2 Experiments with ferric oxyhydroxide, FeOOH The main conclusion from the initial series of tests was that D. palmitatis is much more efficient than S. putrefaciens in reducing the trivalent iron of crystalline scorodite and for this reason it was D. palmitatis that was selected for the subsequent experimental work with FeOOH. In this case, the biological treatment was combined with the chelating strength of EDTA. Results from the combined biological and chemical treatment of FeOOH are shown in Figure 2. The dissolution of iron during the abiotic treatment with the three solutions is presented in Figure 2a. The concentration of aqueous Fe in the control test that was carried out using only the culture medium was constantly below 608
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detection limit. The addition of Na2CaEDTA resulted in a slight mobilization of Fe with 6% dissolution after 50 days of treatment. Using both Na2CaEDTA and NaHCO3, the dissolution increased to 10%. The adsorption of As on FeOOH in the three solutions is presented in figure 2b. In the solution containing only the culture medium, almost all As, i.e. 1.5mM, was adsorbed during the preparation of the slurry and the residual concentration in solution was close to 0.02mM. The addition of Na2CaEDTA and Na2CaEDTA with NaHCO3 retarded slightly the adsorption process and resulted in higher residual concentrations, i.e. 0.08 and 0.17 mM respectively. As seen in Figures 2c and 2d, inoculation of D. palmitatis on the 7th day of treatment had no obvious effect compared to the control experiments. The trend of iron dissolution and that of As adsorption remained essentially similar to that of the corresponding control experiments. A slight mobilization of Fe followed by reprecipitation was only observed during the simple biological treatment. Results from the treatment of FeOOH without adsorbed arsenic are shown in Figure 2e. As seen in the figure, combined treatment with D. palmitatis, Na2CaEDTA and NaHCO3 resulted in the dissolution of FeOOH up to almost 85%. The negative redox potential of the aqueous phase, i.e. -256 mV compared to +89 mV in the corresponding control experiment (Ag/AgCl electrode), indicates that almost all dissolved Fe is in the divalent state. The experimental results, i.e. decrease of redox potential, pH rise from 7.5 to 8.67, decrease of Ca concentration from 100 to 54 mM etc., suggest that dissolution took place according to reactions (1) to (3). The 3 reactions describe the biological reduction of FeOOH by D. palmitatis (reaction 1), the complexation of divalent Fe by EDTA with the simultaneous release of Ca2+ (reaction 2) and the precipitation of Ca2+ as CaCO3 (reaction 3): 2+ D. palmitatis 8 FeOOH + CH3COO- + 15 H+ ⎯⎯ ⎯⎯⎯ ⎯→ 8 Fe + 2 HCO3 +12 H2O (1) 8 Fe2+ + 8 CaEDTA2- ⎯→ 8 FeEDTA2- + 8 Ca2+ 2+
-
+
8 Ca + 8 HCO3 ⎯→ 8 CaCO3 ↓ + 8 H
(2) (3)
During the chemical analyses of solid residues, ferrous iron was detected only in the two bioresidues, 1B and 1B’, of the simple biological treatment. The percentage of reduced iron was very low, i.e. less than 2.5%, in bioresidue 1B containing presorbed arsenic, but it was considerable, i.e. almost 31%, in the bioresidue 1B’ of pure FeOOH. The ferrous precipitate in bioresidue 1B’ was identified by XRD as siderite, FeCO3. It is worthwhile to mention that the other bioresidues produced during the treatment with Na2CaEDTA or Na2CaEDTA plus NaHCO3 were not found to contain any ferrous precipitates, when examined by chemical and XRD analyses. In this second series of tests, FeOOH was found to be more recalcitrant to the biological treatment than scorodite. Approximately 31% of iron was reduced during the simple biological treatment of FeOOH against 80% for scorodite. The presence of Na2CaEDTA had no effect on the bioreduction process. On the contrary, the addition of both Na2CaEDTA and NaHCO3 enhanced the biological activity and resulted in the reductive dissolution of FeOOH up to almost 85%. Another important finding was that presorbed As inhibits FeOOH reduction, even in the presence of Na2CaEDTA and NaHCO3, binding probably the majority of active sites on the FeOOH surface. This is in accordance with previous studies which have shown that bioreduction is suspended when the surface of Fe(III)-oxides is covered by sorbed species and is not any more accessible to the microorganisms (16). 609
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d. As, biological treatment with As 50
100 80 Fe dissolution, %
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0
0
0.30
EDTA+ CO3
D.p.
60
40
EDTA+ CO3
30 20
40 20
10
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Medium
0 0
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EDTA Medium
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5 As in solution, %
10 Fe dissolution, %
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0
60
e. Fe, biological treatment without As Figure 2. Combined biological and chemical treatment of the iron oxide, FeOOH, using D. palmitatis with and without adsorbed As (arrows indicate the day of inoculation) 3.3 Experiments with Lavrion soil The dissolution of iron during the combined biological and chemical treatment of soil is shown in Figure 3a. No Fe was detected in the aqueous phase during the simple biological treatment. On the contrary, a constant increase of Fe concentration was observed when the biological treatment was combined with the presence of Na2CaEDTA and the extraction was more than 60% after 122 days of treatment. The addition of NaHCO3 improved slightly the kinetics of iron dissolution. The release of As is shown in Figure 3b. The simple biological treatment caused only a slight mobilization of As in the order of 3%. In the presence of Na2CaEDTA, the removal of As was more than 70% and the addition of NaHCO3 improved further the extraction up to 90%. 610
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Comparing the results obtained with FeOOH and soil, it is evident that the Fe(III)oxides of soil are less refractory to the biological treatment. For instance, Na2CaEDTA, which was found inefficient in the case of FeOOH, was seen to be very effective during the treatment of soil. Moreover, the presence of As in the contaminated soil sample has not inhibited the biological activity of D. palmitatis, as previously observed during the experiments with FeOOH and adsorbed As. Obviously, the association of As with the Fe(III) oxides in contaminated soils is not a simple surface adsorption phenomenon. Coprecipitation with excess Fe(III) may represent more accurately the association, at least for this particular soil sample. In the case of coprecipitation, As could be enclosed in deeper layers, leaving the outer surface of oxides free and accessible to the bacterial action. Finally, it is worthwhile to mention that this combined biological and chemical treatment resulted in the simultaneous extraction of Pb and Zn up to 89 and 72% respectively (data not shown).
5
20 Medium 0
0 0
50
100 t, days
a. Iron
150
0.4
60
EDTA
40
0.2
20
Medium
As in solution, mM
10
EDTA
As, % extraction
40
0.6
EDTA+ CO3
80
Fe in solution, mM
Fe, % dissolution
60
100
15
EDTA+ CO3
0.0
0 0
50
100
150
t, days
b. Arsenic
Figure 3. Dissolution of Fe and As during the combined biological and chemical treatment of soil 4.
CONCLUSIONS The first series of tests, which was conducted using scorodite as model compound, indicated that D. palmitatis is much more efficient than S. putrefaciens in reducing the trivalent iron of crystalline scorodite. In contrast to scorodite, FeOOH was found to be rather recalcitrant to the biological treatment with D. palmitatis, particularly in the presence of adsorbed arsenic. Bioreduction of FeOOH occurred only when the activity of D. palmitatis was combined with the addition of Na2CaEDTA and NaHCO3 in the aqueous solution. Prior adsorption of As on the surface of goethite inhibited completely the bioreduction process and it was not possible to mobilize sorbed As, even in the presence of Na2CaEDTA and NaHCO3. However, the same treatment scheme was found to be very effective in mobilizing As from the contaminated soil samples. More than 60% of Fe(III) was reduced and almost 90% of As was released in the aqueous solution during experiments with soil, indicating that this combined biological-chemical treatment could provide a promising alternative for the remediation of As contaminated soils. ACKNOWLEDGMENTS This work was carried out in the framework of METALBIOREDUCTION project, with the financial support of the European Commission (Contract No. EVK1-CT-199900033). 611
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REFERENCES
1. Moore, T.J., C.M. Rightmire, and R.K.Vempati (2000), Soil and Sediment contamination, Vol. 9(4), pp. 375-405. 2. US EPA (2002), EPA-542-R-02-004, http:/clu-in.org. 3. Voigt, D.E. and Brantley, S.L. (1996), Applied Geochemistry, Vol.11, pp. 633-643. 4. Legiec I.A., Griffin L.P. et al. (1997), Environmental Progress, Vol.16 (1), pp. 29-34. 5. Alam, M., Tokunaga, S. and Maekawa, T. (2001), Chemosphere, Vol. 43, pp.10351041. 6. EPRI, Electric Power Research Institute (2000), Report 1000203, http://www.epri.com 7. Francesconi, K., Visoottiviseth, P., Sridokchan, W. and Goessler, W. (2002), , The Science of The Total Environment, Vol. 284(1-3), pp. 27-35 8. Lovely, D.R. and Coates, J.D. (1997), Curr. Opinion Biotechnol., Vol. 8, pp. 285-289. 9. Fendorf, S., Eick, M.J., Grossl, P., Sparks, D.L. (1997), Environ. Sci. Technol., Vol. 31, pp. 315-320. 10. Raven, K. P., Jain, A. and Loeppert, R.H. (1998), Environ. Sci. Technol., Vol. 32, pp. 344-349. 11. Cumming, D.E., Caccavo, F., Fendorf, S. and Rosenzweig, R.F. (1999), Environ. Sci. Technol., Vol. 33(5), pp. 723-729. 12. Lovley, D.R., J.D. Coates, D. Saffarini, and D.J. Lonergan (1997), (Winkelman, G., and C.J. Carrano, eds.), Harwood Academic, Switzerland, pp.187-215 13. Lovley, D.R., Phillips, E.J.P. and Lonergan, D.J. (1989), Applied Environmental. Microbiology, Vol. 55, pp. 700-706 14. Papassiopi N., S. Tambouris and A. Kontopoulos (1999), Water, Air and Soil Pollution, Vol. 109, pp. 1-15. 15. Theodoratos P., N. Papassiopi, T. Georgoudis and A. Kontopoulos (2000), Water, Air and Soil Pollution, Vol. 122, pp. 351-368. 16. Urrutia, M.M., Roden, E.E., Zachara, J.M. (1999), Environmental Science and Technology, Vol. 33 pp. 4022-4028.
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Leaching characteristics of heavy metals from sewage sludge by Acidithiobacillus thiooxidans MET K.S. Choa, H.S. Moona, N.Y. Yooa and H.W. Ryub a
Department of Environmental Science and Engineering, Ewha Womans University, Seoul 120-750, Korea b Department of Chemical and Environmental Engineering, Soongsil University, Seoul 156-743, Korea
Abstract An acidophilic, sulfur-oxidizing Acidithiobacillus thiooxidans strain MET allowed metal leaching even at high sludge solids concentration such as 130 g.L-1. It was found that low metal leaching efficiency at high solids concentration was mainly due to an increase in buffering capacity resulting in retardation of pH reduction. Therefore, metal leaching was mainly influenced not by sludge solids concentration, but by the pH (or sulfate concentration per unit sludge mass) of the sludge solutions. The relationship between the pH of the sludge solution and the efficiency of metal leaching was obtained by quantitatively investigating the effect of pH reduction or the amount of sulfate produced per unit sludge mass on leaching of each metal. Keywords: bioleaching, heavy metals, sewage sludge, Acidithiobacillus thiooxidans 1.
INTRODUCTION The treatment and final disposal cost of sewage sludge represent 50% of the overall cost of the wastewater treatment process [1]. Most of the sludge produced is discharged in landfills or the ocean, but alternative methods such as incineration, solidification, composting, and pyrolysis have been recently examined [2]. One of the most economical sludge utilization methods is composting followed by land application. However, heavy metals present in the sludge often hinder agricultural land application of the composted sludge. Total heavy metal content in the sludge is generally about 0.5-2.0% of total dry weight, but in some cases it could reach up to 4% [1, 3, 4, 5]. Numerous methods have been proposed to remove heavy metals from sewage sludge, such as chlorination, use of chelating agent, and acid treatment at high temperatures. However, these methods are generally ineffective in practical applications due to their high cost, operational difficulties, and low metal leachability [6-10]. Microbial leaching with Acidithiobacillus spp. is used alternatively for heavy metal removal from sewage sludge [3-5, 11-15]. The most widely used microorganisms in metal leaching are Acidithiobacillus ferrooxidans and Acidithiobacillus thiooxidans [1, 8, 9, 10, 14, 16, 17]. To develop microbial metal leaching process of sludge, a steady supply of efficient microorganisms must be ensured. Higher process efficiency can be achieved by coinoculation of highly active isolated pure culture to the leaching. Moreover, its application 613
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in microbial leaching process can be optimized by understanding characteristics of the pure culture. Acidithiobacillus thiooxidans MET, which was active in a wide range of pH from acidic to neutral conditions was isolated by enrichment from the anaerobically digested sludge in the presence of elemental sulfur. In this study, the metal leaching efficiency of the bacterium at various sludge solids concentrations was evaluated. Furthermore, a relationship between the pH of the sludge solution and metal leaching efficiency was investigated. 2.
MATERIALS AND METHODS
2.1 Sludge Anaerobically digested, dewatered sewage sludge obtained from a sewage treatment plant located in suburb of Seoul, Korea, was used throughout the study. The sludge used was dried under sun to prevent its decomposition followed by grinding to powders (average particle size: 1 mm). Average concentrations of heavy metals in the sludge samples are as follows: 2,340±40 µg-Cu g-dry sludge-1, 31,870±461 µg-Al g-dry sludge-1, 1,152±31 µg-Cr g-dry sludge-1, 4,529±105 µg-Zn g-dry sludge-1, 829±19 µg-Ni g-dry sludge-1, 222±10 µg-Pb g-dry sludge-1 and 33,020±1,320 µg-Fe g-dry sludge-1. 2.2 Bacterium and medium For the culture of A. thiooxidans MET (KCTC 8928P), modified Waksman (MW) medium of the following composition was used [18]: 3.0 g.L-1 K2HPO4, 0.1 g.L-1 MgSO4.7H2O, 0.3 g L-1 CaCl2.2H2O, 0.01 g.L-1 FeSO4.7H2O, and 10 g.L-1 S° as an energy source. The initial pH of the MW medium was 4.0. 2.3 Effect of inoculation To investigate the effect of A. thiooxidans MET on metal leaching, four incubation systems were prepared in 250mL Erlenmeyer flasks containing 20 g.L-1 sewage sludge, 100 mL MW medium and 10 g.L-1 elemental sulfur as follows:
System I: sterilized for 20 min at 121°C and 1.5atm System II: nonsterilized System III: nonsterilized, inoculated with the mixed sulfur-oxidizing culture (10%, v/v) System IV: nonsterilized, inoculated with A. thiooxidans MET (10%, v/v) The pH of the incubation bath was measured every day. 2.4 Effect of sludge concentration A. thiooxidans MET was cultivated in 500 mL Erlenmeyer flask containing 200 mL MW medium and sewage sludge (50 g.L-1) at 30°C and 180rpm for three days. The resulting culture broth, which had been previously adapted in the sludge, was used as an inoculum to elucidate effect of various sludge concentrations on microbial leaching of heavy metals. This culture (10%, v/v) was inoculated in 500mL Erlenmeyer flasks containing the sewage sludge at various concentrations (20, 50, 90, and 130 g L-1 in dry weight) in 180 mL inorganic salt medium and 10 g-S°.L-1. No adjustment was made for initial pH. Experiments were carried out in duplication at 30°C and 180rpm in a shaking incubator. 614
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2.5 Analyses of metal and sulfate concentrations Ten ml of sample from the flask was centrifuged for 15 minutes followed by filtration of the supernatant to measure concentrations of sulfate and metal in the filtrate. A finely ground dewatered sludge (0.1 g) was suspended in a Teflon vessel containing 5 mL of mixed acid solution, such as fluoric acid, nitric acid, and perhydrochloric acid (4:4:1, v/v) to determine initial concentration of heavy metals in the sludge. The vessel was heated at 150 to 200°C to ensure a complete digestion of the sludge. Concentrations of heavy metal were measured by using Inductively Coupled Plasma Spectroscopy (ICP, Plasma 40, Perkin Elmer, USA). Sulfate concentrations in the medium were determined by Ion Chromatography (Waters 510, USA; conductivity detector, Waters 432; IC-Pak-TM Anion column: 4.6 mm φ x 50 mm L) at 1.2 mL.min-1 and 35°C along with sodium borate/gluconate solution as a mobile phase. 3.
RESULTS AND DISCUSSION
3.1 Effect of inoculation In general, a rapid rate of pH reduction and low final pH of the sewage sludge are indicatives of effective microbial leaching of heavy metals. Four systems prepared at different conditions were analyzed for the pH changes to elucidate effect of A. thiooxidans MET inoculation (Fig. 1).
Figure 1. Effect of inoculation on pH reduction of sludge solution ( ; sterilized sludge, {; nonsterilized sludge, T; nonsterilized sludge with the inoculation of mixed sulfur oxidizing bacteria, ; nonsterilized sludge with the inoculation of A. thiooxidans MET)
No pH change was shown in sterilized sludge indicating the absence of sulfur oxidation (System I). On the other hand, in nonsterilized sludge (System II) an initial lag phase was shown in five days of incubation followed by subsequent decrease in pH, due to the presence of indigenous sulfur oxidizing bacteria in the sludge. The lag phase suggested that these indigenous bacteria could need about five days for initial growth. Inoculation with the mixed sulfur oxidizing bacteria obtained from this study (System III) resulted in rapid pH reduction with a short lag phase (one day). Its final pH was 1.4 in 14 days of incubation. However, in the system inoculated with A. thiooxidans MET (System IV), no initial lag phase was shown and the rate of pH reduction was much faster than that of 615
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System III. The pH decreased to 1.2 in four days of incubation. These results indicated that leaching efficiency of A. thiooxidans MET isolated in this study was much greater than those of the mixed sulfur oxidizing cultures and indigenous sulfur oxidizing bacteria in the sludge. 3.2 Effect of sludge solids concentration With exception of a few studies that examined about 70 g.L-1 sludge solids concentration [15, 17], most of the previous studies in microbial metal leaching employed low sludge solids concentration (13-38 g.L-1). Leaching of heavy metals from the sludge at high solids concentration is less expensive process since large amounts of heavy metals can be extracted from a small volume of the sludge. Therefore, in this research microbial metal leaching by the strain MET was examined at various sludge solids concentrations such as 20, 50, 90, and 130 g.L-1. The experimental results are illustrated in Fig. 2. The initial pH of the sludge solution inoculated with the strain MET increased with increase in sludge solids concentrations. At high sludge concentrations, such as 90 and 130 g.L-1, an increase in sludge pH was observed during the first half day. However, after the initial increase the sludge pH decreased considerably in all the sludge solids concentrations, due to the oxidation of elemental sulfur to sulfuric acid by the strain MET. The rate of pH reduction was higher at lower sludge solids concentration and became lower at pH less than 2.0. The slow acidification at high solids concentration was due to an increase in buffering capacity as described in previous studies [4, 15, 16, 19]. The buffering effects were mainly due to the presence of basic components such as carbonate [19], which hindered microbial leaching efficiency [4, 15, 16]. After 5 to 8 d of leaching, the final pH of the sludge solution decreased to 1.2, 1.3, 1.4, and 1.6 for 20, 50, 90, and 130 g.L1 sludge solids concentrations, respectively. The sludge ORP continuously increased with increasing reaction time and the increase in the rate of ORP was rapid at low solids concentrations (Fig. 2b). The final ORPs obtained for 20, 50, 90, and 130 g.L-1 were 420, 42, 390, and 370 mV, respectively. As shown in Fig. 2c, the concentration of sulfate produced by the strain MET increased with time and, regardless of the sludge solids concentration, was approximately the same at a given time up to five days of incubation. The increase in sulfate concentration, however, was restrained by further incubation. During the initial incubation period the pH reduction of the solution inoculated with the strain MET appeared differently at various sludge solids concentrations, but with similar sulfate concentrations (Fig. 2a and 2c). Since concentration of elemental sulfur in the sludge solution was 10 g.L-1, stoichiometric sulfate concentration should be 30 g.L-1. The results shown in Fig. 2c, therefore, revealed that approximately 50-60% of sulfur had been oxidized during 6-8 days of incubation. The heavy metals removal at different solids concentrations was calculated for six days leaching period (Table 1). Removal efficiencies of heavy metals decreased with increase in sludge solids concentrations. Longer reaction time was generally needed (Fig. 2a) to produce sufficient pH reduction of leaching solution at high sludge solids concentrations such as 90 and 130 g.L-1. With increase in leaching time to eight days, the removal efficiencies of metals at high solids concentrations were comparable to those at low solids concentrations. In addition, even for longer leaching period the removal efficiency of Pb was substantially lower at high solids concentrations. This indicated that leaching of Pb with the strain MET could be difficult at high sludge solids concentration. Nevertheless, these results 616
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confirmed the high efficiency of the strain MET in microbial leaching of heavy metals from the dewatered sludge or the highly concentrated sludge solution. (a) pH
6 5
pH
4 3 2 1 0 (b) ORP
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400 300 200 100
Sulfate (g L-1)
0 (c) Sulfate
20 15 10 5 0
0
1
2
3
4
5
6
7
8
9
Time (d)
Figure 2. pH (a), ORP (b), and sulfate concentration (c) at various sludge solids concentrations during microbial leaching with A. thiooxidans MET ( ; 20 g.L-1. ; 50 g.L-1, ; 90 g.L-1, ; 130 g.L-1) Table 1. Removal efficiency of heavy metals at different solids concentration after 6 days treatment Removal efficiency of heavy metal (%)
Solids conc. (g.L-1)
Zn
Cu
Cr
Pb
Ni
Al
20
99
67
69
99
87
36
50
96
55
65
93
84
33
90
92 (93)
56 (62)
60 (65)
48 (52)
82 (83)
32 (36)
130
91 (91)
51 (57)
47 (62)
30 (33)
84 (84)
29 (33)
Parenthesis: removal efficiency after 8 days treatment.
Fig. 3 shows removal efficiency of the heavy metal at various pHs for all the sludge solids concentrations. The results showed that removal efficiency of metal and solution pH were highly correlated each other with exception of Pb. Low correlation of Pb was mainly due to the precipitation of PbSO4 generated by complex formation of sulfate and Pb ions [20].
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80
60 r2 = 0.97
40
Removal Cr (%)
60
r2 = 0.96
40
20
20 (c)
0
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80 60
60
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40
40 20
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80 2
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100
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Removed Pb (%)
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6
Removed Al (%)
Removed Zn (%)
100
0
pH
Figure 3. Relationship between pH and removal efficiency of heavy metals at different sludge solids concentrations. (a) Zn, (b) Cu, (c) Cr, (d) Pb, (e) Ni, (f) Al ( ; 20 g.L-1, ; 50 g.L-1, ; 90 g.L-1, ; 130 g.L-1)
It should be also noted that the leaching efficiency of Zn and Ni at pH 4 was greater than 20%, whereas that of other metals at the same pH was almost zero. These results suggested easy leaching of Zn and Ni even at high pHs. Nevertheless, the pH of leaching solution should be decreased to 1-1.5 for a sufficient removal of all the types of heavy metals at various sludge solids concentrations. As described in Fig. 2, high buffering capacity at high sludge solids concentrations retarded pH reduction, resulting in low leaching efficiency of metals. Data points illustrated in Fig 4 were obtained by dividing sulfate concentration and amount of leached metal by amount sludge solids at all the solids concentration. By eliminating effect of buffering capacity on metal leaching caused by disparity in solids concentration, a direct interaction between sulfate concentration and metal leaching can be found by this normalization. Results shown in Fig. 4 clearly point out that regardless of sludge solids concentration the amount of metals leached out per unit sludge mass was practically identical at the same sulfate concentration per unit sludge mass. Initially, there was a linear correlation between sulfate concentration per unit sludge mass and amount of leached metal per sludge. For all the metals, however, amount of leached metal per unit sludge mass reached equilibrium at a certain sulfate concentration per unit sludge mass. Minimum sulfate 618
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Leached metal (g kg-sludge-1)
concentrations for maximum metal leaching, i.e., sulfate concentration reaching in equilibrium of the metal leaching, for various metals were in the order of Zn 1.15, Ni 1.22, Al 1.25, Cu 1.29, Cr 1.41, and Pb 2.21 g-SO42- kg-sludge-1. The slopes (g-metal g-sulfate1 ) of initial 1st-order line (before equilibrium) from the relationship between sulfate concentration and amount of leached metal shown in Fig 4 are as follows; 2.63 for Zn, 2.43 for Cu, 3.42 for Cr, 0.85 for Pb, 1.93 for Ni, and 4.42 for Al. The greater the slope value was, the greater influence of sulfate concentration on metal leaching. The influence decreased in the order of Al >> Cr > Zn > Cu> Ni >> Pb. With knowledge on such a relationship, the sole measurement of sulfate concentration in the sludge solution could permit to estimate amounts of leached metal. Al (r2 = 0.86)
10
Zn (r2 = 0.83) Cu (r2 = 0.91)
1
Cr (r2 = 0.91) Ni (r2 = 0.91) Pb (r2 = 0.91)
0.1
0.01 0.01
0.1
1
Sulfate concentration (g kg-sludge-1)
Figure 4. Relationship between sulfate concentration and concentration of heavy metals solubilized ({; Al, ; Zn,
; Cu, ; Ni, U; Cr, S; Pb) 4.
CONCLUSIONS An acidophilic, sulfur-oxidizing Acidithiobacillus thiooxidans MET showed sulfuroxidizing ability at both acidic and neutral conditions, and allowed metal leaching even at a high sludge solids concentration. Metal leaching rate was faster at low sludge solids concentration than that at high solids concentration, due to a faster reduction of pH at low solids concentration. Minimum sulfate concentrations for maximum metal leaching (i.e., sulfate concentrations produced per unit sludge at maximum metal removal) for various metals were in the order of Zn 1.15, Ni 1.22, Al 1.25, Cu 1.29, Cr 1.41, and Pb 2.21 g SO42- per kg sludge. The relationship between the pH of the sludge solution and metal leaching efficiency was also obtained. REFERENCES
1. 2. 3. 4.
R.D. Tyagi, D. Couillard, and F.T. Tran, Proc. Biochem., 26 (1991) 47. R.D. Tyagi, J.F. Blais, J. Auclair, and N. Meunier, Wat. Environ. Res., 65 (1993) 196. D.K. Jain and R.D. Tyagi, Enzym. Microb. Technol., 14 (1992) 376. T.R. Sreekrishnan, R.D. Tyagi, J.F. Blais, and P.G.C. Campbell, Wat. Res., 27 (1993) 1641. 5. J.F. Blais, R.D. Tyagi, and J.C. Auclair, J. Environ. Sci. Heal., A28 (1993) 443. 619
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6. G.G. Oliver and J.H. Carey, Wat. Res., 10 (1976) 1077. 7. D.J. Wozniak and J.Y.C. Huang, Res. J. Wat. Pollut. Control Fed., 54 (1982)1574. 8. K.S.L. Lo and Y.H. Chen, Sci. Total Environ., 90 (1990) 99. 9. R.D. Tyagi, D. Couillard, and F.T. Tran, Environ. Pollut., 50 (1988) 295. 10. T.R. Sreekrishnan and R.D. Tyagi, Proc. Biochem., 31 (1996) 31. 11. R.D. Tyagi and D. Couillard, Bacterial leaching of metals from sludge. p. 537-591. In P.E. Cheremisinoff (ed.), Encyclopedia of Environmental Control Technology, Gulf Pubilishing Co., Texas, 1989. 12. R.D. Tyagi, J.F. Blais, and B. Boulanger, J. Environ. Sci. Heal., A28 (1992) 1361. 13. D. Couillard and G. Mercier, War. Res., 27 (1993) 1227. 14. T.R. Sreekrishnan, R.D. Tyagi, J.F. Blais, N. Meunier, and P.G.C. Campbell, Wat. Res., 30 (1996) 2728. 15. R.D. Tyagi, J.F. Blais, N. Meunier, and H. Benmoussa, Wat. Res., 31 (1997) 105. 16. K.S. Cho, H.W. Ryu, and H.S. Moon, J. Kor. Soc. Environ. Eng., 21 (1999) 433. 17. H.W. Ryu, Y.J. Kim, K.S. Cho, K.S. Kang, and H.Choi, Kor. J. Biotechnol. Bioeng., 13 (1998) 279. 18. K.S. Cho, L. Zhang, M. Hirai, and M. Shoda, J. Ferment. Bioeng., 71 (1991) 44. 19. C. Brombacher, R. Bachofen, and H. Brandl, Appl. Environ. Microbiol., 64 (1998) 1237. 20. G. Mercier, M. Chartier, and D. Couillard, Wat. Res., 30 (1996) 2452.
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Mercury removal by polymer-enhanced ultrafiltration using chitosan as the macroligand E.K. Kuncoro#, T. Lehtonen, J. Roussy and E. Guibal∗ Ecole des Mines d’Alès, Laboratoire Génie de l’Environnement Industriel, 6 avenue de Clavières, F-30319 Alès cedex, France Abstract Chitosan is an aminopolysaccharide produced from crustacean shells. This biopolymer is very efficient at sorbing metal ions through different mechanisms including (a) chelation on its amine group in near-neutral solutions or (b) ion exchange in acidic solutions (protonated amine groups interact with metal anions). Alternatively, dissolved chitosan can be used for the recovery of metal ions using Polymer-Enhanced Ultrafiltration (PEUF) technique. The process makes profit of the ability of the macromolecules to bind metal ions and to be retained by membranes of selected cut-off. The process has been tested in the case of mercury using an Amicon ultrafiltration system. The efficiency of the process has been investigated using membranes of different cut-off (10 kDa, 50 kDa, 100 kDa). The influence of pH, metal concentration, chitosan concentration and pressure has been studied on permeation flux and the retention of both the polymer and the metal. The coupling of metal binding and ultrafiltration allows reaching retention efficiency greater than 95% under appropriate experimental conditions. For mercury, metal recovery occurs by chelation and the optimum pH is close to pH 5-6. The mercury/chitosan binding constant has been determined close to 107.3. Keywords: chitosan, ultrafiltration, mercury, chelation, retention efficiency, pressure drop 1.
INTRODUCTION The need for efficient and economical recovery of metals from dilute effluents is motivated by the drastic development of environmental regulations. To face up the problem of the removal of toxic or valuable metals from dilute effluents, biosorption processes have been developed for the last twenty years. Many biosorbents have been tested for the recovery of a broad range of metals, especially bacteria, fungi, algae [1]. More recently a great attention has been paid to the use of agricultural waste materials, or wastes from food industry [2]. An increasing interest has focused on the use of biopolymers extracted from marine materials: alginate (from algae), chitin and chitosan (from crustaceans). The presence of chelating functions on these biopolymers (carboxylic
#
E.K.K. thanks the French Ministry of Foreign Affairs for his Ph.D fellowship. E.G. The authors thank the European Community for financial support under Growth Program (3SPM project, Contract G1RD-CT2000-00300) for attending the IBS’03 conference. ∗
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functions for alginate, amine groups for chitosan) may explain the growing interest in checking their ability to recover metal ions [3-5]. Chitosan is an amino-polysaccharide that is produced by an alkaline deacetylation from chitin, the most abundant biopolymer in nature after cellulose. The presence of amine groups may explain its high efficiency for the chelation of metal cations in nearneutral solutions [6-7]. However, due to its cationic behavior [8], in acidic solutions the protonation of amine groups leads to interesting ion exchange properties [9-11]. Metal anions are efficiently sorbed by electrostatic attraction. However, chitosan is soluble in most mineral and organic acids (with the exception of sulfuric acid); it is thus necessary to crosslink the polymer to prevent its dissolving in acid media for an easy recovery of loaded material. Alternatively, the biopolymer can be used in its dissolved form providing a suitable filtration system is used for the recovery of loaded macromolecules. This property can be used in polymer enhanced ultrafiltration processes (PEUF) [11-12]. The main element of any membrane separation process is the semipermeable membrane. Certain solution components will pass through the membrane forming the permeate, whereas others will be retained by the membrane forming the retentate or the concentrate. The retention of the component depends on many parameters [13], including solution type, solution composition, pH, temperature, membrane material, pore size, hydrodynamics etc… In many cases, however, the size of the dissolved component is the crucial factor for the retention. In order to improve separation, the metal ions can be bound to macromolecules, thus enlarging the molecular dimensions of the components to be separated [14-18]. This study focuses on the use of chitosan dissolved in HCl solutions for the binding of mercury and the further separation of loaded macromolecules by an AMICON ultrafiltration unit. The efficiency of the process has been investigated using membranes of different cut-off (10 kDa, 50 kDa, 100 kDa). The influence of pH, metal concentration, chitosan concentration and pressure has been studied on both flux and retention of polymer and metal. Ultrafiltration performance can be correlated to the sorption behavior of the biopolymer and it is thus possible to predict the uptake performance of dissolved biopolymer, and to anticipate on the chemical modifications that can be required to increase its efficiency. 2.
MATERIAL AND METHODS
2.1 Materials Chitosan was supplied by Aber Technologies (Plouvien, France). The characteristics of the samples have been previously determined. The deacetylation degree was found to be 87%, using FTIR spectrometry measurements, and the molecular weight was 125,000 g mol-1, using SEC measurements. Mercury nitrate was purchased from Fluka AG (Switzerland). Other common reagents were supplied by Carlo Erba (Italy). 2.2 Ultrafiltration module Ultrafiltration experiments have been performed using an AMICON ultrafiltration module, AMICON 8400, with the following characteristics: membrane diameter 76 mm, volume of solution 200 mL, maximum pressure 75 psi (5.3 bars). The membranes used for these experiments were Amicon membranes (M10, M50 and M100) made of polyethersulfone with different molecular weight cut-off (MWCO): 10,000, 50,000 and 100,000 Da, respectively. Pressure was obtained from local pressurized air network using 622
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suitable manometers. Experiments were performed under pressures ranging between 0 and 2.5 bars. 2.3 Ultrafiltration experiments Chitosan was dissolved in HCl solutions at suitable concentration and the solution was mixed with metal solution at fixed concentrations. The mixture was agitated under pH control for at least 2 hours prior being placed in the ultrafiltration module. For the measurement of flux, the first 10-mL fraction was removed and then the time necessary to filtrate successive (at least 5) fractions of 25 or 50 mL were monitored. The flux (J, L m2. -1 h ) was calculated as the mean value of these 5 time-fractions taking into account the volume passed through the membrane during the nominal times and the surface of the membrane. Samples have been collected from these fractions for the determination of both metal and polymer content. Polymer concentration was determined using a TOC-meter Shimadzu TOC-5000 (Japan) and a calibration curve prepared from pure chitosan solutions. Metal concentration was determined using an Inductively Coupled Plasma spectrometer (JOBIN-YVON JY 2000, France). 3.
RESULTS AND DISCUSSION
3.1 Characterization of permeation properties of ultrafiltration membranes The permeation properties of the membranes have been determined using Milli-Q demineralized water. The fluxes have been measured (Figure 1, left panel)). Flux linearly varied with the pressure. Moreover, the slope of the curves, s, may be correlated to the cut-off, CO (Da), of the membranes according to the equation: s = 0.0111 CO (R2: 0.969). This equation may be used to predict the permeate flux for a given pressure and a given membrane (fixed cut-off): J = 0.0111 CO * P. The right panel in Figure 1 compares experimental data with calculated values.
3000 M50
R = 0.976
2
-2
M100
J = 464.18 P 2
1000
R = 0.971 J = 185.92 P 2
0
J calc. = 0.9866 J exp.
-1
M10
J = 1146.1 P
J calc. (L m h )
2000
-2
-1
J (L m h )
3000
R = 0.963
0 0.5 1 1.5 2 2.5 3 P (bar)
2
R = 0.9698
2000 1000 0 0
1000 2000 -2 -1 J exp. (L m h )
3000
Figure 1. Flux characteristics of ultrafiltration membranes 3.2 Ultrafiltration of chitosan solutions The permeation characteristics of chitosan have been determined on M10 (cut-off: 10,000 Da) membranes at a polymer concentration of 200 mg.L-1. Results are presented on Figure 2. As expected, increasing the pressure increased the permeation flux: it was doubled from 0.5 to 2.5 bars, while the retention rate only slightly varied (around 99%).
623
-1 -2
100
100
80
99
60
J = 21.6 LnP + 59.4 2
R = 0.914
40
J (L/m2 h)
20
R (%)
0
98 97 96 95
Retention Rate (%)
J (L m h )
Bioremediation Environmental Applications
0 0.5 1 1.5 2 2.5 3 P (bar)
Figure 2. Influence of pressure on permeation flux (J) and polymer retention rate (R) (membrane cut-off: 10.000 Da; chitosan concentration: 200 mg L-1)
Table 1 shows the influence of chitosan concentration on the flux and retention rate at pressures (P) 1 and 2 bars for membranes M10, M50 and M100. In most cases, the retention rate exceeded 92%. With the M100 membranes, at low chitosan concentration the retention was lower, in the range 86-89%, whatever the pressure applied to the membrane. Increasing polymer concentration partially restored retention efficiency (above 93%). In the best cases, the retention of the polymer reached 97%, but it was not possible to recover the entire amount of the polymer. This may be explained by the presence of small polymer chains (partially hydrolyzed) that cannot be retained by the membranes, even with the smallest cut-off (M10). This polymer loss could be avoided by a preliminary ultrafiltration of the polymer solution to remove small size polymer chains. Table 1. Permation fluxes (L.m-2.h-1) and retention rates (into brackets, %) for chitosan ultrafiltration in function of polymer concentration, pressure and membrane type P (bar) C (mg/L) M10 M50 M100
1
2
0
50
100
200
0
50
100
200
196.0
154.5 (96.2) 231.7 (96.5) 233.9 (86.5)
142.3 (96.2) 172.2 (93.3) 194.1 (93.2)
131.5 (96.4) 147.8 (96.9) 136.0 (96.8)
362.2
185.4 (94.8) 273.9 (95.1) 280.9 (89.3)
170.5 (96.2) 217.0 (92.4) 272.8 (92.8)
158.2 (97.0) 173.4 (97.2) 161.0 (95.5)
554.0 1358.4
926.7 2253.7
Table 1 also shows the permeation fluxes for selected experimental conditions. These fluxes can be compared to those obtained with pure water filtration. The presence of chitosan strongly decreased the permeation fluxes, especially for large cut-off membranes and for large pressures. Indeed, in the case of M10 membranes at P: 1 bar, the decrease of the permeation flux was between 21 and 33%, but when the pressure increased to P: 2 bars, the decrease reached 49-56%. With the M100 membranes the influence of chitosan on permeation fluxes was much more drastic: at the highest polymer concentration the decrease in permeation fluxes raised to 90-93%. Similar decrease in permeate fluxes have been observed by Tangvijitsri et al. in the case of PEUF of chromate, sulfate and nitrate [15]. This decrease in the flux may be explained by the accumulation of polyelectrolyte
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near the membrane surface (hydrodynamic boundary layer), though the agitation was maintained in the ultrafiltration cell. Figure 3 shows the influence of pH on the ultrafiltration of chitosan (initial concentration: 200 mg L-1) with M50 and M100 membranes under a pressure of 2 bars. The permeation flux shows an optimum at pH 4 that appears to be independent of the membrane type. The permeate flux can be fitted by a quadratic equation. The retention rate varied between 90 and 98%, with a maximum retention obtained at pH 4. Generally, polymers have flexible structure and under certain hydrodynamic conditions the shape of such molecules can change from a sphere to an ellipse or to a slim cylinder so that they will pass through the membrane even though their mass is less than the nominal cut-off of the membranes used. This effect is emphasized by the pH of the solution that changes the conformation of the polymer (hydrogen bonding, electrostatic repulsion) [14]. The presence of metals ions partially neutralizes their charge, which in turn influences the conformation of polymer chains and their overall size. 100
J (M50) J (M100)
150 100 2
50
J = -18.8(pH) + 150.16(pH) - 140.09 2
R = 0.9904
0 1
2
3
4 pH
5
6
7
Retention Rate (%)
-2
-1
J (L m h )
200
98 96 94 92
R (M50)
90 88
R (M100)
1
2
3
4 pH
5
6
7
Figure 3. Influence of pH on permeation flux and retention rate for chitosan solutions (conc.: 200 mg L-1; P: 2 bars; and membranes M50 and M100) 3.3 Influence of pH on mercury removal by PEUF Figure 4 shows the permeation properties (flux, retention rates for mercury and polymer) in function of the pH of the mixed (Hg/chitosan) solutions. The maximum permeation flux (around 150 L.m-2.h-1) was obtained between pH 3 and pH 4. The flux was halved at changing the pH to pH 2 or pH 5-6. These values are comparable to those obtained with chitosan solutions in absence of mercury. For polymer recovery, the maximum retention rates (around 97-98%) were obtained at pH 2.5-3.5: again, these values are comparable to those obtained without mercury. The retention rate decreased to 94-96% when the pH was increased to values 5-6. It is slightly higher than the levels of chitosan recovery in pure polymer solutions, especially at pH 6. This increase in retention efficiency may be explained by the linking effect of metal-polymer interaction that can increase the size of the polymer chains and allows the complementary retention of small size chains, since mercury is supposed to interact with different polymer chains. Indeed, different studies on mercury or copper (in presence of EDTA) recovery by PEI(polyethyleneimine) binding/ultrafiltration have pointed out the change in the stoichiometry between metal ions and amine/imine groups [12,18-20]. These studies have shown the possibility for different chains to contributing to metal ion binding.
625
100
100
150
98
80
96
60
94
40
92
20
90
0
R (C, %)
-1 -2
200
100 J (M50)
50
J (M100)
0 1
2
3
4 pH
5
6
7
1
2 3
4 5 6 pH
R (Hg, %)
J (L m h )
Bioremediation Environmental Applications
7
Figure 4. Influence of pH on permeation fluxes (left panel); and polymer (open symbols) and mercury (closed symbols) retention rates (right panel) (M50 and M100 membranes; P: 2 bars; Chitosan concentration: 200 mg.L-1)
More interesting are the retention curves for mercury. A sigmoid trend was observed at increasing the pH. From pH 2 to pH 3 the retention rate decreased from 15 to 4% and then increased drastically up to pH 6 with a maximum mercury retention that exceeded 95% at pH 6. Blank experiments have been performed with pure mercury solutions (i.e. without chitosan) and mercury retention was less than 4% below pH 4, and progressively increased to 10-14% when the pH was raised to 6. This mercury retention may be explained by a partial adsorption of mercury on the membrane rather than to precipitation of mercury under selected experimental conditions. So, chitosan significantly improved the efficiency of mercury recovery. Optimum pH conditions for mercury recovery appear to be close to pH 6, though the permeation flux is significantly reduced at these pH values. The non-null recovery of mercury at very low pH (around pH 2-2.5) may be explained by a different interaction mechanism. Indeed, at low pH, controlled with hydrochloric acid, the concentration of chloride is high enough to influence the speciation of mercury and to displace the distribution of its species toward the formation of anionic chloro-complexes. Under these acidic conditions amine functions of chitosan are fully protonated and then available for the ionic attraction of anionic species. This interpretation is consistent with mechanisms cited for mercury removal in acidic solutions on cross-linked materials [4]. It is interesting to note that the preliminary ultrafiltration of the polymer solution (prior to contact with metal ion solution) would increase metal retention. Indeed, metal ions can bind to small size chains of the polymer, which are not retained by the membrane. It induces a loss of both metal ions and polymer. According to Rumeau et al., the logarithm of the apparent complex formation constant should vary with the pH of the solution (far from the pKa of the polymer) and the slope of the curve should be n, the number of ligand molecules per metal ligand complex [21]. Similar equation has been established by Juang and Chen in the case of poly(ethylenimine) with several simplifying hypotheses [22]. More specifically, the equation was demonstrated considering that metal ion is only present in the solution in the free form. Obviously, the solution of the problem would be more complex taking into account the speciation of metal ions Different values of n have been tested (n: 1, 2, 3 and 4) using the following equation for the calculation of the apparent complex formation constant (β’n) [21]:
626
Bioremediation Environmental Applications n 1 1⎡ ⎛M⎞ ⎞ ⎤ n⎛ = ⎢(1 − R )(L) ⎜⎜1 − n ⎜ ⎟R ⎟⎟ ⎥ β 'n R ⎢⎣ ⎝ L ⎠ ⎠ ⎥⎦ ⎝
(1)
with R retention yield, M and L metal and ligand concentrations, respectively. Figure 5 shows the plots of the apparent formation constant β’n of the complex Hg/Chitosan versus the pH of the solution using the equation (1) with n = 1 and n = 2. Though each of the stoichiometric ratios tested gave linear trends in the range pH 3-5, the curve Log β’2 versus pH was the only curve that gave consistent slope (Table 2) (slope close to the theoretical stoichiometric ratio). So, the data have been exploited using the equation (2): 2 1 1⎡ ⎛M⎞ ⎞ ⎤ 2⎛ (2) = ⎢(1 − R )(L) ⎜⎜1 − ⎜ ⎟R ⎟⎟ ⎥ β '2 R ⎢⎣ ⎝ ⎝ L ⎠ ⎠ ⎥⎦
2
5 4
1
3 2
3
4 pH
5
Log Beta(1) ( )
3
M100
4
6
3
5
2
4
1
3 2
6
7
3
4 pH
5
Log Beta(2) (o)
7 6
M50
4
8
5
8 Log Beta(2) (o)
Log Beta(1) ( )
5
6
Figure 5. Plot of Log βi as a function of pH for membranes M50 and M100 (M50 and M100 membranes; P: 2 bars; Chitosan concentration: 200 mg.L-1) Table 2. Determination of the stoichiometry of the complex by the equation (1): slope of the curves Log βi versus pH n 1 2 3 4
M50 1.68 1.76 1.30 1.45
M100 1.58 1.66 1.59 1.75
3.4 Influence of Ligand/Metal (L/M) molar ratio on mercury removal by PEUF Figure 6 shows the influence of the ligand/metal ratio on the retention efficiency for mercury on membranes M50 and M100 under a pressure of 2 bars. Metal concentration was varied between 10 and 100 mg Hg L-1, while chitosan concentration was varied between 50 and 200 mg.L-1. The pH of the solution was controlled to pH 5.5. When the concentration of mercury increased, a slight increase in the permeation rate was also observed, whatever the concentration of chitosan in the mixture. As expected increasing chitosan concentration decreased the permeate flux. The positive effect of mercury may be related to the formation of polymer/metal aggregates that prevents the blocking of membrane pores during the formation of the polarization layer. The apparent constant of complexation was calculated using the samples collected with membranes M50 and M100 at pH 5.5 for the different L/M ratios using equation (2). Figure 7 shows the plot of Log β2 for the different L/M ratios. The mean values obtained 627
Bioremediation Environmental Applications
Retention Rate (%)
over 11 values (removing the data obtained at low chitosan concentration with L/M=1 that was not consistent with other experimental points) were log β’2 = 7.29 ± 0.38 for membranes M 50 and M100. 100 80 60 40
R (Hg-M50)
20
R (Hg-M100)
0 0
5
10
15
20
25
L/M (mol/mol)
400 J (M 50/C50)
-2
-1
J (L m h )
300
J (M 100/C50) J (M 50/C100)
200
J (M 100/C100) J (M 50/C200)
100
J (M 100/C200)
0 0
25
50
75 100 -1
Hg Conc. (mg Hg L ) Figure 6. Influence of ligand/metal ratio on mercury retention (left panel) and permeation flux (right panel) at pH 5.5, with membranes M50 and M100, under P: 2 bars
14 Chit: 50 mg/L
M50
LogBeta(2)
LogBeta(2)
12 Chit: 100 mg/L
10
Chit: 200 mg/L
8 6
Chit: 50 mg/L
M100
12
Chit: 100 mg/L Chit: 200 mg/L
10 8 6
0
5
10
15
L/M (mol/mol)
20
25
0
5
10
15
20
25
L/M (mol/mol)
Figure 7. Influence of ligand/metal ratio on Log β2 at pH 5.5, with membranes M50 and M100, under P: 2 bars
Figure 8 shows the influence of L/M ratio on mercury uptake capacity (mercury content on retained polymer) and the plot of uptake capacity (q) versus mercury residual concentration (Ceq). The uptake capacity (mg or mmol per g of polymer) was converted into a molar unit system (mmol metal per mol of polymer). Indeed, for a chitosan sample 628
Bioremediation Environmental Applications
with a 87% deacetylation degree, the molecular weight of the equivalent monomeric unit is 166 g per mol. The mathematical modeling of experimental data was performed using the Langmuir equation: q m b C eq q= (3) 1 + b C eq with b (L.mg-1 or L.mmol-1) and qm (mg.g-1 or mmol.g-1) are Langmuir-model constants. This equation was used by analogy with metal ion sorption on solid particles, though the physical hypotheses of the sorption model were not satisfied in the present case. The equation served to approximate the theoretical maximum uptake capacity that was found to be 579 µmol Hg mol-1 chitosan. The maximum experimental uptake capacity was 499.5 µmol Hg mol-1 chitosan. It means a stoichiometric ratio Hg/Chitosan tending to 1:2, close to the value established above. The maximum theoretical uptake capacity (from Langmuir equation) was close to 698 mg Hg g-1 chitosan, slightly higher than the maximum experimental uptake capacity, which is close to 602 mg Hg g-1 chitosan. For a residual concentration of 20 mg Hg L-1 (0.1 mmol Hg L-1), the uptake capacity was about 298 µmol Hg mmol-1 chitosan (i.e. 360 mg Hg g-1). Under comparable conditions (pH 5-6, and residual concentration tending to 20 mg Hg L-1, not shown), the sorption capacity obtained with chitosan flakes was about 240-250 µmol Hg mmol-1 chitosan (i.e. 290 mg Hg g-1). This increase in mercury uptake capacity by chitosan using the polymer in PEUF (compared to sorption process) may be explained by a greater accessibility to reactive sites when the polymer is dissolved and highly dispersed in the solution. In the solid form the biopolymer maintain a residual crystallinity that restricts the accessibility to sorption sites [23], moreover some functional groups (amine functions) are involved in hydrogen bonding with water and other amine groups (inter and/or intra-molecular hydrogen bonds): it limits the availability of functional groups for other interactions [24]. 600 -0.74
q = 448.8 (L/M) 2
400
M50 M100
R = 0.945
200 0
q (mmol Hg/mol
q (mmol Hg/mol
600
q m = 579.43 mmol Hg/mol b: 10.6 L/mmol
400
2
R : 0.945
0 0
5
10
15
L/M (mol/mol)
20
25
M50 M100 Langmuir
200
0
0.1
0.2
0.3
0.4
Ceq (mmol Hg/L)
Figure 8. Influence of ligand/metal ratio on mercury uptake by the polymer (left panel) and uptake isotherm (right panel) at pH 5.5, with membranes M50 and M100, under P: 2 bars 4.
CONCLUSIONS Polymer enhanced ultrafiltration appears to be a promising technique for the recovery of mercury using chitosan as the chelating agent. The chelation of mercury by chitosan increases with pH. However, due to chitosan precipitation above pH 6, for chitosanenhanced ultrafiltration processes pH might be set below pH 6. The permeation flux and 629
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retention rate are very sensitive to pH (optimum close to pH 4) for chitosan retention. In the presence of mercury both polymer retention and permeation rates are maintained optimum at a value close to pH 4. The formation constant for the complex Hg-chitosan is found close to 107.3. The maximum uptake capacity is close to 600 mg Hg g-1, higher than the levels reached with solid chitosan. This enhancement in sorption capacities may be due to a greater accessibility to sorption sites when chitosan is dissolved. It is also interesting to observe that the dissolving of the polymer allows increasing uptake kinetics. Indeed, a 2-hours contact time prior to ultrafiltration was maintained for the equilibration of the solution (pH stabilization and metal chelation). This contact time was sufficient to achieve a complete uptake of mercury. Under comparable experimental conditions (sorbent dosage and metal concentration), 24 hours of contact were necessary to reach the equilibrium with chitosan in solid form (not shown, in preparation). REFERENCES
1. 2. 3. 4.
B. Volesky and Z.R. Holan, Biotechnol. Prog., 11 (1995) 235. J.R. Deans and B.G. Dixon, Wat. Res., 26 (1992) 469. L.K. Jang, D. Nguyen and G.G. Geesey, Wat. Res., 33 (1999) 2817. Y. Kawamura, M. Mitsushashi, H. Tanibe, and H. Yoshida, Ind. Eng. Chem. Res., 32 (1993) 386. 5. K. Inoue, in: Recent Advances in Marine Biotechnology, Volume 2, Environmental Marine Biotechnology, M. Fingerman, R. Nagabhushanam and M.-F. Thompson, eds., Oxford & IBH Publishing PVT. Ltd, New Delhi, pp. 63-97, 1998. 6. R. Bassi, S.O. Prasher and B.K. Simpson, Sep. Sci. Technol., 35 (2000) 547. 7. M.S. Dzul Erosa, R. Navarro Mendoza, T.I. Saucedo Medina, M. Avila Rodriguez, E. Guibal, Hydrometallurgy, 61 (2001) 157. 8. P. Sorlier, A. Denuzière, C. Viton and A. Domard, Biomacromolecules, 2 (2001) 765. 9. E. Guibal, C. Milot, and J.M. Tobin, Ind. Eng. Chem. Res., 37 (1998) 1454. 10. E. Guibal, T. Vincent, A. Larkin and J.M. Tobin, Ind. Eng. Chem. Res., 38 (1999) 4011. 11. J. Guzman, I. Saucedo, J. Revilla, R. Navarro and E. Guibal, Langmuir, 18 (2002) 1567. 12. R.-S. Juang and C.-H. Chiou, J. Membr. Sci., 165 (2000) 159. 13. K.E. Geckeler and K. Volchek, Environ. Sci. Technol., 30 (1996) 725. 14. B.L. Rivas, E. Pereira, I. Moreno-Villoslada, Prog. Polym. Sci., 28 (2003) 173-208. 15. S. Tangvijistri, C. Saiwan, C. Soponvuttikul and J.F. Scamehorn, Sep. Sci. Technol., 37 (2002) 993. 16. K. Volchek, E. Krentsel, Y. Zhilin, G. Shtereva and Y. Dytnersky, J. Membr. Sci., 79 (1993) 253. 17. B.L. Rivas and I. Moreno-Villoslada, J. Membr. Sci., 178 (2000) 165. 18. J. Müslehiddinolu, Y. Uluda, H. Önder Özbelge and L. Yilmaz, J. Membr. Sci., 140 (1998) 251. 19. Y. Uludag, H. Önder Özbelge and L. Yilmaz, J. Membr. Sci., 129 (1997) 93. 20. R.-S. Juang and M.-N. Chen, J. Membr. Sci., 119 (1996) 25. 21. M. Rumeau, F. Persin, V. Sciers, M. Persin and J. Sarrazin, J. Membr. Sci., 73 (1992) 313. 22. R.-S. Juang and M.-N. Chen, Ind. Eng. Chem. Res., 35 (1996) 1935. 23. E. Piron, M. Accominotti and A. Domard, Langmuir, 13 (1997) 1653. 24. S. Despond and A. Domard, Personal communication (2002).
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Microbial recovery of copper from printed circuit boards of waste computer by Acidithiobacillus ferrooxidans K.S. Choa, M.S. Choia, J.H. Honga, D.S. Kima, H.W. Ryub, D.J. Kimc, J.S. Sohnc and K.H. Parkc a
Department of Environmental Science and Engineering, Ewha Womans University, Seoul 120-750, Korea b Department of Chemical and Environmental Engineering, Soongsil University, Seoul 156-743, Korea c Minerals Utilization & Materials Division, Korea Institute of Geoscience and Mineral Resources, Daejon 305-350, Korea Abstract The bioleaching of copper contained in the printed circuit boards (PCB) of waste computers by A. ferrooxidans was studied. The amount of copper leached from PCB shreds increased with the addition of ferrous ion and reached up to 5,190 mg.L-1 when the initial concentration of Fe2+ ion was 7 g.L-1. As the microbial leaching progressed, pale brown precipitate was observed to form in the solution. Based on the total amount of copper, both in solution and precipitate, the optimal addition of ferrous ion for the leaching of copper was around 7 g.L-1. When citric acid was not added, only about 37 wt% of the total leached copper remained dissolved; however, the amount of dissolved copper increased to greater than 80 wt% in the presence of citric acid. This fact indicates that the addition of a complexing agent (citric acid) to the bioleaching solution can raise the solubility of the leached metal ions. Keywords: microbial recovery, copper, printed circuit board, Acidithiobacillus ferrooxidans 1.
INTRODUCTION While electronic engineering has helped automate industry and commerce, it has also drastically increased electronic waste. In South Korea, it is estimated that the cumulative number of personal computers now exceeds more than 20 million units; however, only over half are currently in use. In addition, various types of electronic equipment, including personal computers, are discarded every year. However, the rate of reuse and/or recovery of valuable components contained in them is very low. For example, of the approximately one-half million personal computers discarded in South Korea last year, only 1.5% of them were reused. Considering the large number of units now being stockpiled and shortening computer life spans, personal computer waste is expected to grow. The most valuable components found in waste computers are usually contained in the PCB. Generally, one personal computer contains ca. 0.3 ~ 0.6 g of gold and 2.3 ~ 9.0 g of silver in its PCB's. In addition, there are several hundred mgs of copper included in each 631
Bioremediation Environmental Applications
gram of PCB. Therefore, it is essential - for both economic benefit and utilization of resources - to recover these components before disposal. Bioleaching is a process in which heavy metal compounds (which are included in a solid matrix), change into soluble species by the product of their metabolic regulations [1, 2]. This process is very simple since it is usually run under atmospheric temperature and pressure. Other benefits include high dissolution efficiency and simple facilities. In this research, the bioleaching of copper contained in the PCB of waste computer by A. ferrooxidans has been attempted as a fundamental study for an effective application of microbial leaching process for the removal and recovery of heavy metal components from PCB waste. 2.
MATERIALS AND METHODS
2.1 Waste PCB The PCB's used in this study were obtained from a local electronics market. They were shredded using stainless steel blades with no special pre-treatment except for a brief air-cleaning to remove dust. Shreds containing relatively high amounts of copper were collected after discarding shreds composed of plastics. The selected shreds were sieved. Shreds in the size range of -14/+20 mesh were collected and employed in bioleaching. The content of copper contained in the PCB shreds in this size range was measured to be 759.3±225.9 mg per gram of dry shreds. 2.2 Microorganism and culture The growth medium used for A. ferrooxidans (ATCC 19859) was 9K medium. This medium was composed of mineral salts [(NH4)2SO4 3.0 g.L-1, K2HPO4 0.5 g.L-1, MgSO4⋅7H2O 0.5 g·L-1, KCl 0.1 g·L-1, Ca(NO3)2 0.01 g·L-1] and 45 g.L-1 of FeSO4⋅7H2O, which served as the energy source for A. ferrooxidans. The pH of medium was adjusted at 2.0 using 1N H2SO4. A. ferrooxidans was grown in the 9K medium at 30°C for 3 ~ 4 days. After filtering the medium solution for the removal of precipitate, the cells were harvested by centrifuging at 8,000 rpm for 20 min. The harvested cells of A. ferrooxidans were suspended in the mineral salts medium again and this concentrate was used in the experiments. 2.3 Oxidation of ferrous ion by A. ferrooxidans in the leachate of PCB Waste PCB contains, in addition to copper, several other elements, e.g. Ni, Co, Cd, etc, which may dissolve in the acidic culture medium and inhibite the activity of A. ferrooxidans. To investigate the influence of chemically dissolved metals on the oxidation rate of ferrous ion by A. ferrooxidans, the following tests were carried out: Prepared PCB shreds were put into the mineral salts medium, which had been adjusted at pH 2.0. The mixture, containing 50 g.L-1 PCB shreds, was stirred at 180 rpm maintaining the temperature at 30°C for 48 h. During this period, the pH of leachate was continuously monitored and readjusted at 2.0 using a 2 N H2SO4 solution. After 48 hours, the leachate was filtered and mixed with the mineral salts medium at the volume ratio of 0, 20, 40, 60, 80, and 100%, respectively. 100 mL of each mixture was then put into a 250 mL conical flask. 45 g.L-1 of FeSO4.7H2O was then added to each mixture. After adjusting the pH of each mixture at 2.0 with an H2SO4 solution, A. ferrooxidans concentrate was inoculated until the initial cell concentration of 30 mg/L was reached. The inoculated mixtures were cultured in a shaking incubator (180 rpm) and controlled at 30°C. Changes in pH, ORP 632
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(Redox Potential), and the concentrations of Fe2+ and Fe3+ ions with time were recorded. Each experiment was repeated two times and the average was reported. 2.4 Bioleaching of copper by A. ferrooxidans 50 g.L-1 of PCB shreds was put into a 250 mL conical flask which contained 100 mL of mineral salts medium. After controlling the Fe concentration using FeSO4.7H2O, the initial pH solution was adjusted at 2.0 by H2SO4. A. ferrooxidans concentrate was inoculated to the solution to obtain an initial cell concentration of 30 mg/L, and it was cultivated in a shaking incubator (180 rpm) at 30°C. The variations of pH, ORP, and concentrations of Fe and Cu of the solution with time were examined. At the end of the bioleaching experiment, the residue remaining in the flask was recovered. After separating the reaction precipitate from the PCB shreds and drying it at 60°C, its content of copper was measured by extracting the copper from the precipitate using a 0.1 N HCl solution. All experiments were conducted twice and the averaged results are reported below. 2.5 Analysis Fe2+ and Fe3+ ions were analyzed by o-phenanthroline method with an absorption wavelength of 510 nm [3]. In addition, the contents of copper and other heavy metals in the leachate of PCB shreds were measured by an atomic absorption spectrophotometer (Plasma 40, Perkin Elmer, USA). 3.
RESULTS AND DISCUSSION
12 0 % 20 % 40 % 60 % 80 % 100%
10 8 6 (a)
4 2 0
0
12
24
36
48
Time (h)
60
72
84
Fe oxidation rate (g.L-1.h-1)
Fe2+ concentration (g.L-1)
3.1 Influence of leachate on the Fe oxidation activity of A. ferrooxidans Fig. 1 shows the influence of the leachate of waste PCB on the Fe oxidation activity of A. ferrooxidans. The oxidation rate of ferrous ion by A. ferrooxidans in the 9K medium, with no addition of leachate, was ca. 0.15 g·L-1·h-1. A slight decrease was observed in the tests that were carried out in the presence of leachate. However, the oxidation rates generally fell in the range of 0.15 ~ 0.13 g·L-1·h-1 with the addition of leachate to the medium from 20 to 100% (v/v). 0.18 0.16 0.14 0.12 0.10 0.08 (b) 0.06 0.04 0.02 0.00
0
20
40
60
80
100
Addition of leachate (%, v/v)
Figure 1. Changes of the Fe2+ concentration with time (a), and the Fe oxidation rate (b) according to the volume addition of leachate
Table 1 shows the heavy metal contents in the leachate of PCB. It should be noted that the content of copper is the greatest at about 548 mg.L-1, and those of other metals are below 2 mg.L-1. It should also be noted that the Fe oxidation activity of A. ferrooxidans 633
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was almost not demoted by these levels of concentrations of heavy metals [4-6]. Therefore, it was determined that the leachate of PCB did not seriously affect the bioleaching process by this bacterium. Table 1. Concentration of heavy metal ions in the leachate Heavy metal Concentration . (mg L-1) N.D., not detected
Cu
Zn
Fe
Cd
Ni
Co
548.00
1.61
1.13
N.D.
0.11
0.14
10 Fe (g.L-1) 2+
(a)
8
Cu concentration (g.L-1)
Fe2+ concentration (g.L-1)
3.2 Bioleaching of copper from PCB Fig. 2 shows the temporal changes of the concentrations of ferrous ion and copper ion in the leaching solution. This solution initially contained 50 g.L-1 of PCB shreds, 0 ~ 9 g.L1 of ferrous ion, and was inoculated with A. ferrooxidans. Without the inoculation of microorganism, the leached amount of copper was observed to be less than 1,000 mg/L (data not shown).
0 3 5 7 9
6 4 2 0
0
1
2
3
4
Time (d)
5
6
7
6 (b)
5 4 3 2 1 0
0
1
2
3
4
5
6
7
Time (d)
Figure 2. Time profiles of Fe2+ (a), and Cu2+ concentration in the leachate (b) for different initial amount of ferrous ion
The oxidation rates of ferrous ion estimated from the slopes of the lines in Fig. 2(a) were 1.47, 1.46, 1.44, and 1.48 g.L-1.d-1 in conditions in which the initial concentrations of ferrous ion were 3, 5, 7, and 9 g.L-1, respectively. This indicates that the oxidation rate of ferrous ion by A. ferrooxidans was virtually unaffected by the initial amount of Fe2+ ion within the experimental conditions. However, the oxidation rate obtained in these conditions was almost half of that observed in the condition of 100% leachate of PCB (0.13 g.L-1.h-1 3.14 g.L-1.d-1). The reason for this decrease may be the result of an inhibition effect by PCB shreds on the transfer of materials such as oxygen and/or carbon dioxide in solution. In other words, A. ferrooxidans is an aerobic and autotropic bacterium that uses carbon dioxide as the carbon source. In addition, if gaseous nutrients such as oxygen and carbon dioxide are not sufficiently available, the oxidation activity of ferrous ion deteriorates. Similar results regarding the deactivation of A. ferrooxidans by the inhibition of mass transfer in a slurry system have also been reported in the bioprocesses of extracting metallic components from ores, removal of iron impurities from clay minerals, and coal desulfurization [2, 7, 8]. In the absence of ferrous ion, the maximum amount of copper leached by A. ferrooxidans was ca. 2,550 mg.L-1. The amount of copper in the leachate increased with 634
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the addition of ferrous ion and reached 5,190 mg.L-1 when the initial concentration of Fe2+ ion was 7 g.L-1. However, the amount of leached copper was reduced to 3,846 mg·L-1 when the initial addition of ferrous ion was 9 g.L-1. Accordingly, it seemed that excessive addition of ferrous ion deteriorated the leaching activity of A. ferrooxidans contrarily. The leaching mechanism of copper from PCB shreds by A. ferrooxidans is similar to that of metal sulfide. For example, Fe2(SO4)3 formed by A. ferrooxidans oxidizes the elemental copper contained in PCB to cupric ion through reaction (1), which resulted in the dissolution of copper. The regeneration of FeSO4 suggests that this bioleaching process is cyclical. (1) It is also reasonable to presume that part of the copper was directly leached out chemically under the experimental conditions of the present study by the following reaction: (2) In fact, during the leaching experiment, the pH of leachate was observed to slightly but continuously increase with time close to 3.0 (data not shown), which is probably due to the hydroxide ions that formed in reaction (2). The increase in the pH of leachate can be substantiated as well by the fact that, unlike the case of metal sulfide, no sulfuric acid was formed during the oxidation process of copper (reaction (1)). By the way, the monitoring of the ORP of leachate revealed that the potential of solution did not change and was almost maintained at ca. 450 mV vs. Ag/AgCl throughout the leaching reaction (data not shown). It is also true that the addition of ferrous ion to the system furthered the dissolution of copper to some extent. Considering all this, it is reasonable to conclude that the microbial leaching of copper from PCB by A. ferrooxidans was accomplished mainly by the indirect oxidation of elemental copper through the prior oxidation of ferrous ions by A. ferrooxidans. This process was partially accompanied by a reaction in which copper was directly oxidized chemically. 3.3 Reaction precipitate As the microbial leaching progressed, pale brown precipitate was observed in the solution. Since it was considered highly possible that a part of the leached copper was included in it through physicochemical adsorption and/or substitution in the lattice structure, the content of copper in the precipitate was also analyzed to compare with the amount of copper ions in a dissolved state. Fig. 3 shows the variation of the copper contents in solution and precipitates with the initial concentration of ferrous ion. It can be seen that that in all the tests that were carried out with initial addition of Fe2+ ion, the contents of copper in the precipitates were even higher than those remaining in solution. Therefore, after the bioleaching of PCB shreds by A. ferrooxidans, subsequent treatment of reaction precipitates in a proper way should be sought in an effort to develop a more efficient method for copper recovery. However, this should not take undue time and energy since the reaction precipitate was observed to be easily extractable in its metallic component by dilute mineral acids. Based on the total amount of copper both in solution and precipitate, the optimal addition of ferrous ion for the leaching of copper was around 7 g.L-1, and the recovery efficiency of copper compared to its initial amount in the PCB shreds was 24%.
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12
Cu content (g.L-1)
10
Cu content in solution Cu content in precipitate Total Cu content
8 6 4 2 0
0
3
5
7
9
Fe concentration (g.L-1) 2+
Figure 3. Variation of the contents of copper in solution and precipitate with the initial addition of ferrous ion 3.4 Effect of complexing agent As mentioned previously, a large portion of the leached copper ions was found to be fixed again in the precipitate. To increase the solubility of copper and improve the efficiency of this bioleaching process, addition of a complexing agent was attempted. Table 2 shows the distributions of leached copper in solution and precipitate when the bioleaching was conducted under the same conditions as those of Fig. 3 in the absence and presence of 1 g.L-1 of citric acid. As can be seen, when citric acid was not added, only about 37 wt% of the total leached copper remained dissolved; however, it increased to greater than 80 wt% in the existence of citric acid. This indicates that the addition of a complexing agent like citric acid to the bioleaching solution can raise the solubility of the leached metal ions, which enables subsequent recovery processes, including solvent extraction, to be carried out more efficiently. Table 2. Effect of the addition of citric acid on the state of leached copper ions
4
Condition
Cu in solution (%)
Cu in precipitate (%)
Without citric acid
37.0 ± 3.1
63.0 ± 5.1
With citric acid
81.6 ± 1.8
18.4 ± 1.6
CONCLUSIONS Present fundamental studies for the microbial leaching of copper from PCB's generated from waste computers have demonstrated the possibility that the valuable metallic components contained in wastes can be recovered effectively through a bioleaching process. Hereafter, more systematic investigations will be needed regarding the modification and optimization of the leaching process, the development of an economical method for the recovery of metals from the leachate, and the utilization of the leaching residues in a profitable way.
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REFERENCES
1. W. Sand, T. Gehrke, P.G. Jozsa, and A. Schippers, Hydrometallurgy, 59 (2001) 159. 2. K. Bosecker, FEMS Microbiol. Rev., 20 (1997) 591. 3. N.H. Furman, Standard Methods of Chemical Analysis, R.E. Krieger Publishing Co., New York, 1975. 4. J. Magnin, F. Baillet, A. Boyer, R. Zlatev, M. Luca, A. Cheruy, and P. Ozil, Can. J. Chem. Eng., 76 (1998) 978. 5. H. Huber and K.O. Stetter, Appl. Environ. Microbiol., 56 (1990) 315. 6. L. Harahuc, H.M. Lizama, and I. Suzuki, Appl. Envron. Microbiol., 66 (2000) 1031. 7. H.W. Ryu, K.S. CHO, Y.K. Chang, S.D. Kim, and T. Mori, J. Ferment. Bioeng., 80 (1995) 46. 8. G.F. Andrews, M. Darroch, and T. Hansson, Biotechnol. Bioeng., 32 (1988) 813.
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Oxidation of iron, sulfur and arsenic in mine waters and mine wastes: an important role for novel Thiomonas spp. Kris Couplanda, Fabienne Battaglia-Brunetb, Kevin B. Hallberga, Marie-Christine Dictorb, Francis Garridob and D. Barrie Johnsona a
School of Biological Sciences, University of Wales, Bangor, LL57 2UW, U.K. b Biotechnology Group, Environment & Process Division, BRGM, 3 avenue C. Guillemin, B.P. 6009, 45060 Orleans cedex 2, France
Abstract Bacteria belonging to the genus Thiomonas (Tm.) are usually considered to be neutrophilic, except for the acidophile Tm. cuprina. These microbes are capable of oxidizing elemental sulfur and reduced inorganic sulfur compounds (RISCs), such as thiosulfate and, as such, were once classified as Thiobacillus spp. Recently, a number of isolates that are related to known Thiomonas species, as determined by 16S rRNA gene sequence homology and physiological characteristics, have been obtained from various iron-containing acidic and arsenic-contaminated neutral mine sites from around Europe. Several of the isolates were obtained by virtue of their ability to catalyze the oxidation of Fe (II) and one isolate was obtained following growth with As (III) as electron donor. To determine if these two phenotypes are restricted only to bacteria isolated from iron- or arsenic-containing sites, we have initiated a study looking into the ability of each of the iron-oxidizing isolates to catalyze the oxidation of As (III), and of the As (III)-oxidizer to oxidize Fe (II). In addition to the new isolates, the type strains of Tm. perometabolis and Tm. intermedia have been included in this study. Preliminary results indicate that the ability to oxidize Fe (II) and As (III) is widespread amongst this phylogenetically related group of microorganisms. Also, data indicate that the ability to grow at relatively low pH (as low as pH 3) is also widespread amongst these bacteria. Aside from the evolutionary implications of iron and arsenic oxidation, these results indicate that there exists a variety of microorganisms that can be applied to remediate iron- and arsenic-containing mine waters of varying physico-chemical properties. Keywords: acid mine drainage, arsenic, moderate acidophiles, Thiomonas 1.
INTRODUCTION The role of bacteria and archaea in mediating dissimilatory transformations of iron and sulfur in acidic, metal-rich waters such as acid mine drainage is well documented. Though much of the published research has focused on Acidithiobacillus ferrooxidans, other mesophiles such as Leptospirillum ferrooxidans, "Sulfobacillus montserratensis" and the archaeon Ferroplasma also oxidize ferrous iron, Fe (II), to ferric, Fe (III). Some ironoxidizers can also oxidize elemental sulfur and reduced inorganic sulfur compounds 639
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(RISCs) to sulfate, while other sulfur-oxidizers (most notably Acidithiobacillus thiooxidans) are incapable of oxidizing iron. The genus Thiomonas (Tm.) was first proposed by Moreira and Amils [1] for some bacteria that were classified, at that time, as Thiobacillus spp. The justification for the reclassification was on the basis of phylogeny (based on 16S rRNA gene sequences) and physiological characteristics (mixotrophic growth of the four species studied: Tm. intermedia, Tm. perometabolis, Tm. thermosulfata and Tm. cuprina). Of these, only Tm. cuprina was classed as an acidophile, though Shooner et al. [2] reported that Tm. thermosulfata could grow at pH 4.3. Whereas all Thiomonas spp. can grow autotrophically with sulfur or RISCs as a source of energy, none has been reported to oxidize Fe (II). Novel Thiomonas-like bacteria were isolated from water draining a deep coal mine in south Wales [3]. The mine water at source was pH 6.8, but fell to 3.0 as a result of ironoxidation and hydrolysis. No extremely acidophilic iron-oxidizers (Acidithiobacillus, Leptospirillum, etc.) were detected in this mine water, but colonies of ferric iron-encrusted bacteria grew on pH 4 ferrous iron/thiosulfate solid medium inoculated with enrichment cultures from the AMD. The novel Fe (II)- and RISC-oxidizing Thiomonas isolates were classified as "moderately acidophilic" since they grew at pH 3 but not at pH 2. Arsenic is frequently found in relatively high concentrations in mineral leaching environments, due to the oxidative dissolution of arsenopyrite (FeAsS) and arsenical pyrite. Arsenic exists mainly in two oxidation states, As (III) and As (V), of which As (III) is more toxic. While a number of bacteria that reduce As (V) to As (III) have been described, relatively little is known about As (III)-oxidizing microorganisms. Santini et al. [4] isolated nine strains of As (III)-oxidizing bacteria (four autotrophic, five heterotrophic) from gold mine sites in Australia. The chemoautotrophs were all α-Proteobacteria, while the heterotrophic isolates were all β-Proteobacteria. Elsewhere, As (III)-oxidizing bacteria have been implicated in the immobilization of arsenic (as Fe (III)-As (V) precipitates) in acidic mine waters draining the Carnoules mine in France [5, 6]. The abiotic catalysis of As (III) by iron (III) is a thermodynamically viable reaction, though this may require a catalyst, such as the surface of pyrite [7]. In this paper, several novel strains of moderately acidophilic Thiomonas isolates from AMD-impacted or arsenic impacted waters are described, all of which have been found to oxidize iron (II), RISCs and As (III). 2.
MATERIALS AND METHODS
2.1 Bacterial strains The moderate acidophiles used in this study were isolated from mine sites in various European locations (Table 1). Several quite different media types were utilized and include R2 agar [8], originally developed for growth of neutrophiles, ferrous iron/thiosulfate/tryptone soya broth overlay solid medium ("FeTo", pH ~ 4.5), and yeast extract overlay solid medium ("Yeo", pH ~ 3.0). The latter two media are described in more detail in [9]. Further information on isolates Ynys1 and CAsO1-b6 can be found in [3] and [5], respectively. Type strains of Thiomonas were obtained from various culture collections, and included Tm. intermedia ATCC 15466, Tm. perometabolis ATCC 23370, and Tm. cuprina DSM 5495. Media recommended by the culture collections were used to grow the type strains.
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Table 1. Source and method of isolation of the moderately acidophilic strains used in this study Isolate Ynys1 WJ68
NO115
PK44 CAsO1-b6
Origin AMD (pH 6.8), Ynysarwed coal mine, south Wales AMD (pH 5.5) in a constructed wetland at the Wheal Jane tin mine, Cornwall, England Seepage from waste dumps (pH 3.5), Killingdal copper mine, Roeros, Norway Biofilm from an adit draining disused copper mine (pH 2.7), north Wales Arsenic-containing mine spoil (pH 6), disused Cheni gold mine, France
Method of Isolation Direct isolation on FeTo Direct isolation on FeTo
Direct isolation on R2 agar
Direct isolation on YEo Enrichment in liquid medium containing As (III) and isolation on solid medium with lactate and As (III)
2.2 Iron (II) Oxidation by isolate WJ68 WJ68 was grown in 5 mM ferrous sulfate/5 mM sodium thiosulfate/0.025% (w/v) tryptone soya liquid medium ("FeT", initial pH 4.5) in a shake flask at 20°C. Following observable oxidation of iron (change of culture medium from colorless to rust color), the bacteria were inoculated, in triplicate, into fresh FeT and Fe (II) concentrations in culture medium were periodically determined with ferrozine [10]. 2.3 Oxidation of As (ΙΙΙ) by NO115 Strain NO115 was routinely grown in selective liquid medium [5] supplemented with 0.1 g/l yeast extract, 1 mM sodium thiosulfate and 100 mg/l As (III) (Yeast extract Thiosulfate Arsenite Medium - YTAM) in shake flasks at 25°C. In order to test the influence of As (III) oxidation on bacterial growth, three variants of the above culture medium were prepared: (1) YAM, yeast extract and As (III); (2) YM, yeast extract only; and (3) AM, As (III) only. Respective media in shake flasks were inoculated with NO115 and incubated at 25°C. As (V) and total arsenic were analyzed by Atomic Absorption Spectrophotometry (Varian SpectrAA 300). As (III) and As (V) were separated using PDC/MIBK [5]. Total bacteria were enumerated under an optical light microscope using a Thoma cell. 2.4 Growth of moderate acidophiles on other substrates The moderately acidophilic isolates were subcultured in a minimal liquid medium (MMa) containing the following (g/l); KH2PO4 (2.5), (NH4)2SO4 (1), MgSO4.7H2O (0.1), FeSO4.7H2O (0.005), NaOH (0.55), nitrilotriacetic acid (0.25), trace elements, yeast extract (0.1), Na2S2O3 (to final concentration of 1 mM) and glucose (to final concentration of 5 mM). All components were mixed and adjusted to pH 6.0 before sterilization by autoclaving for 20 minutes at 120°C. Isolates were also tested for growth in a range of variations of the above medium: without sodium thiosulfate; without glucose; 10mM lactate in place of glucose; and 5mM succinate in place of glucose. Isolates were also grown on a solid form of the lactate-supplemented medium made with 5 g/l agarose (Sigma type 1, low EEO). All cultures were incubated at 20°C.
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2.5 Phylogeny of moderate acidophiles and known Thiomonas species Bacterial 16S rRNA genes were amplified using protocols described elsewhere [11]. Purified PCR products were sequenced using a BeckmanCoulter Dye Terminator Cycle Sequencing kit and a CEQ8000 Genetic Analysis System (Beckman Coulter, U.K.). The resulting gene sequences were compared with those available in GenBank using BLAST [12]. A phylogenetic tree was constructed using the neighbor joining function of ClustalX [13], with gaps excluded from the analysis.
Figure 1. Phylogenetic tree showing the relationship of the moderately acidophilic isolates used in this study to each other, and to other selected β-Proteobacteria. The GenBank accession numbers of the sequences not determined in this study are: Tm. cuprina, U67162; Thiomonas sp. CAsO1-b6, AF460990; Tm. thermosulfata, U27839; Ynys1, AF387302; Gallionella ferruginea, L07897; and Leptothrix discophora, L33975. Scale bar represents 0.05 inferred nucleotide substitutions per site for the horizontal branch lengths between any two microbes 3.
RESULTS
3.1 Origin and isolation of Thiomonas isolates A group of microbes were isolated using different media from various miningimpacted sites in Europe (Table 1). Some were initially identified by their apparent ability to oxidize Fe (II) (Ynys1 and WJ68), one for its ability to oxidize As (III), and two were isolated as heterotrophs. Phylogenetic analysis revealed that all of these microbes were related to bacteria of the genus Thiomonas (Figure 1). The 16S rRNA genes from the type species of Thiomonas, Tm. intermedia, as well as from Tm. perometabolis were also sequenced. The latter two had genes of >99% identity, and also shared similar homology to the genes from WJ68 and Ynys1. The gene from a third Thiomonas strain, Tm. thermosulfata, was ~96% identical to this group of microbes. These microbes formed a group of Thiomonas isolates, designated as Group 1. A second group of Thiomonas-like isolates was also detected (Figure 1), and these microbes have similar 16S rRNA genes (~97% identity) with each other but only about ~92% to those of Group 1 thiomonads. The gene sequence available in public databases from Tm. cuprina is only about 92% identical to the others, and therefore lies outside of either Group.
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3.2 Oxidation of iron by Thiomonas strain WJ68 As isolate WJ68 was found to be highly similar to the iron-oxidizer Ynys1, the ability of WJ68 to oxidize Fe (II) in liquid media was investigated. When WJ68 was inoculated into FeT liquid medium, approximately 12% of Fe (II) was oxidized to ferric iron over a period of about 12 hours (Figure 2). During the course of iron oxidation by WJ68, the bacteria were highly motile and cell numbers increased (data not shown). During further incubation for another 24 hours, an additional 6% of the Fe (II) was oxidized (not shown). In abiotic controls, no noticeable decrease in Fe (II) was observed over a 5-day period.
Figure 2. Mean oxidation of Fe (II) in FeT liquid medium by triplicate cultures of WJ68. The error bars show the standard deviation for each time point
During the oxidation of iron, the pH of the culture medium decreased to 3 within 8 hours, with no further pH change ocurring between 8 and 12 hours. This decrease is to be expected from the hydrolysis of the Fe (III) produced: Fe3+ + 3H2O → Fe(OH)3 + 3H+ In cultures of WJ68 where the pH of the medium was initially set a 3.0, no oxidation of iron occurred, nor was any growth of the microbes observed. It thus appears that pH 3 is the lower limit for growth of this moderate acidophile. An upper pH limit for growth was not determined as abiotic oxidation of Fe (II) occurs at an appreciable rate above pH 5 (the starting pH of the culture shown in Figure 2). 3.3 Oxidation of As (ΙΙΙ) by Thiomonas sp. NO115 The close phylogenetic relationship between NO115 and the As (III)-oxidizer CAsO1-b6 prompted investigation of As (ΙΙΙ) oxidation by NO115. Oxidation was complete within 48h in both YAM and AM cultures (Figure 3A), and was more rapid in the presence of yeast extract (YAM). Bacterial growth was very poor in the yeast extractfree AM medium (Figure 3B). Growth of NO115 was stimulated by As (III) (Figure 3B); bacterial concentrations were greater in As (III)-containing medium YAM than in As (III)free medium (YM). These results suggest that energy from As (ΙΙΙ) oxidation supports the growth of NO115. Reproducible growth of NO115 with only As (III) as energy source has not yet been confirmed.
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Figure 3. Influence of arsenic on NO115 growth: (A) evolution of As (V) and (B) bacterial growth. YAM: medium with yeast extract and As (III); YM medium with only yeast extract; AM medium with only As (III). Error bars represent standard error of means (3 replicates) 3.4 Growth of all isolates on various media types Further comparison of the Thiomonas isolates included examination of their ability to grow on with various substrates (Table 2). While all strains tested grew in thiosulfate/yeast extract liquid medium containing glucose or succinate, the most successful growth medium, in terms of cell numbers, of all strains tested contained lactate. Strains formed cream-colored, domed colonies on the agarose-gelled form of this medium. Table 2. Growth of Thiomonas bacteria in liquid medium (MMa) containing 0.1 g/l yeast extract (YE), 5 mM glucose and 1 mM sodium thiosulfate (thio), or the variants below. The relative final cell number is indicated by the number of asterisks (*). Iron or arsenic oxidation is indicated by Y. Growth inhibition by arsenic is indicated by I, and n.t. means not tested Isolate Tm. intermedius Tm. perometabolis Ynys 1 WJ68 NO115 PK44 CAsO1-b6
Growth in medium containing Oxidation of YE, thio, YE, YE, thio, YE, thio, As YE, thio Fe (II) glucose glucose lactate succinate (III) ** * ** *** *** Y Y ** * ** *** *** Y I ** * ** *** *** Y n.t. ** * *** *** *** Y Y ** * ** *** *** Y Y * * ** *** *** Y n.t. ** * ** *** *** Y Y
Consistent with the phylogenetic classification, heterotrophic growth of all isolates with glucose was poor in the absence of thiosulfate (Table 2). Evidence that these isolates could use thiosulfate as a source of energy, and not simply as a growth factor (i.e. a source of reduced sulfur), came from experiments in which the bacteria were incubated in medium MMa containing additional thiosulfate (added to 5 and 10 mM). For all strains, cell numbers were found to increase in proportion to the amount of available thiosulfate. 644
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Further evidence of thiosulfate utilization was provided by the decrease of culture media pH to <3 (from pH 6). All of the Thiomonas strains were able to grow on FeTo plates, and catalyzed the oxidation of Fe (II), turning the medium red/orange. Like those in Group 2, all Thiomonas Group 1 bacteria tested were able to oxidize As (III) (Table 2) with the exception of Tm. perometabolis, which appeared to be inhibited by arsenic. 4.
DISCUSSION Bacteria belonging to the genus Thiomonas are capable of oxidizing sulfur and RISCs, and are also able to grow mixotrophically [1] or heterotrophically [2]. Most of the Thiomonas species have relatively high pH optima for growth (5-7) and, except for Tm. cuprina, are not usually considered to be acidophilic. With the exception of Tm. cuprina, Thiomonas species have not been considered important in mining environments. Recently, however, Thiomonas-like Fe (II)-oxidizing (Ynys1 [3]) and (As III)oxidizing (CAsO1-b6 [5]) bacteria have been isolated from mine water-impacted streams. Due to their ability to grow at pH 3 to 5, Ynys1 was considered to be a moderate acidophile. Analysis based on 16S rRNA gene sequences revealed that Ynys1 was closely related to Tm. thermosulfata, and that the nearest relative of the As (III)-oxidizing strain b6 was Ynys1. This information prompted the current comparative study of these two isolates with other, more recent, Thiomonas-like isolates from various European mineimpacted sites and with the type strains of the Thiomonas species Tm. intermedia and Tm. perometabolis. Unfortunately, Tm. thermosulfata is not available in any public culture collection and Tm. cuprina did not grow during this study. Even though the more recent isolates were originally obtained on a variety of media, partial 16S rRNA gene sequence homology showed that they were other all related to Thiomonas species. For completeness of the phylogeny, the 16S rRNA genes of the type species of Thiomonas, Tm. intermedia, and also of the type strain of Tm. perometabolis were sequenced (only the middle portion of the latter gene sequence was available in public databases [14]). The latter two microbes shared gene homology of >99%, indicating that they represent the same species. This is in keeping with the DNA:DNA homology between these two of greater that 70% [15]. The 16S rRNA genes from two of the isolates, Ynys1 and WJ68, from mine sites had nearly 100% identity with the two culture collection strains, indicating that they too are probably the same species (Tm. intermedia has precedence over Tm. perometabolis). These strains form what we have designated as “Thiomonas Group 1”, which includes Tm. thermosulfata (~96% identity to the others). On the other hand, two of the isolates (PK44 and NO115) were more closely related to the As (III)-oxidizer CAsO1-b6 (~97%) and form “Thiomonas Group 2”, which shares only about 92% identity with microbes of Group 1. Tm. cuprina shares only about 92% identity to Group 2 microbes (and even less to Group 1 microbes), and therefore is excluded from either group until further study can be carried out on this microbe. Based on 16S rRNA gene sequence analysis, further studies are required to correctly address the taxonomy of these phylogenetically related microbes. The various Thiomonas strains studied here shared a number of physiological characteristics. Successful growth of all strains in liquid medium was only observed in the presence of thiosulfate, and increased concentrations of thiosulfate resulted in higher growth yields of all the strains. Growth of all strains was enhanced in the presence of lactate or succinate. Also, all of the isolates and the type strains of both classified species that grew appeared to oxidize Fe (II), as evidenced by color change on Fe (II)-containing 645
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solid medium (and supported, in the case of isolates WJ68 and Ynys1, by data from liquid cultures). That the latter property has not previously been ascribed to Thiomonas species may be due to the fact that they have generally been grown in media of higher pH than those used in the present study. Above pH 4.5, the abiotic oxidation of Fe (II) is very rapid, making conclusions about biological iron oxidation difficult to draw. Lastly, each of the Thiomonas strains tested here were able to oxidize As (III) to As (V), except Tm. perometabolis whose growth was inhibited by arsenic. This oxidation reaction has been shown to yield energy for growth for NO115, as has also been shown for CAsO1-b6 [5]. The fact that these microbes can catalyze the oxidation of both Fe (II) and As (III) is of major interest for the remediation of mine drainage. As these microbes are able to grow over a wide range of pH, they are suitable candidates for the bioremediation of mine waters, which tend to vary significantly in physico-chemical parameters [16] but which often contain elevated concentrations of dissolved iron and arsenic. ACKNOWLEDGEMENTS K.C. is grateful to the NERC (U.K.) and Rio Tinto Technology for provision of a research studentship (ref. NER/S/C/2001/06450). K.B.H. and D.B.J. thank the DTI (U.K.; ref. BTL/20/71) for partial support of this work. REFERENCES
1. D. Moreira and R. Amils, Int. J. Syst. Bacteriol. 47 (1997) 522. 2. F. Shooner, J. Bousquet and R. D. Tyagi, Int. J. Syst. Bacteriol. 46 (1996) 409. 3. F. Dennison, A. M. Sen, K. B. Hallberg and D. B. Johnson, Biohydrometallurgy: Fundamentals, Technology and Sustainable Development, V.S.T. Ciminelli and O. Garcia Jr. (eds.), Elsevier, Amsterdam, (2001) 493. 4. J. M. Santini, L. I. Sly, A. M. Wen, D. Comrie, P. De Wulf-Durand and J. M. Macy, Geomicrobiol. J. 19 (2002) 67. 5. F. Battaglia-Brunet, M. C. Dictor, F. Garrido, C. Crouzet, D. Morin, K. Dekeyser, M. Clarens and P. Baranger, J. Appl. Microbiol. 93 (2002) 656. 6. M. Leblanc, B. Achard, D. B. Othman, J. M. Luck, J. BertrandSarfati and J. C. Personne, Appl. Geochem. 11 (1996) 541. 7. J. Barrett, D. K. Ewart, M. N. Hughes and R. K. Poole, FEMS Microbiol. Rev. 11 (1993) 57. 8. D. J. Reasoner and E. E. Geldreich, Appl. Environ. Microbiol. 49 (1985) 1. 9. K. B. Hallberg and D. B. Johnson, Hydrometallurgy (2003) in press. 10. D. R. Lovley and E. J. P. Phillips, Appl. Environ. Microbiol. 53 (1987) 1536. 11. N. Okibe, M. Gericke, K. B. Hallberg and D. B. Johnson, Appl. Environ. Microbiol. 69 (2003) 1936. 12. S. F. Altschul, T. L. Madden, A. A. Schaffer, J. Zhang, Z. Zhang, W. Miller and D. J. Lipman, Nucleic Acids Res. 25 (1997) 3389. 13. J. D. Thompson, T. J. Gibson, F. Plewniak, F. Jeanmougin and D. G. Higgins, Nucleic Acids Res. 25 (1997) 4876. 14. D. J. Lane, A. P. Harrison Jr., D. Stahl, B. Pace, S. J. Giovanni, G. J. Olsen and N. R. Pace, J. Bacteriol. 174 (1992) 269. 15. A. P. Harrison Jr., Int. J. Syst. Bacteriol. 33 (1983) 211. 16. D. Banks, P. L. Younger, R.-T. Arnesen, E. R. Iversen and S. B. Banks, Environ. Geol. 32 (1997) 157.
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Oxidation of metallic copper by Acidothiobacillus Ferrooxidans K. Lilova and D. Karamanev Department of Chemical and Biochemical Engineering, The University of Western Ontario, London, Ontario N6A 5B9, Canada Abstract This is a study on the kinetic aspects of the oxidation of zero-valence copper by immobilized Acidothiobacillus ferrooxidans. Two different mechanisms of oxidation were considered: the direct and the indirect one. An original bioreactor was used in order to grow A. ferrooxidans in an iron sulphate-free media, which was required for the study of the direct mechanism. It was shown that the direct oxidation exists, but it is too slow. 1.
INTRODUCTION The bacterium Acidothiobacillus ferrooxidans is a chemolithotrophic organism that can use different inorganic reactions as an energy source [1]. These reactions include the oxidation of ferrous ions: 2FeSO4 + H2SO4 + ½O2 = Fe2(SO4)3 + H2O (1) the oxidation of sulfur:
2S° + 3O2 + 2H2O = 2H2SO4 (2) or chemical compounds in which sulfur is at a valency lower than +6: 4FeAsS + 13O2 + 6H2O = 4H3AsO4 + 4FeSO4 (3) The above shows that A. ferrooxidans can oxidize inorganic substrates in both liquid (aqueous solution) and solid (crystal) forms. While the mechanism of aqueous ferrous iron oxidation has been relatively well understood [2,3], the oxidation mechanism of crystallic compounds such as metal sulfides, has been a subject of intensive debates. Two different mechanisms of sulfide oxidation have been proposed [1]: direct one, based (in the case of pyrite) on the following reaction: 4FeS2 + 14O2 + 4H2O = 4FeSO4 + 4H2SO4 (4) where the sulfides are oxidized by the direct action of the immobilized bacterium. The indirect mechanism postulates that the oxidation of sulfide is a pure chemical process: 7Fe2(SO4)3 + FeS2 + 8H2O = 15FeSO4 + 8H2SO4 (5) The products of reaction (5) are then oxidized by A. ferrooxidans: sulfur - to sulfuric acid according to reaction (2) and ferrous sulfate - to ferric sulfate according to reaction (1). The produced ferric sulfate is then reused for the chemical oxidation of pyrite (reaction (5) which closes the cycle of iron ions oxidation-reduction. Some fundamental 647
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bioleaching kinetics works, published in the last several years [4-6], strongly tipped the balance towards the indirect mechanisms. In addition to the above-mentioned substrates, it is also known that A. ferrooxidans is able to use oxidize zero-valence metals such as copper [7], and alloys such as constantane and even stainless steel [8]. This type of oxidation is usually unwanted because it results in biocorrosion. However, some new applications of metal oxidation, to be used for beneficial purposes, are emerging. These include the recycling of electronic components, biomachining and some possible mining applications.
Figure 1. Direct biological oxidation of copper
Since in the case of metal oxidation, the substrate is in a solid, insoluble form, there is also a possibility for both direct and indirect oxidation. The direct oxidation (Fig. 1) would be based on the following reaction, when the metal is copper: (6) Cu° + 2H+ + ½O2 = Cu2+ + H2O The indirect metal biooxidation (Fig. 2), similarly to the case of sulfide oxidation, involves chemical reaction with ferric ions: Cu° + Fe2(SO4)3 = CuSO4 + 2FeSO4 (7) followed by reoxidation of FeSO4 (1). In addition, some chemical oxidation of copper by sulfuric acid in the presence of oxygen can be expected.
Figure 2. Indirect biological oxidation of copper 648
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The main goal of this work was to study the kinetics of the zero-valence copper oxidation by A. ferrooxidans. In particular, the importance of the direct oxidation will be estimated. 2.
MATERIALS AND METHODS
2.1 Bioreactors 2.1.1 Bioreactor for the cultivation of A. ferrooxidans on copper sulfide An immobilized ore bioreactor was used for the cultivation of A. ferrooxidans on copper sulfide in the absence of iron salts. The bioreactor was described in details earlier [9]. It was a vertical cylindrical vessel with a diameter of 4.5 cm and a total volume of 560 mL. The cylinder was separated vertically into two semicylindrical sections by means of a rectangular piece of non-woven polyester textile. It had a porosity of 99.5%, fibre diameter of 50 microns and a thickness of 2 cm. The textile was hold in place by two pieces of stainless-steel mesh, forming a sandwich-like structure. The air was sparged, using a membrane pump, at the bottom of one of the sections by means of a perforated tube with a flow rate of 60 L/h. The liquid substrate was delivered by a peristaltic pump with flow rates between 50 and 250 mL/h. 2.1.2 Bioreactor for the copper oxidation This bioreactor had a similar geometry as the one used for CuS oxidation. The main difference was in the vertical wall. In the experiment with copper oxidation, the vertical wall in the bioreactor was formed from copper foil, which was oxidized by microorganisms. The circulating liquid flow in the reactor provided a good oxygen supply, combined with low shear stress and well defined velocity profiles. The experiments were carried out in two bioreactors, one for study of the corrosion due to the bacterial oxidation and a sterile (control) one, operating under the same conditions except for the absence of bacteria. All the bioreactors were operated at ambient temperature (21°C) and a pH of 2.0. 2.1.3 Materials The mineral salts solution was prepared according to the Silverman and Lundgren 9K media [10]. Ferrous sulfate was not added to the solution, which contained: ammonium sulfate 3 g, potassium phosphate (K2HPO4) 0.5 g, magnesium sulfate 0.5 g, potassium chloride 0.1 g, copper sulphide 10 g, deionized water 1 L, sulfuric acid - to obtain pH of 2.5. All the chemicals were analytical grade and were purchased from VWR Inc. The copper used was in the shape of foil with thickness of 0.5 mm and a purity of 99.9%. 2.1.4 Analytical methods The concentrations of ferrous and ferric ions in aqueous solutions were determined by a spectrophotometric method using sulphosalycilic acid [11]. The spectrophotometer used was Cary-50 (Varian Inc.). The concentrations of total iron and copper in liquid were determined by atomic absorption (Varian SpectrAA spectrometer). The microscopic pictures of copper, oxidized by microbial cells, were taken by a scanning electron microscope Hitachi. The elemental analysis of the surface was performed by energy dispersive X-ray analysis.
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3.
RESULTS AND DISCUSSION
3.1 Study of direct mechanism of copper biooxidation In order to study the process of direct microbial oxidation of copper, it was first necessary to prepare a microbial culture, growing in the absence of aqueous iron ions. It was decided to transfer an A. ferrooxidans culture from a ferrous sulfate growth solution to pure, chemically synthesized, copper sulfide. In order to eliminate iron from the growth media, a system was developed, in which it was possible to continuously wash the liquid phase from the bioreactor, containing microorganisms and copper sulfide, without washing out the microorganisms and CuS. The bioreactor used was based on the recently proposed concept of ore immobilization [12]. It was a modified airlift bioreactor, in which the non-permeable vertical wall was replaced by a highly porous, liquid-permeable fibrous non-woven textile (Fig. 3). The porous wall separates vertically the bioreactor into two sections, one of which is aerated, while the other one is not. Because of the difference in the hydrostatic pressure in the aerated and non-aerated sections, the liquid flows in both horizontal and vertical directions. The vertical flow is directed upwards in the aerated section and downwards in the non-aerated one. In the horizontal direction, liquid flows from the non-aerated towards the aerated section through the porous wall (Fig. 3).
Figure 3. Immobilized ore bioreactor
Once the solid particles (copper sulfide) were introduced to the bioreactor in the form of water slurry, they quickly become entrapped into the pores of the wall. After inoculation of the bioreactor with A. ferrooxidans, the cells attach spontaneously to the surface of sulfide crystals, forming a biofilm. This two-level immobilization structure (cells fixed on sulfide; sulfide fixed in the pores of textile) prevents the washing-out of cells and particles from the bioreactor, when liquid is continuously fed and withdrawn. The bioreactor was filled with Silverman and Lundgren 9K media containing no FeSO4. After aeration was started, 5 g of CuS were added, which made liquid black and non-transparent. It took approx. 30 min for the liquid to become transparent again, which indicated the almost complete entrapment of solids in the porous wall. At that point, 10% v/v inoculate, containing A. ferrooxidans and some dissolved iron sulfate, was added to the bioreactor. After another 24 hours, a continuous pumping of 9K media with a pH of 2.5 into the bioreactor was started. The flow rate was 100 mL/h. Close to constant CuS 650
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oxidation rate was obtained 5 days after the start of the experiment. Under these conditions, the rate of copper sulfide oxidation was 1.4 mg/L.h. There was a visible formation of biofilm in the bioreactor. After the steady-state oxidation of CuS was reached, the microbial culture was transferred from the immobilized ore bioreactor, containing copper sulfide, to the bioreactor with copper foil. Steady state conditions were reached when the dissolved copper concentration in the liquid leaving the bioreactor became constant. A control reactor was set up, working under the same conditions as the copper bioreactor, with the only difference being the absence of A. ferrooxidans in the former. Samples of the copper foil were taken periodically from both the bioreactor and the control reactor. A scanning electron microscope was used to analyze the surface of copper foil after biooxidation in the bioreactor and after chemical oxidation in the control reactor. The scanning electron micrographs of copper with A. ferrooxidans and without microorganisms for the same period of time are shown in Figs. 4 and 5. The difference between the two figures is obvious. The flower-like formations on the surface of the biologically-treated copper were found to contain carbon, phosphorus, oxygen, and presumably hydrogen. It was calculated that the ratio of the molar fractions of different elements corresponds to copper phosphate plus biomass. It was assumed that each microbial cell, attached to the surface of copper, oxidizes Cu to Cu2+, which is followed by the formation of copper phopsphate.
Figure 4. Scanning electron micrograms of copper treated by A. ferrooxidans
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Figure 5. Scanning electron microgram of copper kept under sterile conditions 4.
CONCLUSIONS The results of this study show that A. ferrooxidans can oxidize zero-valence copper via direct mechanism, in the absence of intermediate oxidants such as iron ions. At least some of the products of the reaction form insoluble, flower-like crystal structure, covering the copper surface. Our preliminary data indicate that the process of direct copper biooxidation is quite slow, much slower than the indirect copper oxidation. These results will be important for different processes, involving biooxidation of metallic copper.
REFERENCES
1. A.Torma, Adv. Biochem. Eng., 6 (1977) 1. 2. M. Nemati, S. T. L. Harrison, G. S. Hansford and C. Webb, Biochem. Eng. J., 1 (1998) 171. 3. W. J. Ingledew, Biochim. Biophys. Acta, 683 (1982) 89. 4. W. Sand, T. Gehrke, P.-G. Jozsa and A. Schippers, Hydrometallurgy, 59 (2001)159. 5. H. Tributsch, Hydrometallurgy, 59 (2001) 177. 6. G. S. Hansford and T. Vargas, Proc. Metall., 9A (1999) 13. 7. Y. Uno, T. Kaneeda and S. Yokomizo, JSME Int. J., 39 (1996) 837. 8. Ignatiadis and D. Morin, Proc. Metallurgy, 11A (2001) 95. 9. N. Chong, D. Karamanev and A. Margaritis, Biotechnol. Bioeng., 80 (2002) 349. 10. M. P. Silverman and D. G. Lundgren, J. Bact., 77 (1959) 642. 11. D. Karamanev, L. Nikolov and V. Mamatarkova, Miner. Eng., 15 (2002) 341. 12. D. Karamanev, C. Chavarie and R. Samson, Biotechnol. Bioeng., 57 (1998) 471.
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Process monitoring of biodesulfurization of high sulfur coal in packed columns using molecular ecology methods F. Gómez1, J. Cara2, M.T. Carballo2, A. Morán2, R. Amils1,3 and F.J. García Frutos4* 1
Centro de Astrobiología (CSIC-INTA), Torrejón de Ardoz, 28850 Madrid Universidad de León; Ingeniería Química. Avda. de Portugal, 41, 24071 León, Spain 3 Centro de Biología Molecular, U. Autónoma de Madrid, Cantoblanco, 28049 Madrid 4 CIEMAT, Fossil Fuels Department, Avda. Complutense 22, 28040 Madrid, Spain 2
Abstract Biodesulfurization of inorganic pyritic sulphur in coals has been extensively studied for years, and large-scale biodepyritization has been described in heap- and in slurryleaching processes. Although reaction rates are much faster in slurry reactors with higher efficiency of pyrite elimination achieved in shorter periods of time, there is a controversy about its economic feasibility. On the other hand, the economics of coal processing using heap-leaching may be competitive in cost and operation, even though it requires longer residence times with lower efficiencies than those obtained with slurry reactors. Given that heap bioleaching processes take place in non-sterile substrates with complex native microbial populations, it would be convenient to have the means to control the microbial ecology of the system thus optimising the process to obtain high desulfurization efficiencies. The introduction of molecular ecology techniques (FISH, DGGE) to the direct identification and quantification of microorganisms in natural environments opened the possibility of using these methodologies to control complex biotechnological processes like bioleaching. The development of methodologies for the molecular ecology characterisation of coal biodesulfurization processes allowed the microbial population changes to be followed during the bioleaching of a coal sample with high sulfur content in a packed-bed reactor using an inoculum of adapted native microorganisms from a coalmine. Preliminary results showed that these methodologies could be useful for the control of bioleaching processes and for the identification of critical components of the system. Keywords: coal biodesulphurization, packed bed reactor, iron oxidation, microbial monitoring, fluorescence in situ hybridization 1.
INTRODUCTION Sulfur content, among other factors, determines the energy yield of coal. During combustion, sulfur is mainly transformed into SO2, a pollutant gas that is the main cause of acid rain. The sulfur content in coals is of greater concern to engineers and * e-mail:
[email protected]
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environmentalists than any other of its components. Different technologies have been developed to reduce its presence before combustion. Desulfurization can be achieved by physical, chemical and biological methods. Biological methods are considered of interest because they can be carried out in simple installations with low energy consumption, eliminating pyritic sulphur, which is finely disseminated in the carbonaceous matrix, and part of the organic sulphur [1]. Biodesulfurization may be considered a number of biochemical reactions catalysed, in general, by aerobic microorganisms in an aqueous medium and resulting in the oxidation of the sulfur content into soluble sulphate [2, 3]. Most of the studied processes have been carried out in a stirred reactor, giving pyretic sulphur elimination efficiencies higher than 90%. The main problems associated with stirred reactors, especially the high cost of stirring (4), prevent its industrial application. It can, however, be run on a fixed bed, which would make industrial application easier, although the residence times are longer than with stirring, and the efficiencies lower. The biological desulfurization of coal is influenced by physico-chemical and biological factors. The physico-chemical parameters include pH, temperature, iron concentration in the leachate, etc., and have been the subject of several studies in the last few years, both in stirred reactor process [5, 6] and on packed-bed reactors [7, 8]. The biological factors, that is the study of the types of microorganisms that can be used and the enhancement of their activity in the process has also been widely studied for stirred reactors [9, 10], which is not, however, the case for the packed system. Dissimilatory microbial sulfur oxidation is an important metabolic activity for different chemolithotrophic microorganisms [11], many of which share the same habitat. The control of sulfur dissimilatory microbial populations during the bioleaching process is important for the optimisation of the method. Until recently only conventional microbial methodologies (isolation from enrichment cultures, phenotypic identification and physiology) could be used to study bioleaching processes. These methodologies have a strong bias as a result of the absolute requirement of isolation of microorganisms in selective media prior to their characterization. The introduction of molecular biology techniques, mainly in situ hybridization using specific fluorescent probes (FISH) and denaturating gradient gel electrophoresis able the resolve PCR amplified rRNA genes using primers with different phylogenetic specificity (DGGE), has produced an authentic revolution in microbial ecology, in general, and in the study of complex habitats like bioleaching, in particular [12,13]. The use of both classical isolation and molecular ecology techniques is an effective combination to advance models and control procedures in biohydrometallurgy processes [14]. In this study the methodologies required for the identification and quantification of the microorganisms present in different steps of a biodesulfurization treatment of a pyritic coal in a packed-bed system have been developed. A preliminary report on the use of these methodologies is presented. 2.
EXPERIMENTAL Biodesulfurization was carried out in fixed packed beds with columns of 150 cm high and 10 cm in diameter. A filter was placed at the end of each column to retain coal particles larger than 0.5 mm. Percolation liquid was fed into the packed column with a peristaltic pump. 654
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2.1 Coal Characteristics The coal used for the bioleaching experiments was industrial semianthracite from northern Spain. The sample came from the gravity middlings, with a particle size in the range 0.5-12 mm. Its chemical analysis is shown in Table 1. 2.2 Chemical Analysis Redox potential, dissolved oxygen concentration and pH of the leachates were measured using specific electrodes. Ferrous and ferric iron and total iron concentrations were measured colorimetrically [8]. Elemental analyses, the caloric value and the organic and inorganic sulfur concentration of the coal samples were measured as described in [8]. Table 1. Coal analysis Ash (%) Volatile matter (%) Moisture (%) C (%) H (%) N (%) Total sulphur (%) Pyritic sulphur (%) Organic sulphur (%) Sulphur in sulphates (%) Caloric Value (KJ/kg)
43.24 12.15 1.12 47.92 2.21 1.20 3.11 2.30 0.81 <0.01 18530
2.3 Microbial culture The microbial inoculum was prepared by incubating a native mixture of active desulfurization microorganisms kept in the lab for several years by serial transfers and then adapted to the experimental coal sample (5% suspension) by incubation in a modified 9K medium with ferrous sulphate at pH 2 ± 0.2 and at 30ºC in an Erlenmeyer flask stirred at 100 rpm for 10 days. The number of bacteria was counted in a Thoma chamber. 2.4 Biodesulfurization procedure Two packed columns, each with a capacity for 5 kg of coal, were used. The pH of the packed bed was first stabilised to a constant value of 1.5 using diluted H2SO4. The process began by inoculatig the first column with an adapted bacterial solution containing 2.8 x 108 cells/ml. The columns were washed with the percolated solution every five days. This percolated liquid was stored in a stirred tank during this time and pumped into the column again. The cycle was repeated over a period of 60 days, total 12 cycles, keeping constant the temperature, 20ºC, and the pH, 1.5. The leachate was purged whenever the iron concentration rose above 3000 ppm, a limit determined in previous experiments to avoid jarosites precipitation decreasing the process efficiency [8]. Any resulting loss of volume was made up with distilled water and the pH readjusted to 1.5. When the redox potential was over +500 mV, the leachate of the first column was used to inoculate the second column. At the end of the biodesulfurization step, the columns were washed with diluted HCl, in order to flush away the iron salts formed during the pyrite oxidation. Each column was then washed with enough distilled water to ensure complete removal of the HCl solution.
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2.5 Molecular ecology methods The methodology used for fluorescence in situ hybridization was basically the one originally described by Amann [13] with some modifications to adapt it to acidic environments [14]. Cell recovery from 15 ml column effluents was done by centrifugation after denaturation with formaldehyde (37%) for 4 hours at 4ºC. Samples were filtered through GTTP Millipore filter (0.22 µm), washed with 10 ml of minimal Mackintosh medium [15] to eliminate excess of formaldehyde and heavy metals, followed by a 10 ml wash with PBS buffer (130 mM NaCl, 10 mM sodium phosphate pH 7.2) and air dried. The fluorescent probes used for hybridization are those described in Table 2. The protocols for PCR amplification and gel electrophoresis in denaturating conditions were the ones described in [14]. 3.
RESULTS AND DISCUSSION
3.1 Development of microbial ecology methodologies (FISH and DGGE) for the study of coal biodesulfurization processes As mentioned in the introduction, the use of conventional microbiological techniques has severe limitations for a complete enumeration of the microbial diversity present in any given system, and most important, for the quantification of their populations. The introduction of PCR based technologies prompted the accumulation of rDNA sequences from many different microorganisms, thus allowing the development of complementary molecular ecology techniques like FISH and DGGE. One of the goals of this work was to evaluate the possibility of adapting existing molecular ecology techniques to the study of biodesulfurization processes. Fig. 1 shows the total population in the leachate using DAPI as a universal stain. Fig. 2 shows an in situ hybridization performed with a fluorescent probe specifically designed for α-Proteobacteria.
Figure 1. DAPI (universal) stain for a leachate sample 656
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Figure 2. In situ hybridization of the sample shown in Fig. 1 using α-proteobacteria probe labeled with CY3
As can be seen this specific probe quantifies the bacteria related with specific bioleaching groups present in a leachate from a biodesulfurization experiment in a coal packed-bed system. Complementary tests showed that all the fluorescent probes listed in Table 2 have the described level of specificity and are able to perform in the experimental acidic conditions required for biodesulfurization. Table 2. Fluorescence hybridization probes used in this work Probes Arch 915 Alf968 Acdp821 Thio1 Ntr50
Group Archaeal domain α-proteobacteria Acidiphillim spp.
Acithiobacillus spp. Leptospirillum, Nitrospira group Beta42A Β-proteobacteria Gamma42A γ-proteobacteria Non338 No target (control for nonspecific binding) Eub338 Bacteria domain
Sequence 5´-GTG CTC CCC CGC CAA TTC CT-3´ 5´-GGT AAG GTT CTG CGC GTT-3´ 5´-AGC ACC CCA ACA TCC AGG AGC ACA CAT-3´ 5´-GCG CTT TCT GGG GTC TGC-3´ 5'-CGC CTT CGC CAC CGG CCT TCC-3'
Reference [16] [17] [18] [19] [20]
5´-GCC TTC CCA CTT CGT TT-3´ 5´-GCC TTC CCA CAT CGT TT-3´ 5´-ACT CCT ACG GGA GGC AGC-3´
[21] [21] [22]
5´-GCT GCC TCC CGT AGG AGT-3´
[23]
Unfortunately the use of complementary DGGE techniques, which allow direct and comparative measurement of the level of diversity corresponding to samples obtained in different conditions (14), could not be used in this study because the leachate obtained fror coal desulfurization treatment liberates inhibiting compounds for the PCR amplification step. Different protocols have been used to avoid this problem, but so far none has allowed 657
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us to amplify externally added control DNA. Due to the interest of this technique further experiments will be performed to eliminate this interference. 3.2 Microbial ecology of the bioleaching packed bed columns The original idea was to use the native chemolithotrophic microbes present in the coal sample to promote its biodesulfurization. Preliminary results showed that although some microbial growth was obtained after incubation of coal samples at acidic leaching conditions, the number of cells was extremely low, and difficult to detect using hybridization probes with different specificity. Accordingly, it was decided to use bioaugmentation with a natural microbial consortia isolated from the same mine site and grown in the lab on pyritic contaminated coal which had shown a high level of desulfurization efficiency [8]. An aliquot of this consortium was used as inoculum for the coal treatment in packed columns. A pre-adaptation of the consortium to the coal was promoted using the conditions described in the experimental section. Hybridization done with an aliquot of the inoculum used to seed the column did not show the presence of very active microorganisms. The hybridization analysis of the leachate from the first column after thirty days of biodesulfurization showed that 55% of the DAPI stained prokaryotes gave positive hybridization with the α-Proteobacteria probe. 2% of the active cells gave a positive signal with the Leptospirillum probe. No hybridization was obtained using a β- or γ-Proteobacteria probe nor with the specific probe for Acidithiobacillus ferrooxidans. Fig. 3 shows the evolution of the populations of α-Proteobacteria and Leptospirillum ferrooxidans over time. A net increase of the iron oxidizing Leptospirillum which rose up to 20% after 50 days of bioleaching was observed, which coincides with the appearance of oxidized iron in the leachate (Fig. 4). The number of α-Proteobacteria decreased with time, reaching its lowest value after 50 days. At longer leaching times, 60 days, the number of α-Proteobacteria increased dramatically (up to 55%), while the number of iron oxidizing bacteria decreased to undetectable values. This change in the bacterial population could be correlated with a decrease in the redox potential of the solution (data not shown).
Figure 3. Evolution of the microbial population during the biodesulphurization experiment in column 1
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Figure 4. Evolution of the iron separated in column 1 leachates
The leachate obtained after 35 days (7 cycles) of column 1 was used to inoculate column 2. The amount of active α-Proteobacteria increased to 45%, and a low value for iron oxidizing Leptospirillum, 4%, was obtained (Fig 5).
Figure 5. Evolution of the microbial population during the biodesulphurization experiment in column 2
After 10 days, the numbers of α-Proteobacteria decreased to a 7% value, while the activity of L. ferrooxidans reached a 41% value. This increase in iron-oxidizing activity followed the increase of ferric iron concentration in the leachate (Fig. 6).
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Figure 6. Evolution of the iron separated in column 2 leachates.
As in the first column, longer periods of bioleaching decreased the number of ironoxidizing bacteria while the α-Proteobacteria rose again. This change in the microbial population could be correlated, as before, with a change in the redox potential of the leachate. 3.3 Coal biodesulfurization Table 3 shows the evolution of sulfur content in the treated coal samples. Pyritic desulfurization was more effective in the second column, 30%, than in the first one, 13.5%. When the desulfurization efficiency is calculated using the total sulphur content, the numbers are lower than those calculated using the pyrite concentration, indicating that some insoluble sulphur compounds are produced which can not be removed after the final acidic wash. Table 3. Total sulphur content and the different forms of sulphur present in the coal before and after treatment. All results are expressed on a dry basis Sample Original coal Column-1 Column-2
Total sulphur % (% desulfurization) 3.11 2.96 (4.8%) 2.38 (23.4%)
Pyritic S. % (% desulfurization) 2.30 1.99 (13.5%) 1.61 (30%)
Organic S. % (% desulfurization) 0.81 0.80 (1.2%) 0.70 (13.6%)
Sulphur in sulphates % <0.01 0.17 0.07
Some removal of organic sulfur can be also observed (Table 3) but given that the organic sulfur is calculated as the difference between the total sulfur and the inorganic sulfur (pyritic and sulfate), care should be taken before any conclusion concerning the removal of this recalcitrant sulfur contamination is reached. 4.
CONCLUSIONS Fluorescence hybridization probes with different domain, kingdom, phylum, genus and species specificity can be used to identify and quantify prokaryotes involved in the biodesulfurization of pyritic contaminated coals (FISH). The acidic conditions used for bioleaching and the high concentration of heavy metals present in the solution do not interfere with the hybridization efficiency nor with the specificity. 660
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PCR amplification using specific primers for the identification of microbial populations present in coal leachates was not possible, probably due to inhibition by compounds released during the process. Therefore, this important complementary molecular ecology tool cannot be used until a protocol for removing the inhibitory substances is implemented. Preliminary pilot experiments of coal biodesulfurization in packed bed columns using enriched chemolithotrophic microorganisms from a coalmine showed that FISH technology can be used to follow the microbial population evolution over time. Two sets of experiments showed that the iron oxidizing bacteria of the Nitrospira group, L. ferrooxidans, increased during the bioleaching, while an unidentified α-Proteobacteria that appeared in high numbers at the beginning of the leaching process first decreased and then increased at the end. The possible correlation between this fact and the evolution of redox potential is under study. The chemolithotroph Acidithiobacillus ferrooxidans has not been found to be associated with the biodesulfurization of the pyritic coal used in this work. Using this protocol pyritic biodesulfurization values of 30% have been achieved. Pilot experiments for biodesulfurization in heaps are recommended to evaluate the potential of this process. Molecular ecology methods can be used for monitoring the evolution of microbial populations associated with biohydrometallurgical processes. ACKNOWLEDGEMENTS This work was supported by the European Coal and Steel Community Coal RTD Programme (Contract number 7220-PR-098). REFERENCES
1. Rossi G. Fuel 72 (1994) 1581-1592. 2. Monticello D.J, Finerty W.R. Ann. Rev. Microbio. 39 (1985) 371. 3. Cara J, Aller A, Gómez E, García A.I, Sánchez M.E. Ingeniería Química, 386 (2002) 237-242. 4. Orsi N, Rossi G, Tríos P, Valenti P.D, Zecchin A. Resour. Conserv. Recyc. 5 (1991) 211-230. 5. Andrews G.F, Noah K.S, Glenn A.W, Stevens C.J. Fuel Processing Technol. 40 (1994) 283. 6. Martínez O, Aller A, Alonso J, Gómez E, Morán A. Eighth International Conference of Coal Science, Oviedo, Spain; (1995).p. 1749-53. 7. Morán A, Aller A, Cara J, Martínez O, Encinas J.P, Gómez E. Fuel Processing Technol. 52 (1997) 155-164. 8. Cara J, Aller A, Otero M, Morán A. Appl. Microbiol Biotechnol, 55 (2001) 49-54. 9. Olsson G, Pott B.M, Larsson L, Holst O, Karlsson H.T. Fuel Processing Technol. 40 (1994) 277. 10. Rossi G. Microbial desulfurization of coal. In: Rossi G (ed) Biohydrometallurgy. McGraw-Hill, Hamburg, (1990) pp 528-538. 11. Ehrlich H.L. Geomicrobiology, fourth edition. Marcel Dekker Inc., New York. (2001) 12. Manz, W., Amann, R., Ludwig, W., Wagner, M. and Schleifer, K.-H. Phylogenetic oligodeoxynucleotide probes for the major subclasses of proteobacteria: problems and solutions. Syst. Appl. Microbiol. 15 (1992) 593-600. 13. Amann, R.I. In situ identification of microorganisms by whole cell hybridisation with rRNA-targeted nucleic acid probes. In: Akkermans, A.D.L., Van Elsas, J.D., De 661
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Bruijn, F.J. (eds.), Molecular Microbial Manual. Kluwer Academic Publishers, Dordrecht, (1995) 3.3.6, pp.1-15. 14. González-Toril E, Gómez F, Rodríguez N, Fernández-Remolar D, Zuluaga I, Marín I, Amils R. Hydrometall., in press. 15. Mackintosh, M.E.. Nitrogen fixation by Thiobacillus ferrooxidans. J. Gen. Microbiol. 105, (1978) 215-218. 16. Stahl, D.A., Amann, R. Development and application of nucleic acid probes. Nucleic Acid Techniques in Bacterial Systematic. E. Stackebrandt and M. Goodfellow (eds.) John Wiley and Sons Ltd., Chichester, UK, (1991) pp: 205-248. 17. Neef, A. Anwendung der in situ Einzelzell-Identifizierung von Bakterien zur populationanalyse in komplexen mikrobiellen biozönosen. Doctoral thesis (1997) Trchnische universität München. 18. Peccia, J., Marchand, E.A., Silverstein, J., Hernandez, M. Development and application of small-subunit rRNA probes for assessment of selected Thiobacillus species and members of the genus Acidiphilium. Appl. Environ. Microbiol. 66 (2000): 3065-1390. 19. Stoffels, M. Not publissed. 20. Daims, H., Nielsen, J.L., Nielsen, P.H., Schleifer, K.H., Wagner, M. In situ characterization of Nitrospira-like nitrite-oxidizing bacteria active in wastewater treatment plants. Appl. Environ. Microbiol. 67 (2001): 5273-5284. 21. Manz, W., Amann, R., Ludwig, W., Wagner, M., Schleifer, K.H. Phylogenetic oligodeoxynucleotide probes for the major subclasses of Proteobacteria: problems and solutions. Syst. Appl. Microbiol. 15 (1992): 593-600. 22. Wallner, G., Amann, R., Beisker, W. Optimizing fluorescent in situ hybridization with rRNA-targeted oligonucleotide probes for flow cytometric identification of microorganisms. Cytometry 14: (1993) 136-143. 23. Amann, R.I., Krumholz, L., Stahl, D.A. Fluorescent-oligonucleotide probing of whole cells for determinative, phylogenetic, and environmental studies in microbiology. J. Bacteriol. 172 (1990): 762-770.
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Regeneration of hydrogen sulfide using sulfate-reducing bacteria for photo catalytic hydrogen generation Yui Takahashia*, Koichi Sutob, Chihiro Inoueb, Tadashi Chidab a
Department of Geoscience and Technology, Tohoku University, Aoba, Aramaki, Aoba-ku, Sendai, 980-8579, Japan b Department of Environmental Studies, Tohoku University, Aoba, Aramaki, Aoba-ku, Sendai, 980-8579, Japan
Abstract Recently, a process of hydrogen generation using ZnS particles as photocatalyst has been developed [1]. In this process, hydrogen sulfide dissolved in alkaline solution is changed into poly-sulfide ion with the simultaneous evolution of hydrogen gas. In order to obtain the cyclic operation of the photo catalytic hydrogen generation system, it is necessary to convert again poly-sulfide ion into hydrogen sulfide. This paper proposes the use of sulfate reducing bacteria (SRB) for the regeneration of hydrogen sulfide in the process of the photo catalytic hydrogen generation. The batch cultivation experiments of SRB, which were isolated from the cooling tower of a geothermal power plant, were carried out using a culture medium containing only poly-sulfide ions as sulfur compounds. SRB were able to grow actively in the medium and to use poly-sulfide ion as electron acceptor. Converting poly-sulfide ion into hydrogen sulfide by SRB enables to recycle sulfur as a source material. So the photo catalytic hydrogen generation system can be viable. Keywords: sulfate reducing bacteria, poly-sulfide, hydrogen sulfide, sulfur cycle, regeneration, and hydrogen generation. 1.
INTRODUCTION Making use of hydrogen energy has been considered as one of the means to solve the energy and environmental problems. Recently, the photocatalyst, which generates hydrogen with sunlight, has been developed [1] and the hydrogen generation from hydrogen sulfide has become viable. In the photocatalytic reaction, a free electron and a positive hole are arising by exposing a semiconductor to light radiation energy. The free electrons transfer to the hydrogen ions at the interface, and the holes accept electrons from the sulfide ions oxidized at the interface. If Na2S is used as the reducing agent, the generation of hydrogen is described by the following reaction:
* Corresponding author. Mailing address: Aoba, Aramaki, Aoba-ku, Sendai, 980-8579, JAPAN. Phone: +81-(0)22-217-7404, E-mail:
[email protected]
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Na2S + H2O → 2Na+ + HS- + OH(1) 2+ 2HS → 2H + S2 + 2e (2) + 2H + 2e → H2 (3) As a result of these reactions, hydrogen sulfide is converted into hydrogen and polysulfide ion [1, 2]. Under weak acid conditions, elemental sulfur precipitates according to the following reaction: 2HS- → H+ + HS- + S + 2e(4) At present, a lot of sulfur in excess of a demand is produced and scrapped all over the world. Therefore, it is important to reduce elemental sulfur into hydrogen sulfide, thus obtaining sulfur recycling in the system of hydrogen generation. In nature, sulfur has various forms, and many kinds of microorganisms play important roles in the global sulfur cycle. Sulfate reducing bacteria (SRB) are specific microorganisms reducing sulfate ion to hydrogen sulfide during their growth. These bacteria grow under anaerobic conditions, utilize organic acids as electron donors and have the ability to use sulfate as a terminal electron acceptor. Some species can also use a variety of other sulfur compounds including sulfite, thiosulfate and elemental sulfur, but there is no evidence up to now that SRB can use poly-sulfide ion as a substrate. Applications of SRB are used in the mining industry, where wastewater often contains a large amount of sulfate ion. The selective precipitation of metals using hydrogen sulfide produced biologically by SRB has been proposed and the scaling-up of this process is currently in progress [3-6]. However, there is scarcely any instance about other industrial use of SRB except the mining industry. This study proposes the use of SRB for the regeneration of hydrogen sulfide in the process of the photocatalytic hydrogen generation. The batch cultures of SRB using a culture medium containing only poly-sulfide ion as sulfur compounds was carried out. 2.
MATERIALS AND METHODS
2.1 Microorganisms, culture conditions and isolation Bacterial samples were obtained from the sludges in the cooling tower of the Uenotai geothermal plant (Tohoku Electric Power Co. Inc. Japan) at Akita prefecture, Japan. The samples were taken at 5cm depth from the surface of the sludge (June 18, 2000). To maintain anaerobic conditions, the samples were stored in 20ml tubes filled up to the lid, sealed with screw caps and kept at temperature lower than 5°C during their transportation to the laboratory. A nutritive salt solution contains: KH2PO4, 0.5g; NHCl4, 1.0g; Na2SO4, 1.0g; CaCl2.2H2O, 0.1g; MgSO4.7H2O, 2.0g and yeast Extract, 1.0g per liter. After autoclaving at 121° for 15min and cooling to room temperature, the salt solution was mixed with the aseptically filtered (0.2 µm pore size) solution containing: sodium lactate (70%), 5.0g/l; CH3COONa.3H2O, 2.8g/l; FeSO4.7H2O, 0.2g/l; sodium thioglycolate, 0.1g/l and sodium ascorbate, 0.1g/l. The resultant solution referred to the basic medium was used for the cultivation. The pH of the medium was adjusted to 7.2~7.4 by addition of a NaOH solution. All reagents were supplied from Wako Pure Chemical Industries. Ltd. Japan.
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8ml of the medium was introduced into sterilized 10ml test tubes equipped with screw caps and gassed headspace with N2. Then, the each sample sludge of about 5gram was inoculated into the test tubes, which were sealed and incubated at 30°C in a shaker. A black precipitate (FeS) is formed during the growth of SRB. When the growth of SRB was confirmed, the culture medium was diluted and spread on the same medium with 1.5% agar under anaerobic conditions. Those plates were incubated at 30°C in an anaerobic jar with the gas pack being able to absorb oxygen and to generate carbon dioxide (AnaeroPack Kenki: Mitsubishi Gas Chemical Co. Ltd. Japan). Individual colonies were picked and were transferred into the basic medium to examine whether the strains have sulfate reducing activity. The subcultures of the isolates were cultivated as described above. For the following experiments, inoculums were obtained from the subcultures. Desulfovibrio desulfuricans IFO 13699 used as standard strain was obtained from the culture collection of the Institute for Fermentation, Osaka (IFO), Japan. The strain was also cultivated and examined in the same way. 2.2 Utilization of organic substrates as electron donor The ability of isolated SRB to utilize various organic substrates were tested using modified basic media containing (a) 5g/l of sodium lactate (70% solution), (b) 7g/l of CH3COONa.3H2O or (c) 1.2g/l of ethanol as a sole electron donor. In order to prevent substrates of subcultures from having an effect on these tests, these cells were gathered by centrifugation of 1ml subcultures solution. The pellet was washed and resuspended in the modified basic medium. Cultivations were carried out in the same way as subcultures, and the growth of the cells was confirmed by the formation of black precipitate. 2.3 PCR-restriction fragment length polymorphism (RFLP) analysis
Isolates were cultured in 200ml sealed vessels at 30°C in anaerobic condition. The cells were harvested at the stationary phase by centrifugation. DNA extraction and purification from the cell pellets were performed by using InstaGene Matrix (Bio-Rad Co. Ltd. Japan). The universal primers used for PCR amplification were fD1 (5’CCGGATCCGTCGACAGAGTTTGATCCTGGCTCAG-3’) and rD1 (5’CCAAGCTTCTAGAAAGGAGGTGATCCAGCC-3’), which covered almost all of 16S rRNA region of eubacteria [7]. For PCR amplification, 40µl of DNA extraction, were mixed with 10 µl of 10×PCR buffer (20mM Tris-HCl [pH=8.0], 100mM KCl, 0.1% EDTA, 1mM DTT, 0.5% Tween 20, 0.5% Nonidet D-40, 50% Glycerol), 8 µl of dNTP (each of 2.5mM dATP, dTTP, dCTP, dGTP), 1 µl of two primer solutions 10pmol/ml, 1 µl of Ex-Taq DNA polymerase and 39 µl of deionized water. The PCR reagents were obtained from TakaraBio Inc. (Japan). Amplification was carried out in a Gene Amp PCR System 2400 (Applied Biosystems Japan Ltd.) for 30 cycles with each cycle consisting of 15s at 94°C, 15s at 55°C and 90s at 72°C. PCR products were loaded together with a λ/HindIII molecular size marker on a 1% agarose gel to evaluate the PCR. Amplified 16S rRNA genes were digested by restriction endnuclease; HaeIII[GG|CC], HhaI[GCG|C] and RsaI[GT|AC] (BioChain Institute, Inc. USA). After the resultant RFLP products were separated on 2% agarose gel, these were stained with ethidium bromide and were visualized by UV excitation. An 100bp DNA ladder (TakaraBio Inc.) was used as a molecular size marker. The fragment sizes were determined by a densitograph (ATTO Co. Ltd. Japan: model AE-6920-MLF) and its software.
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2.4 Generation of hydrogen sulfide from poly-sulfide A PS medium containing sodium lactate (70%), 10.0 g/l; CH3COONa.3H2O, 5.6 g/l; FeCl2.4H2O, 0.0029 g/l and MgCl2.6H2O, 1.8 g/l instead of both FeSO4.7H2O and MgSO4.7H2O was used for the test of hydrogen sulfide generation from poly-sulfide. 85ml of the PS medium was introduced into each sterilized 100ml vial and gassed headspace with N2. Modifications of the PS medium from the basic medium were removal of all sulfate ions, increased substrates and decreased ferrous ion constituents. These modifications enabled to detect generation of hydrogen sulfide and growth of SRB more clearly. Poly-sulfide ion is usually provided as metal sulfides, which combine with more sulfuric atoms than common one. Poly-sulfide ion has a straight chain structure of sulfuric atoms, [S-S-S…]2-, and is decomposed by addition of acids. In this study, the potassium sulfide reagent (Wako Pure Chemical Industries, Ltd. Japan) that was used as a source of poly-sulfide had low purity (>30%). This means that the solution contains sulfide ion and other sulfuric chemical species. The potassium sulfide solution was prepared as follows: First, 1.1g (about 0.1mol as K2S) of the potassium sulfide powder was dissolved in 100ml of deionized water, and dispersed by ultrasonic waves. After the solution was stirred for ten minutes, it was filtered with a membrane filter (pore size 0.2 µm). The pH of the solution was about 11.5. The culture medium for these experiments had a volume of 100ml and was prepared by adding 15ml of the potassium sulfide solution into the PS medium. The pH of the culture solution was 7.8. Inoculums were treated as described in section 3.1. After inoculation, the vials were sealed with butyl rubber stoppers and aluminum seals. Those batch cultures were incubated at 30°C with shaking. Control experiment without inoculation was done in the same way. 1.5ml of the liquid was sampled once a day using syringes equipped with 25G needles. 2.5 Chemical analysis Concentration of hydrogen sulfide was measured by using a high-performance liquid chromatography (HPLC). The samples were filtered through 0.2 µm syringe filter unit Dismic (ADVANTEC Co. Ltd. Japan). HPLC analysis was performed using a liquid chromatography set (TOSO Co. Ltd. Japan) equipped with multiple-wavelength UV detector set at 220 nm. The mobile phase was 4mM K2HPO4 (pH 9.1). Samples (5 µl) were injected on to a guard column: TSK-GUARD COLUMN IC-A and a column: TSKGEL IC-ANION-PW at 40°C. The flow rate of solvent was 1.2ml/min. 3.
RESULTS AND DISCUSSION
3.1 Characterization of isolated strains Four strains were isolated from the samples of the sludge collected at the cooling tower of the geothermal plant, and were designated as the strain U shown in Table 1, in which the characteristics on shape of isolates are presented. The colonies on the plates were very small (below 1mm in diameter) except strain U1, and for this reason, colonies were taken after two weeks of incubation to obtain a sufficient amount of biomass for each isolate. The isolated strains were mostly rods. It was difficult to isolate other SRB, which grow without formation of colonies or cannot use of lactate and acetate as electron donors.
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Table 1. Characteristics of isolates Strain
Color of colony
Shape of colony
Shape of cell
U1
white
star spread thinly
short rod
U2
translucence
very fine spots
rod
U42
translucence
U61
translucence
Round diameter <1mm
coccus or short rod
very fine spots
rod
Table 2 shows what the strains U and the standard strain can use as the electron donor. Every strain can use lactate, and the strains U1 and U42 can use acetate too, but only U42 cannot use ethanol. The strains U1 and U42 are the complete oxidation type of SRB, while U2 and U61 are the incomplete oxidizers. It is suggested that the strain U42 is the genus Desulfomonile, because of its shape and ability not to use ethanol [8-10]. Table 2. Utilizations of substrates U1
U2
U42
U61
D. desulfuricans
Lactate
++
++
++
++
++
Acetate
++
-
++
-
-
Ethanol
+
++
-
++
++
incomplete
incomplete
Oxidation type complete incomplete complete ++ utilized, + poorly or slowly utilized, - not utilized
Figure 1 shows the PCR-RFLP patterns by using three different restriction endnucleases. The PCR-RFLP patterns of each restriction endnuclease indicate that the strains U2 and U61 are identical with D.desulfuricans IFO 13699. The PCR-RFLP patterns of strains U1 and U42 are clearly different from that of D. desulfuricans, therefore it is supposed that they belong to different genera. The results are consistent with the utilization of substrate. Therefore, the strain U1 is Desulfobacterium, Dusulfobacter, Desulfosarcina or Desulfotomaculum, and the strains U2 and U61 are Desulfovibrio and U42 is Desulfomonile.
Figure 1. PCR-RFLP patterns by using different restriction endnucleases. (a) HaeIII, (b) HhaI, (c)RsaI. Lane D, D. desulfuricans. Lanes (M); molecular size marker; 2000, 1500, 1000, 900, 800, 700, 600, 500, 400, 300, 200, 100bp from the top
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3.2 Generation of hydrogen sulfide Fig. 2 shows the concentration of hydrogen sulfide produced during poly-sulfide reduction by SRB. In all the experiments at the beginning of the reaction, about 20mg/l of hydrogen sulfide were already detected. A probable reason is that poly-sulfide was reduced by chemical reducers (sodium thioglycolate and sodium ascorbate) contained in the medium. To check this assumption, the above reducing compounds were added to the potassium sulfide solution, and hydrogen sulfide generation was confirmed by HPLC analysis. The amount of hydrogen sulfide generation using D.desulfuricans was lower than that produced by the four isolates of SRBs. The results suggest that these SRBs can use poly-sulfide as electron acceptor instead of sulfate. The amount of hydrogen sulfide produced from poly-sulfide was higher than that from sulfate (date not shown). When SRB use poly-sulfide as electron acceptor, there is no need to produce APS (adenosyl-5phospho-sulfate) on their metabolism. Therefore, they can efficiently use electrons and energy which are usually consumed for reducing sulfate ion to APS [11-16]. Although the amount of generated hydrogen sulfide was different from one isolated strain to another, however it seems that poly-sulfide reduction occurred with all the isolated species of SRB. However further research is required to confirm this metabolism avoiding the initial production of hydrogen sulfide due to the presence of chemical reducers. Future research should also address problems such as separation of sulfur from the potassium sulfide solution and so on.
Figure 2. Production of hydrogen sulfide from poly-sulfide; , U61; +, D.desulfuricans; 4.
, U1;
, U2;
, U42;
, control
CONCLUSION Four SRB strains were isolated from the sludge of the geothermal plant. Two isolates are the incomplete oxidation type of SRB. It is assumed that these belong to D. desulfuricans. Two other isolates are complete oxidizers. All isolates can use poly-sulfide as an electron acceptor. But, the reduction capability of poly-sulfide differs from one isolate to 668
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another. So, it seems possible that these SRB could be used for the reduction of polysulfide in the process of hydrogen generation obtaining the cyclic operation of sulfur compounds. ACKONWLEDGEMENTS This work was financially supported in part by a Grand-in-Aid for Scientific Research (no. 14103016) from Japan Society for the Promotion of Science. REFERENCES
1. T. Arai, S. Sakima, H. Yoshimura, K. Shinoda, B. Jeyadevan, K. Tohji, A. Kasuya and Y. Nishina, Proc. Int. Symp. on Cluster Assembled Mater. IPAP Conf. Series 3 (2001) 75. 2. K. Shinoda, T. Arai, H. Ohshima, B. Jeyadevan, A. Muramatsu, K. Tohji and E. Matsubara, Materials Transactions, 43 (2002) 1512. 3. C. García, D.A. Moreno, A. Ballester, M.L. Blázquez and F. Gonzáles, Minerals Engineering, 14 (2001) 997. 4. F. Glombitza, Waste Management, 21 (2001) 197 5. S. Foucher, F. Battaglia-Brunet, I. Ignatiadis and D. Morin, Chemical Engineering Science 56 (2001) 1639. 6. J.R. Lloyd, J. Ridley, T. Khizniak, N.N. Lyalikova and L.E. Macaskie, Appl. Environ. Microbiol. 65 (1999) 2691 7. W.G. Weisburg, S.M. Barns, D.A. Relletier and D.J. Lane, J. Bacteriol. Jan. (1991), 697. 8. R. Devereux, M. Delaney, F. Widdel and D.A. Stahl, J. Bacteriol. Dec. (1989) 6689. 9. V. J. Fowler, F. Widdel, F. Pfennig, C. R. Woese and E. Stackebrandt, Syst. Appl. Microbiol. 8 (1986) 32 10. R. Devereux, S.H. He, C.L. Doyle, S. Orkland, D.A. Stahl, J. LeGall and W.B. Whitman, J. Bacteriol. 172 (1990) 3609 11. H. Biebl and N. Pfennig, Arch. Microbiol. 112 (1977) 115 12. H. Cypionka, Arch. Microbiol.152 (1989) 237 13. W. Dilling and H. Cypionka, Arch. FEMS. Microbiol. Lett. 71 (1990) 123 14. R. M. Fitz and H. Cypionka, Arch. Microbiol.152 (1989) 369 15. C. Fukusaka, Nature 192 (1961) 427 16. J.M. Odom and H.D. Peck, FEMS. Microbiol. Lett. 12 (1981) 47
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Remediation of sites contaminated by heavy metals: sustainable approach for unsaturated and saturated zones Ludo Diels, J. Geets, J. Vos, K. van Broekhoven, L. Bastiaens Environmental Technology Centre, Flemish Institute for Technological Research, Vito, Boeretang 200, B – 2400 Mol, Belgium Abstract Large areas of soil and groundwater bodies are contaminated by heavy metals due to the emissions by mining activities, metal processing, surface treatment, electronic or paint industry etc. This makes heavy metal contamination one of the largest environmental problems. As the scale of contamination will not allow metal removal techniques in a cost effective way, metal immobilisation could be an approach to control the risks and spreading of these contaminants. The addition of soil additives (as silicates, phosphates, carbonates, iron oxides, etc) can reduce the bioavailability of metals in the unsaturated zone and allow normal plant growth. Such an immobilisation process will reduce the spreading of metals to the groundwater and to the air. However in case of groundwater pollution the immobilisation of metals became a possible alternative to pump & treat technology only the last 5 years. Mostly immobilisation could be performed by Permeable Reactive Barriers in which metal sorbing or metal precipitation inducing compounds are added. This paper deals with the creation of reactive zones in groundwaters and aquifers containing besides heavy metals also high sulphate concentrations. A metal immobilising activity can be induced by Sulphate Reducing Bacteria (SRBs). Metals will precipitate as non-soluble metal sulphides in situ in the aquifer in a way that they become immobilised. In order to induce the bacteria a carbon source (electron donor) must be provided which stimulates these bacteria and their activity and which ensures the consumption of oxygen by microbial metabolism. This paper describes the effect of different electron donors and the influence of the concentration of this on SRB activity and metal precipitation. Furthermore, the importance of the sulphate concentration, the control of ORP (oxidation reduction potential), and the irreversibility of the metal sulphide were studied. The evaluation is done by using batch tests. Several substrates like lactate, ethanol, methanol, and also the slow release compound HRC® can be used. Acetate shows a lower induction rate which can be increased by adding some other fermentable compounds as yeast extract or lactate. Molasses shows always a fast induction of the sulphate reduction process. But the final reached metal concentrations are higher than those obtained with other pure substrates. It is generally known that substrate concentrations play an important role in the anaerobic sulphide production since too high concentrations will induce methanogenic conditions resulting in failure of metal removal. 671
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Sequential extraction tests showed that the metals are removed from the leachable and exchangeable fractions and are shifted mainly to the Fe-Mn-oxide fraction which indicates a strong irreversibility of the precipitation process. However in the case of molasses a larger part of the metals remains in the organic fraction due to some complexation processes. Nickel showed a very slow rate of metal immobilisation and it took more time before it was really tightly bound to the Fe-Mn-oxide fraction. The method could also be used in case of high ORP (e.g. 400 mV) and low pH (e.g. 3). The addition of a redoxmanipulating compound resulted in the induction of sulphate reducing activity of the SRBs. 1.
INTRODUCTION Former mining activities, non-ferrous metals processing, surface treatment and electronic industries lead to the contamination of soils and groundwaters with heavy metals. Depending on the industrial activities, these metals are sometimes accompanied by organic pollutants like BTEX or chloro-aliphatics (VOCLs, volatile organic compounds). Also the presence of anthropogenic iron or even arsenic can play a role in this contamination. Metal emissions caused large and diffuse contaminations of the soil (unsaturated zone). Leakage of landfills and rainfall were responsible for the penetration of these contaminants into the groundwater. After long periods of exposure the heavy metal plumes became very large and are threatening groundwater reserves. Especially mining activities and non-ferrous metals processing lead to large landfills and contaminated lagoons (often with acidic pH). Not only leakage of these systems caused a continuous threat but risks are also related to possible disasters as dike breakdowns. Two recent disasters are well known in Europe as the Baja Mare and Donana accidents [1] in Romania and Spain respectively. Since metals cannot be degraded, the only way to control the problem is to remove and concentrate them or to reduce their bioavailability and mobilityspreading. The removal of metals from soil [2] or groundwater [3] is possible but not very cost-effective. Further these techniques are leading to large amounts of waste that must be deposited. This paper deals with the use of immobilisation techniques for the reduction of risks related to the presence of heavy metals in soil and groundwater. The addition of soil additives (as silicates, phosphates, carbonates, iron oxides, etc) can reduce the bioavailability of metals in the soil and allow normal plant growth. Such an immobilisation process will reduce the spreading of metals to the groundwater and to the air [4-6]. In case of groundwater pollution the metals can be treated in situ in the aquifer in a way that they become immobilised. Several in situ treatments and pilot operations using Permeable Reactive Barriers comosed of comost and limestone have been published [7 9]. The compost will reduce the ORP and provide electron donors for Sulphate Reducing Bacteria (SRBs). Also the polymeric humic acids can play a role in immobilising the metals by binding them to their functional groups. The limestone is added to control the pH. Also sorption barriers can be constructed based on metal binding compounds as zeolites, silicates, apatites, limestone etc. [10, 11]. In many cases the heavy metals contaminated groundwater contains also high concentrations of sulphates (acid mine drainage, effluents from metal processing activities etc.). In this case the immobilisation can be done by inducing SRBs to reduce sulphates into sulphides which will precipitate the heavy metals. In order to induce the bacteria a carbon source (electron donor) must be provided to support the growth and to ensure 672
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oxygen consumption. It is important to make good evaluations about the feasibility of this technology as a quite low ORP is necessary and methanogenic activity must be avoided. Accurate chemical analysis of the inorganic and organic content is necessary [12]. Feasibility (batch and column) tests will be presented. Special attention will be paid to the use of different carbon sources and their effectiveness in stimulating sulphate reduction. Evaluations of the in situ precipitates show the irreversibility of the metal precipitates on the aquifer. In some cases the low pH and high ORP will hinder the SRB growth. Special attention will be paid to the addition of ORP reducing compounds in order to overcome this problem. Also the tolerant concentrations of toxic heavy metals will be examined. 2.
MATERIALS AND METHODS Several tests were done in order to verify different aspects of in situ bioprecipitation of heavy metals. Therefore natural groundwaters (with different characteristics) from a surface treatment site and from two metal processing sites were tested. This allowed to control the effect of different kinds and concentrations of substrates, influence of heavy metals of sulphate concentration and of pH and ORP etc. In order to define the conditions under which sulphate reduction and heavy metal bioprecipitation could take place, several batch tests were set up. Column tests were already presented [13]. The batch tests were done with 80 g aquifer material (undisturbed samples) and 186 ml groundwater. Besides an abiotic control (containing HgCl2), a microcosm containing solely aquifer and groundwater was studied in all tests. In addition, acetate was supplied as electron donor at different concentrations to determine the optimum amount of substrate. In other tests, the metal immobilization was studied after addition of different e-donors, such as acetate, molasses, lactate and HRC® to stimulate the SRB population. Furthermore, strain Desulfovibrio desulfuricans 8301 was inoculated to create optimal conditions for sulphate reduction. Finally, extra nutrients and sulphates were added in the form of Postgate C medium [14]. All manipulations were done in an anaerobic chamber. Aquifer and groundwater samples were taken as anaerobic as possible and stored under anaerobic conditions. 3.
RESULTS AND DISCUSSION
3.1 Role of sulphate concentration Sulphate must be present in order to induce the SRBs for sulphate reduction and it must be present in a stoechiometrical sufficient amount in order to precipitate the metals present in the groundwater. From several experiments it was shown that a sulphate concentration of at least 200 mg/l was necessary to induce the SRBs. Test with pure cultures of sulphate reducing bacteria showed that at concentrations below the 200 mg sulphate/l only hydrogen could be used as electron donor. Table 1 shows the results of Zn removal in a test done on a groundwater containing only 74 mg sulphate /l. The results show that with the addition of Postgate C (containing an extra sulphate source) Zn removal could be induced (from 1800 µg Zn/l to 25 µg Zn/l). The addition of extra bacteria (Dd8301 = Desulfotomaculum desulfuricans) also showed an immediate metal removal probably due to the presence of HS- on the biomass. Condition 6 containing 68 mM acetate and inoculated with strain Dd8301 shows no blackening (as a result of FeS precipitation). The ORP in condition 5 and 7 decreased to – 300 mV. In condition 6 the ORP dropped after 12 weeks until -290 mV. Probably the produced HS was only sufficiently available for Zn removal. 673
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Table 1. Zn concentration after induction of sulphate reduction in a groundwater with a low sulphate concentration (74 mg/l) Zn 1. Aquifer + groundwater
T0 1870
T4 1400
T8 1510
T12 644
2. Aquifer + groundwater + 0.5 mM HgCl2
1840
1630
1940
2270
3. Aquifer + groundwater + 1 ml K-acetate (25%) 4. Aquifer + groundwater + 5 ml K-acetate (25%)
1350 1570
922 884
1200 1160
1500 434
5. Aquifer + groundwater + 1 ml K-acetate (25%) + Dd8301
38
14
14
25
6. Aquifer + groundwater + 5 ml K-acetate (25%) + Dd8301
16
144
118
14
1170
51
20
26
7. Aquifer + groundwater + Postgate C medium
Tx: x is number of weeks of incubation. Zn concentration is presented in µg Zn/l. Grey colouring indicates the appearance of a black colour.
3.2 Role of substrate concentration In a second test a groundwater containing a higher sulphate concentration (506 mg SO4/l) was used. Different substrates and substrate concentrations were used. The results are presented in table 2. Conditions 3 and 5, both with a low acetate concentration resulted in a low redox potential of about – 230 mV and a complete removal of Zn (from 103,000 µg Zn/l to 15 µg Zn/l). In the conditions 4 and 6 no metal removal was observed and a high gas production was detected. This is probably because methanogenic conditions were created which did not led to metal removal. However it should be noted that redox measurements do not indicate methanogenic conditions. The extra addition of SRBs only resulted in an initial immediate removal of some metal due to the presence of some HS- on the biomass. Table 2. Zn concentration after induction of sulphate reduction with different substrate concentrations Zn (µg Zn/l) 1. Aquifer + groundwater 2. Aquifer + groundwater + HgCl2 3. Aquifer + groundwater + acetate 4. Aquifer + groundwater + 5 x acetate 5. Aquifer + groundwater + acetate + Dd8301 6. Aquifer + groundwater + 5 x acetate + Dd8301 ORP (mV) 1. Aquifer + groundwater 2. Aquifer + groundwater + HgCl2 3. Aquifer + groundwater + acetate 4. Aquifer + groundwater + 5 x acetate 5. Aquifer + groundwater + acetate + Dd8301 6. Aquifer + groundwater + 5 x acetate + Dd8301
T0 101,000 109,000 103,000 103,000 93,100 96,100
T8 79,200 94,200 82,800 109,000 77,200 91,600
T20 49,000 62,400 15 90,000 12 72,800
- 117 232 - 146 - 65 - 194 - 103
- 60 - 50 - 70 - 78 - 78 - 96
- 36 - 104 - 229 - 90 - 259 - 88
3.3 Influence of the electron donor In a following test several carbon sources were used as electron donor. The results are presented in table 3. It is known that e.g. acetate can only be used by some specific SRBs where lactate is mostly degraded into acetate and CO2. So in some cases it is expected that 674
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some fermentative bacteria are degrading the substrates and that finally the produced hydrogen is consumed by the SRBs. This might also be the case by using HRC® (Hydrogen Release Compounds = polylactate ester). This is shown by the difference in metal removal rate in presence of acetate compared to lactate. Already after 6 weeks lactate led to efficient metal precipitation and a Zn concentration below the Flemish groundwater standard of 500 µg Zn/l. Only after 16 weeks acetate led to a 93% removal (from 187,000 µg Zn/l to 12,770 µZn/l). Ethanol and molasses (about 42% of sucrose) showed a similar Zn removal as lactate. The use of polylactate (HRC®) was very efficient although the removal rate was more slowly than lactate. This is due to the slow release behaviour of the HRC® compound. The addition of Postgate C nutrients and yeast extract increased the metal removing rate for lactate and very spectacularly for acetate. This indicates that some fermentation could occur on the yeast extract of the Postgate C medium not only resulting in extra e-donors for the SRB but also in a very fast consumption of possible oxygen concentrations. The use of aquifer and groundwater in a different weight ratio (condition 4 and 11) did not influence the metal removal results indicating that metal toxicity would not be an inhibiting event. It was also observed several times that after a nearly complete metal removal the metal concentration in the groundwater increased after a while, even to concentrations higher than the Flemish standard. This is probably due to the development of a new equilibrium and some re-solubilisation. Table 3. Zn concentration after induction of sulphate reduction with different substrate concentrations Zn (µg Zn/l) 1. Aquifer + groundwater 2. Aquifer + groundwater + HgCl2 3. Aquifer + groundwater + Na-acetate 4. Aquifer + groundwater + Na-lactate 5. Aquifer + groundwater + ethanol 6. Aquifer + groundwater + methanol 7. Aquifer + groundwater + molasse 8. Aquifer + groundwater + HRC® 9. Aquifer + groundwater + PostgateC* + Na-acetate 10. Aquifer + groundwater + PostgateC* + Na-lactate 11. Aquifer + groundwater (80g/100ml) + Na-lactate
T0 207,000 188,000
T6 172,000 173,000
T8 145,900 153,100
T10 153,800 159,600
T13 148,300 148,600
T16 144,200 149,900
187,000
145,000
15,740
22,890
25,320
12,770
199,000
399
610
230
360
< 10
192,000
909
460
240
360
< 10
194,000
478
5670
6,260
7,010
10,240
192,000
227
350
100
330
1,920
181,000
670
380
230
1,130
126,000
15,500
220
150
240
50
84,400
24
60
50
160
890
146,000
149
160
3,100
1,670
< 10
Tx: x is number of weeks of incubation time; * Postgate C without sulphate and carbon source.
3.4 ORP-control Table 4 shows the results of a test done on a groundwater containing several heavy metals and with a pH around 3.3 and a very high ORP of about 400 mV. Some first tests 675
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did not show any metal removal or sulphate reducing activity because the ORP could not be decreased. Therefore several trials were done in order to decrease the redox and to increase the pH. Especially the addition of a redox-manipulating compound (RMC) helped in creating reducing conditions. The addition of a redox-manipulating compound in condition 3 and 6 lead to a pH increase and a decrease in redox potential. However when acetate was added as e-donor the pH increase and ORP decrease were higher than without the extra substrate indicating the influence of microbial (i.e. sulphate reducing) activity. The removal of the metals is presented in figure 1. The figure shows that when aquifer was brought into contact with the groundwater a new equilibrium was formed and resulted in a release of Ni and Zn in the water phase. Neither the redox-manipulating compound (RMC) nor the acetate could induce bioprecipitation reactions. Only the presence of both RMC and acetate could induce the sulphate reduction and metal bioprecipitation process. Ni and Zn were removed to very low concentrations. Also Cu, Cd and Cr could be removed to very low concentrations (results not shown). Table 4. pH and ORP in function of different conditions and redox manipulation Conditions pH 1. Groundwater 2. Aquifer + groundwater + HgCl2 3. Aquifer + groundwater + RMC 4. Aquifer + groundwater + acetate 5. Aquifer + groundwater 6. Aquifer + groundwater + acetate + RMC
T0 3.1 3.3 3.5 3.8 3.3 3.9
T3 3.3 3.4 3.8 3.9 3.2 4.0
T7 3,2 3.5 3.9 3.6 2.0 4.2
T12 3.2 3.4 4.1 3.6 3.2 5.4
T16 3.0 ND 4.4 3.6 3.3 5.4
T26 3.1 3.3 4.4 3.6 2.3 5.5
Conditions ORP 1. Groundwater 2. Aquifer + groundwater + HgCl2 3. Aquifer + groundwater + RMC 4. Aquifer + groundwater + acetate 5. Aquifer + groundwater 6. Aquifer + groundwater + acetate + RMC
T0 440 392 339 294 321 287
T3 455 380 363 336 441 315
T7 402 280 57 215 409 274
T12 262 292 32 280 340 -58
T16 167 ND - 80 188 208 -123
T26 300 329 205 294 320 -12
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Figure. 1. Metal removal by in situ bioprecipiation at low pH en high ORP 3.5 Irreversibility of immobilised metals Table 5 presents the results of a sequential extraction done on the solid residues of some batch tests containing aquifer and groundwater and poisoned by HgCl2 (Control), supplied by molasses (Condition 1) or by methanol (Condition 2). The sequential extraction procedure is described by A. Tessier et al. [12]. Two different extractions (on two identical samples) were done either under anaerobic either under aerobic conditions. The results show that Zn is completely removed from the leachable and exchangeable fraction and nearly completely from the carbonate fraction after induction of the in situ bioprecipitation process. Under aerobic conditions, the metal is also removed from the exchangeable fraction and its concentration in the carbonate fraction represents only 20% of the original concentration. Most of the remaining metal content could be found in this stable Fe-Mn-oxide fraction. However most of the stable metal sulphides would be expected in the organic fraction. In the case of molasses an increase of Zn concentration could be observed in the organic fraction which is probably also due to complexing of metals to the non-fermented residual organic molasses fraction. A decrease in Ni content from the leachable and exchangeable fraction could be observed in both ISBP stimulated conditions. The decrease was not so high as in the case of Zn and a large difference was observed between the molasse and methanol substrates. This indicates that the Ni does not form immediately a very stable precipitate on the aquifer material. Further the precipitates formed with the methanol substrate are more stable than the ones formed with the molasses. Table 5. Ni and Zn concentrations in different fractions after sequential extraction of the aquifer material Control Anaerobic extraction Leachable fraction Exchangeable fraction Carbonate fraction
Molasse
Methanol Zn
Ni
Zn
Ni
Zn
Ni
60.0 145 9.2
161 397 31.5
38,2 172 11,1
0,57 0,62 3,52
7.22 64.3 4.84
0.46 0.91 4.08 677
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Control Aerobic extraction Exchangeable fraction Carbonate fraction Fe-Mn-oxide fraction Organical fraction Residual fraction Total concentration
Molasse
Methanol Zn
Ni
Zn
Ni
Zn
Ni
234 18.4 13,5 4.6 6.5
641 77.7 181 62.9 51
230 20,5 19,1 42,6 8,4
1,56 14,9 510 106 68
82.2 9.37 19.3 61.7 10
1.61 12.7 341 72.7 57
207
810
227
524
213
553
Metal concentrations in mg/kg
4.
CONCLUSIONS The results show that in situ heavy metal immobilisation is an alternative for PRBs composed of compost, limestone or adsorbents. Furthermore, this process known as in situ bioprecipitation (ISBP) can also be used at larger depths where the PRB construction becomes impossible. The process is a clean system with no addition of unknown compounds and in case of saturation another reactive zone can be installed elsewhere. The results show that the concentration of sulphates is important in order to induce the SRBs. In the case of low concentrations (< 200 mg SO42-/l) only hydrogen can be used as electron donor. This volatile substrate can be delivered either in a physical way by gas injection or by providing compounds that are degraded into hydrogen. Several substrates like lactate, ethanol, methanol, and also the slow release compound HRC® can be used. Acetate, not frequently used by sulphate reducing bacteria, shows a lower induction rate which can be increased by adding some other fermentable compounds as yeast extract or lactate (probably also polylactate). Molasses shows always a fast induction of the sulphate reduction process. But the final reached metal concentrations are higher than those obtained with other pure substrates. The substrate concentrations are crucial since too high concentrations will induce methanogenic conditions without any metal removal. Irreversibility tests showed that the metals are removed from the leachable and exchangeable fractions and a shift of the metals to the Fe-Mn-oxide fraction occurred indicating a strong irreversibility of the precipitation process. However in the case of molasses a larger part of the metals stays in the organic fraction due to some complexation processes. Nickel shows a very slow rate of metal immobilisation and also from the irreversibility tests it is shown that it takes more time before Ni is really tightly bound to the Fe-Mn-oxide fraction. Here again a looser binding is observed in the case of molasses. The method can also be used in case of high ORP (e.g. 400 mV) and low pH (e.g. 3). The addition of a redox manipulating allowed the induction of sulphate reducing activity of the SRBs. The principle of in situ bioprecipitation can also be used for the bioreduction of metals as chromate [15] and arsenate (personal communication). The use of several SRB genus specific probes targeting the 16S rDNA indicated in many cases the presence of Desulfosporosinus and Desulfovibrio as SRBs [16]. ACKNOWLEDGEMENTS The financial support of Bekaert nv and Umicore nv is acknowledged. 678
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REFERENCES
1. J. Grimault, J. Ferrer, E. Macpherson (1999) The mine tailing accident in Aznacollar. The Science of the Total Environment 242, 3. 2. L. Diels, M. De Smet, L. Hooyberghs and P. Corbisier, Heavy metals bioremediation of soil. In: Molecular Biotechnology, 12 (1999) 149. 3. T. Pümpel, C. Ebner, B. Pernfuss, F. Schinner, L. Diels, Z. Keszthelyi, A. Stankovic, J. A. Finlay, L.E. Macaskie, M. Tsezos and H. Wouters, Treatment of rinsing water from electroless nickel plating with a biologically active moving-bed sand filter, Hydrometallurgy, 59 (2001) 383. 4. J. Vangronsveld, N. Spelmans, H. Clijsters, D. Van der Lelie, M. Mergeay, P.Corbisier, J. Bierkens, E. Adriaensens, R. Carleer and L. Van Poucke, Physicochemcial and biological evaluation of the efficacy of in situ metal inactivation in contaminated soils. 5. J. Vangronsveld, J. Colpaert, K. Van Tichelen (1996) Reclamation of a bare industrial area contaminated by non-ferrous metals: physico-chemical and biological evaluation of the durability of soil treatment and revegetation. Environ. Pollut. 94, 131. 6. Karczewska, T. Chodak and J. Kaszubkiewicz, The suitability of brown coal as a sorbent for heavy metals in polluted soils, Applied Geochemistry, 11 (1996) 343. 7. S.G. Benner, D. W. Blowes, C. J. Ptacek and K.U. Mayer, Rates of sulphate reduction and metal sulphide precipitation in a permeable reactive barrier, Applied Geochemistry, 17 (2002) 301. 8. D.J.A. Smyth, D.W. Blowes, S.G. Benner and C.J. Ptacek, In situ treatment of metal contaminated groundwater using permeable reactive barriers, In Bioremediation of inorganic compounds, Leeson A. et al. International in situ and on-site bioremediation symposium 6(9) (2001) 71. 9. D. Nuyens, L. Bastiaens, J. Vos, J. Gemoets and L. Diels, Heavy metal in situ bioprecipitation & adsorption on a manufacturing site, In Bioremediation of inorganic compounds, Leeson A. et al. International in situ and on-site bioremediation symposium 6(9) (2001) 87. 10. L. Diels, N. Van der Lelie and L. Bastiaens, New developments in treatment of heavy metals contaminated soils, Re/Views in environmental Science & Biotechnology, 1 (2002) 75. 11. J. Carrera, A. Alcolea, J. Bolzicco, C. Knudby, C. Ayora (2001) An experimental geochemical barrier at Aznalcollar, in Thornton S. & Oswald S. (Eds.) Proceedings of the 3rd International Conference on Groundwater quality, 18-21 June 2001 Sheffield, UK, 407-409. 12. Tessier A., Campbell, T. G. C., Bission M. (1979) Sequential extraction procedure for the speciation of particulate trace metals. Anal. Chem. 51, 844. 13. L. Diels, D. Van der Lelie, J. Gemoets, D. Springael, J. Geets, J. Vos and L. Bastiaens, In situ bioprecipitation of heavy metals in groundwater. IBS2002. 14. J.R. Postgate, The Sulphate reducing bacteria, (1984) Cambridge University Press, Cambridge. 15. J. Gemoets, C. Gielen, N. Hermans, Y. Vermoortel and M. Carpels, (2003) Evaluation of the potential for natural attenuation and in situ bioprecipitation of chromium in groundwater. In Consoil2003 in preparation, 1641. 16. J. Geets, L. Diels, K. Van Geert, E. Ten Brummeler, P. van den Broek, W. Ghyoot, K. Feyaerts and W. Gevaerts 2003) In situ metal bioprecipitation from lab scale to pilot tests. In Consoil 2003 in preparation. 679
15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Removal of chromium(VI) through a two-step process using sulphur-oxidising and sulphate-reducing bacteria E. Donati, M. Viera, G. Curutchet Centro de Investigación y Desarrollo de Fermentaciones Industriales (CINDEFI-CONICET), Facultad de Ciencias Exactas, Universidad Nacional de La Plata, 47 y 115 (1900) La Plata Abstract The environmental and biological effects of chromium are dependent upon its oxidation state. Chromium(III) is essentially non toxic but chromium(VI) is poisonous (except some chromium(VI) complexes) to most living organisms. Most chemical and biological treatments involve the reduction of chromium(VI) to chromium(III). Recently we have proposed the use of a sulphur-oxidising bacterium (Acidithiobacillus thiooxidans) for chromium(VI) reduction and a pool of sulphate-reducing bacteria (Desulfovibrio sp.) for the subsequent chromium(III) precipitation. In this work, we analyse this process in a semi-continuous mode with a first reactor (under aerobic conditions) containing a A. thiooxidans culture using elemental sulphur as the energy source and a second one (under anaerobic conditions) containing Desulfovibrio cells using lactate as the carbon and energy source. A medium containing 10, 15 or 30 mg.l-1 of chromium(VI) was added continuously to the first reactor and the effluent from this reactor, supplemented with lactate, was finally introduced into the second reactor. Chromium(VI) was determined by the diphenylcarbazide method and total chromium concentration was determined using atomic absorption spectrophotometry. Chromium(VI) was almost totally reduced by the action of A. thiooxidans whereas total removal of chromium by the two-step process was higher than 85%. Keywords: Acidithiobacillus thiooxidans, Desulfovibrio, chromium removal 1.
INTRODUCTION Hexavalent chromium was classified as a primary contaminant because of its mobility in soil and groundwater and its reported harmful effects on organisms including humans. Waste-waters containing Cr(VI) are generated by many industrial processes as ore processing, electroplating, leather-tanning processes among others [1-2]. The reduction of toxic Cr(VI) leads to the formation of stable and non-toxic Cr(III); this reduction followed by precipitation or immobilization can be produced by chemical or biological action [3-10]. The oxidation of sulphur by two species of chemoautotrophic acidithiobacilli, Acidithiobacillus ferrooxidans and Acidithiobacillus thiooxidans [11], generates a series of intermediate sulphur compounds with high reducing power [12]. These reducing 681
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compounds can be used in the reduction of Cr(VI) under aerobic (both species) and anaerobic (A. ferrooxidans) conditions [13]. Sulphate-reducing bacteria can be used to remove toxic metals in waters of moderate pH through the formation of insoluble metal sulphides [14]. This group of anaerobic microorganisms uses sulphate as terminal electron acceptor for respiration and organic compounds as carbon/energy sources [15]. There are few reports on Cr(VI) reduction by Desulfovibrio bacteria [10, 16-18] and on Cr(III) precipitation, although in this last case using an undefined culture of sulphate-reducing bacteria [19]. Recently, we have reported a preliminary bioprocess to remove Cr(VI) combining the reduction and subsequent precipitation of 5 mg.l-1 of Cr(VI) [20]. In this work, we analyse this two-step process using two stirred reactors with A. thiooxidans and Desulfovibrio sp. respectively, in a semi-continuous mode at higher Cr(VI) concentrations. 2.
MATERIALS AND METHODS
2.1 Microorganisms The strain of Acidithiobacillus thiooxidans (DMS 11478) was maintained in iron-free 9 K medium [21] with sulphur as energy source, without pH control. Sulphate reducing bacteria Desulfovibrio sp. (ATCC 49975) was maintained in Postgate B medium. The inoculum for the reactor was prepared filtering 10 ml of 15 dayold Postgate B culture and adding to 120 ml of Postgate C medium (without iron). 2.2 Combined process: chromium(VI) reduction and chromium(III) immobilisation A. thiooxidans culture was carried out in a reactor vessel containing 16 g of powdered sulphur and 0.8 l of iron-free 9 K medium (pH 5.1) at 30°C and 400 rpm. pH was maintained through the automatic addition of KOH 1.0 M. After 6 days, a iron-free 9 K medium containing Cr(VI) was added with a peristaltic pump at a flow rate of 150 ml.day1 . Cr(VI) concentrations assayed were 10, 15 and 30 mg.l-1. The effluent of this reactor was filtered through 0.45 µm membrane and was stored in a reservoir. Sterile lactate was added to this reservoir to reach a concentration of 3 g.l-1. Desulfovibrio sp. culture was carried out in a similar reactor vessel containing 0.8 l of Postgate C medium without iron (pH 7.5±0.2) at 30°C and 400 rpm. In order to obtain anaerobic conditions, sterilized N2 was continuously bubbled. pH was maintained with the automatic addition of H 2 SO 4 0.1 M. After 8 days, solution from the reservoir was pumped at a flow rate of 150 ml.day-1 into the Desulfovibrio sp. reactor. The experimental set-up is showed in Figure 1. 2.3 Analytical methods Free (not attached) bacterial population was determined by using a Petroff-Hausser counting chamber in a microscope with a contrast phase attachment. Cr(VI) was determined by the diphenylcarbazide method [22] and total chromium concentration was determined using atomic absorption spectrophotometry. Samples from Desulfovibrio cultures were digested with HNO3 (30 minutes) before chromium determination. Sulphate concentration in cultures and in the reservoir was determined by a turbidimetric method [23]. Lactate was determined with a commercial kit based on the oxidation to piruvate.
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The production of protons by A. thiooxidans was measured by the amount of KOH needed to keep the pH at 5.1. Solid residues were examined under a scanning electron microscope with an energy-dispersive X-ray (EDXA) probe.
Figure 1. Diagram of the experimental set-up. A.t.: Acidithiobacillus reactor; D.sp.: Desulfovibrio reactor 3
RESULTS AND DISCUSSION Figure 2 shows the evolution of bacteria numbers and proton productions in the A. thiooxidans reactor at the three Cr(VI) concentrations assayed. It can be seen that there was a decrease in the number of bacteria and in the production rate of protons when Cr(VI) started to be fed into the reactor.
Figure 2. A. thiooxidans reactor: acid production and free-bacterial population evolutions. Arrows show the beginning of feeding with chromium solution.
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The overall process of sulphur oxidation by bacteria yields sulphate as end product. Although the mechanism of sulphur oxidation by sulphur-oxidising bacteria is not very well known yet, it is widely accepted that the acid production occurs mainly during the first steps of this process by bacteria attached on the sulphur particles. Then, some intermediate compounds such as sulphite and polythionates are oxidised, mainly by free bacteria but these steps are not accompanied by an important acid production [13, 24]. It can be seen in Figure 2 that there was a correlation between the number of free bacteria and the acid production, during the first days of growth. However, in spite of the similar free bacterial populations, the acid production was higher in the reactor fed with 15 mg.l-1 Cr(VI) than in the reactor fed with 10 mg.l-1. This could be related to a higher number of attached bacteria but the reason for this behaviour is not known. In each experiment, it was observed (data not shown) that Cr(VI) concentration slowly increased in the reactor during the first 4-10 days of feeding. This fact suggests that the amount of reducing compounds present in the cultures at the moment of feeding was not enough to completely reduce Cr(VI). After that, a decrease of Cr(VI) concentration and an increase of Cr(III) concentration were observed (data not shown) indicating that more reducing compounds were being produced in the culture. These compounds can only have been produced through bacterial activity showing that there were viable cells in the culture in spite of the fact that the population of free bacteria tended to diminish (figure 2). Besides, there was a small acid production even after 20 days which means that there was bacterial activity probably due to viable cells attached to sulphur particles. The evolution of Desulfovibrio sp. vs. time in the three systems is shown in Figure 3. The number of free bacteria was not affected significantly by the presence of chromium, except in the case of the highest concentration.
Figure 3. Desulfovibrio reactor: Evolution of free bacterial population and sulphate concentration. Arrows show the beginning of feeding
In this step, a decrease in the sulphate concentration was also achieved. The effluents from A. thiooxidans reactor had high concentrations of sulphate (from 8 to 10 g.l-1) and the 684
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final sulphate concentrations were about 3-5 g.l-1. There was also a regulation of the pH, because of the alkalinisation effects of the sulphate-reducing bacteria cultures. Table 1 shows Cr(VI) and Cr(III) concentrations in the inlet and outlet liquids in both reactors. It can be seen that there was no Cr(VI) entering the Desulfovibrio reactor for 10 and 15 mg.l-1 but there was a small amount of Cr(VI) not reduced by the first reactor entering the second one at 30 mg.l-1. In this case, the concentration of Cr(VI) in the effluent from the Desulfovibrio reactor was zero indicating that Cr(VI) was reduced by sulphate-reducing bacteria. In this way, Cr(VI) was completely reduced at all the concentrations assayed. Table 1. Inlet and outlet concentrations of Cr(III) and Cr(VI) in both reactors for the three feeding concentrations assayed (mg.l-1) At: Acidithiobacillus thiooxidans. Dsp: Desulfovibrio sp.
Inlet At reactor Outlet At reactor Inlet Dsp reactor Outlet Dsp. reactor
Experiment 1 (10 mg.l-1) Cr (III) Cr(VI) 0 10.3 4.9 0 4.9 0 1.3 0
Experiment 2 (15 mg.l-1) Cr (III) Cr(VI) 0 15.1 5.5 0 5.5 0 1.1 0
Experiment 3 (30 mg.l-1) Cr (III) Cr(VI) 0 33.4 10.4 1.85 10.4 1.85 3.8 0
Cr(III) was found in the final effluent, at an average concentration of 1 mg.l-1, reaching 3 mg.l-1 when the inlet solution contained 30 mg.l-1 Cr(VI). This could indicate a limitation for this process at concentrations higher than 30 mg.l-1 if a complete removal of chromium, independently its oxidation state, is required. Table 2 shows the percentages of chromium removed in each reactor. Chromium could have been removed in A. thiooxidans reactor by biosorption and/or precipitation of Cr(III) or biosorption of Cr(VI). There are reports on the adsorption of Cr(VI) on biomass [25], but we have found in previous work [13], that chromium adsorbed on Acidithiobacillus exists mainly as Cr(III). The remaining of soluble chromium, fed into the Desulfovibrio reactor, was removed as insoluble Cr(III) compounds, reaching a removal efficiency (in both reactors) higher than 85%. Table 2. Percentages of total chromium removal in both reactors for the three concentrations assayed At: Acidithiobacillus thiooxidans. Dsp: Desulfovibrio sp. Influent Concentration 10 mg.l-1 15 mg.l-1 30 mg.l-1
% Removal in At reactor 51.3 63.5 68.8
% Removal in Dsp reactor 36.6 29.5 17.6
As in previous report [20], EDXA-analysis of solid residues of the Desulfovibrio reactor confirmed that chromium was present mainly as Cr2S3 although chromium(III) phosphates were also found. Comparing both reactors, it can be seen that the percentage of removal reached by A. thiooxidans increased with the inlet concentration, and the opposite occurs with the Desulfovibirio reactor (the higher the inlet concentration the lower the percentage of removal). This could be due to the toxicity of heavy metals to the cells.
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The combination of A. thiooxidans reactor with a second reactor with Desulfovibrio provided a higher efficiency in the removal of chromium and also a reduction in sulphate concentrations and an increase in the pH of the effluent, which can be important in the treatment of some industrial effluents. Sulphate-reducing bacteria have been used to reduce Cr(VI) [10, 16-18]. When the effluent from the A. thiooxidans reactor, having 2 mg.l-1 of Cr(VI) was fed into the second reactor, no Cr(VI) was detected in the outlet (see table 1), indicating that Desulfovibrio sp. was able to reduce Cr(VI). To study this capacity, we fed the second reactor with iron-free 9 K medium supplemented with lactate and 30 mg.l-1 Cr(VI). We found that after 1 day of feeding, the number of bacteria decreased, the consumption of acid needed to keep the pH value in the reactor stopped and Cr(VI) started to accumulate in the culture. Besides, morphological changes in the cells were corroborated by microscopic observations. The redox potential of the medium increased from –240 mV to 150 mV indicating the absence of reducing conditions needed for sulphate-reducing bacteria growth. These results show that Desulfovibrio sp is more susceptible to Cr(VI) than A. thiooxidans, being necessary the introduction of a first step of reduction by sulphur-oxidising bacteria prior to the precipitation of Cr(III) by Desulfovibrio cells. Summarising, combining the action of both bacteria, A. thiooxidans and sulphatereducing bacteria, an effective removal of chromium is achieved as a result of a first step of Cr(VI) reduction and Cr(III) precipitation/adsorption and a second step of residual Cr(III) precipitation. In this way, it was possible to decontaminate higher concentrations of Cr(VI) than that used with one reactor alone. ACKNOWLEDGEMENTS Financial support from the Agencia Nacional de Promoción Científica y Tecnológica is acknowledged. REFERENCES
1. E.N. Lawson. In: Biotechnology Comes of Age, Australian Mineral Foundation, Glenside, 1997, pp. 302-303. 2. V.J. Sundar, J.R. Rao and C. Muralidharan, J. Cleaner Production 10 (2002) 69. 3. R. Melhorn, B. Buchanan and T. Leighton. In: Emerging Technology for Bioremediation of Metals, Lewis Publishers, Boca Raton, 1994, pp. 26-37. 4. J. Rajwade and K. Paknikar. In: Biotechnology Comes of Age, Australian Mineral Foundation, Glenside, 1997, pp. 221-226. 5. P.B. Salunkhe, P.K. Dhakephalkar and K.M. Paknikar, Biotechnol. Lett. 20 (1998) 749. 6. O. Muter, A. Patmalnieks and A. Rapoport, Proc. Biochem. 36 (2001) 963. 7. J.S. Mc Lean, T.J. Beveridge and D. Phillips, Environ. Microbiol. 2 (2000) 611. 8. C.R. Myers, B.P. Carstens, W.E. Antholine and J.M. Myers, J. Appl. Microb. 88 (2000) 98. 9. H. Guha, K. Jayachandran and F. Maurrasse, Environ. Pollution 115 (2001) 209. 10. F. Battaglia-Brunet, S. Foucher, A. Denamur, I. Ignatiadis, C. Michel and D. Morin, J. Ind. Microbiol. Biotechnol. 28 (2002) 154. 11. D.E. Rawlings, Biomining: Theory, Microbes and Industrial Processes, SpringerVerlag, Berlin, 1997. 12. R Steudel, In: H.G. Schlegel and B. Bowien (eds.), Biology of Autotrophic Bacteria, Science Tech. Publ., 1989, pp. 289-303. 686
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13. M. Quintana, G. Curutchet and E. Donati, Biochem. Eng. J. 9 (2001) 11. 14. D. K. Newman, T.J. Beveridge and F.M. Morel, Appl. Environ. Microbiol. 63 (1997) 2022. 15. G. Voordouw, Appl. Environ. Microbiol. 61 (1995) 2813. 16. D.R. Lovley and E.J. Phillips, Appl. Environ. Microbiol. 60 (1994) 726. 17. L. Fude, B. Harris, M.M. Urrutia and T.J. Beveridge, Appl. Environ. Microbiol. 60 (1994) 1525. 18. B.M. Tebo and A.Y. Obraztsova, Fems Microbiol. Lett. 162 (1998) 193. 19. C. White, A.K. Sharman and G. M. Gadd, Nat. Biotechnol. 18 (1998) 572. 20. M. Viera, G. Curutchet and E. Donati, International Biodeteriodation & Biodegradation, in press, 2002. 21. M.P. Silverman and D.G. Lundgren, J. Bacteriol. 77 (1959) 642. 22. P. Urone, Anal. Chem. 27 (1955) 1354. 23. A.E. Greenberg, R. Rhodes Trussell, L.S. Clesceri (eds.) Standard Methods for the Examination of water and wastewater, 16th. Ed. American Public Health Association , Washington, 1985, pp.467-468. 24. K.B. Hallberg, M. Dopson and E.B. Lindström, J. Bacteriol. 178 (1996) 6. 25. T. Srinath, T. Verma, P.W. Ramteke and S.K. Garg, Chemosphere 48 (2002) 427.
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Removal of Mn(II) ions by manganese-oxidizing fungus at neutral pHs in the presence of carbon fiber K. Sasakia, M. Endob, K. Takanoc and H. Konnob a
b
Otaru University of Commerce, Otaru, 047-8501 Japan Graduate School of Engineering, Hokkaido University, Sapporo, 060-8628 Japan c Public Health Institute of Hokkaido, Sapporo, 060-0819 Japan
Abstract A manganese-oxidizing fungus was isolated from Komanoyu hot spring, Hokkaido, Japan, and identified as a type belonging to Phoma. sp. The fungus was found to oxidize Mn(II) ions with decreasing concentration of organic carbon sources in medium, and completely oxidized 60 ppm of Mn(II) ions within 170 hours at pH 7.3 and 25˚C. The presence of carbon fiber shortened the time needed for complete oxidation of Mn(II) ions to half, and made the oxidation reaction steady. Similar shape of PET fiber did not enhance the oxidation, indicating that geometry of fiber is not important but the material exhibits functionality. The results suggest that the carbon fiber does not simply serve as a bed for the microorganisms. During the microbial oxidation of Mn(II) ions under the low concentrations of carbon sources, the colonies did not grow markedly. The recovered Mnproducts were identified to be Mn(IV) species by XPS and poorly crystallized ramsdellite (γ-MnO2) by XRD. Hydrogen peroxide, which was thought to be a metabolic product of the fungus, was detected in the medium within a few hours of oxidation experiment. Time variation of the concentration of hydrogen peroxide was similar to that of Mn(II), and the concentration decreased more quickly in the presence of carbon fiber. It is reported that Mn-peroxidase is produced by the fungus and catalyzes the redox reaction between Mn(III)/Mn(II) and H2O2/H2O, and Mn(III) species eventually decompose to Mn(IV) and Mn(II). There is a possibility that the carbon fiber plays an important role in the catalytic reactions by enzyme. The application of carbon fiber to the treatment of manganese mine drainage by microorganism is promising. Keywords: Mn-oxidizing fungus, carbon fiber, organic carbon source, hydrogen peroxide 1.
INTRODUCTION Treatment of manganese-rich mine water costs high because such water must be treated by the following steps: alkalization to pH higher than 8.5, aeration to oxidize Mn(II) ions to MnO2 precipitate, and neutralization to discharge the treated water into rivers. Accumulation of manganese in the natural environment suggests that the application of biological treatment to Mn-rich mine water is a promising subject. It is an interesting process also in view of the monitoring natural attenuation. Many investigators reported the manganese sediments formed biologically. In such places pH is mostly 6-9 and many kinds of Mn-oxidizing bacteria, such as Leptothrix, Arthrobactor, Bacillus, 689
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Pseudomonas, Micrococcus, and so on, were found and classified by morphologic and genetic investigations [1]. However, Mn-oxidizing fungi have not been investigated so often as bacteria. Exceptionally, white rot and litter-decaying fungi attracted some researchers’ interest in the decomposition of agricultural chemicals as well as Mnoxidation. It is significant and promising to explore the new fungal species having oxidizing activity in the environment of high Mn(II) ion concentration. The fungal cell wall is superior for the accumulation of fine particles and mechanically stronger than that of bacteria. This leads to the advantages in recovery of Mn oxides by fungal oxidation. In addition, biogenic Mn-oxides have significantly higher adsorption capacity for metals and larger surface area than abiotically precipitated Mn oxides [2]. Such kinds of oxides may contribute in controlling the mobility of metals in natural sediments. One of the shortcomings of biological treatment is a slow rate. It has been reported that carbon fiber accelerates the rate of sewage water treatment with activated sludge and is utilized as a "bed". This effect was observed both on a laboratory scale and a small plant scale [3]. Accordingly, it is also significant to investigate if the fiber is useful for the recovery of biogenic Mn-oxides. One of our research groups isolated a new Mn-oxidizing fungus from Mn-rich springs in Hokkaido, Japan. In the present work, the fungal accumulation of Mn oxides from manganese-rich water was investigated in the presence of carbon fiber. 2.
MATERIALS AND METHOD
2.1 Isolation and identification of microorganism Sampling site was Komanoyu, located in the southern Hokkaido prefecture in Japan. Komanoyu is a narrow and shallow watercourse with spring water. The 10 g sediment taken from there was put into 50 cm3 phosphate buffer (6 g Na2HPO4, 4.5 g KH2PO4, 0.5 g sodium thioglycollate in 1 L distilled water, pH 7.0). After shaking it, the turbid water was inoculated on the peptone-yeast extract-glucose (PYG) medium (3.56 g HEPES, 0.5 g peptone, 0.5 g yeast extract, 0.5 g glucose, 0.6 g MgSO4.7H2O, 0.07 g CaCl2, 24 mg MnSO4.5H2O, 15 g agar in 1 L distilled water, pH 7.0). The brownish fungal colony grew on the medium at 20˚C in a week and was transferred to the new PYG medium as a strain. The isolated strain is maintained on the PYG medium at 20˚C. The morphometrical characteristics of the fungus were observed using the microscope. The fungus has the extended hypha with septum and forms a pycnidium in a mature colony. The oval conidium is formed in the pycnidium, and its lump is pushed out from the pycnidium. The sexual cycle was not observed. From these observations the species could not be identified but speculated to belong to Genus Phoma (Coelomycetes). The genetic identification was carried out precisely by the standard and well established methods, and the final sequence alignment (600 bases) was obtained. The closely related species to the fungus were searched from GenBank by BLAST (Altschul et al., 1997). To construct the phylogenic tree, data were analyzed by neighbor-joining method (Saitou and Nei, 1987). The distance matrix was calculated using the method of Jukes and Cantor (1969). Statistical support for clades in distance was estimated using the bootstrap method (Felsenstein, 1985). Phylogenic analysis showed that Ampelomyces hunuli (Coelomycetes), Phoma glomerata (Coelomycetes) and Microsphaeropsis amaranthi (Coelomycetes) were closely related species with the difference of five bases in ITS1 and ITS2 regions. The fungus and 690
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M. amaranthi were formed a branch with the low bootstrap value, 34% supporting the monophyly of the clade. 2.2 Carbon fiber and PET fiber Commercially available carbon fiber (Showa Denko, H-20ST) was used. It is a yarntape of about 2-3 mm wide and the diameter of single fiber is 7-10 µm. The tape was cut into about 30 cm long and tied to a loop to avoid loose threads during experiments. Sizing agents on the fiber surface were removed by heating at 500˚C for 2 h in air. By this treatment, the surface became hydrophilic. Commercially available polyester fiber (Mitsubishi Rayon, D66T136) was also used for comparison. The material is pure polyethylene telephthalate. It is a yarn-type of about 136 single-fibers and the diameter of single fiber is 7 µm. Since the surface was covered with oil, they were washed with ethanol, water, and acetone, and then autoclaved before use. 2.3 Microbially mediated oxidation of Mn(II) ions Since the fungus is unknown species, the experiment was first attempted with the medium for manganese-oxidizing bacteria, Leptothrix discophora. A 500 cm3 Erlenmeyer flask was filled with 100 cm3 of the following medium at initial pH 7.3: MgSO4.7H2O 0.6 g, CaCl2.2H2O 0.07 g, MnSO4.5H2O 0.0746-0.329 g (17-75 ppm as Mn), peptone 0-0.5 g, yeast extract 0-0.5 g, glucose 0-0.5 g, HEPES 3.57 g per one liter of distilled water. This medium is practically the same with the PYG medium used above, but content of three organic carbon sources was changed systematically. Parallel experiments were carried out with and without the sterilized fiber added into the flask. Finally, 0.5 g(wet weight)/100 cm3 of fungus was inoculated into each flask. All cultures were installed in a rotary shaking culture-apparatus TB-16 (Takasaki Kagaku) at 25±2˚C and have been incubated within 240 h under light shielding. A sterilized experiment was also carried out in the presence of NaN3. At intervals, the supernatants were sampled, and filtered by membrane filter of 0.2 µm pore size, then diluted with a hydrochloric acid solution. Dissolved Mn species were determined by atomic absorption spectrometry (Hitachi, Z-6100). The concentration of H2O2 was determined by spectrophotometry [4]. 2.4 Characterization of precipitates formed by microbial treatment Precipitates were analyzed after drying at room temperature by a powder X-ray diffractometer (XRD, RIGAKU Rint-2000 with a monochromator, Cu Ka, 40 kV, 25 mA) and an X-ray fluorescence analyzer (XRF, JEOL JSX-3220Z). The precipitates with carbon fiber were observed by a field emission type scanning electron microscope (FESEM, JEOL JSM-6300F) at an acceleration voltage of 2-3 kV after evaporating a thin platinum layer on the sample. They were also analyzed by XPS (VG Scientific, ESCALAB Mk II). After evacuating to better than 10-5 Pa for 15 minutes, the sample was transferred into an analyzer chamber of better than 5 x 10-8 Pa, and cooled below –150˚C, then irradiated with Mg Kα X ray (14 kV, 20 mA). The binding energies, EB, were calibrated with EB[Au 4f7/2] = 84.0 eV. 3.
RESULTS AND DISCUSSION First, the effect of organic carbon concentration on the removal of Mn(II) ions by the fungus was investigated. Around pH 7.3, the chemical oxidation of Mn(II) ions did not 691
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occur at all. In the PYG medium containing 0.50 g/L of each organic carbon source and 17 ppm of Mn(II) ions, only 4 ppm of Mn(II) ions were oxidized after 150 h. However, by reducing each organic carbon source content to 1/10, 100% of Mn(II) ions were oxidized after 150 h. Clearly a shortage of organic carbon source brings about the oxidation of Mn(II) ions by fungus, suggesting that the fungus begins to oxidize Mn(II) ions after the consumption of organic carbon sources. Under the same conditions, by adding carbon fiber into the medium, 17 ppm of Mn(II) ions were completely oxidized within only 40 h, indicating that the biological Mn(II)-oxidation was markedly enhanced by the carbon fiber. Accordingly, the medium containing 0.050 g/L of each organic carbon was used in the following experiments. Next, the most suitable combination of carbon sources was determined. The best results were obtained with peptone and yeast extract excluding glucose from the PYG medium (hereafter referred to as PY medium), that is, 17 ppm of Mn(II) ions were completely oxidized in 32 h. When the carbon fiber was put into the PY medium, the same concentration of Mn(II) ions was oxidized within only 22 h. The removal of yeast extract from the PYG medium scarcely enhanced the biological oxidation rate of Mn(II) ions, and resulted in a large scattering of the oxidation data. The carbon fiber improved reproducibility of the data but it did not accelerate the rate in this case. It has been reported that the oxidation of Mn(II) ions by some kinds of Mn-oxidizing fungi is achieved by catalytic reactions of extracellular laccase as a terminal electron acceptor of oxygen and of Mn-peroxidase as a terminal electron acceptor of hydrogen peroxide [5]. Since glucose is a reducing agent, exclusion of glucose from the PYG medium might be beneficial to reduce the decomposition of hydrogen peroxide. Yeast extract seemed to be essential to advance the oxidation of Mn(II) steadily. So far the fungus was found to be tolerant to maximum 60 ppm of Mn(II) ions in the PY medium at pH 7.3, and could completely oxidize them within 200 h. Figure 1 shows the oxidation curves of Mn(II) ions with the carbon or PET fiber and without fiber in the PY medium of the initial Mn(II) ion concentration of 50 ppm. The data points are the averages of four runs and error bars are ±1σ. These results clearly demonstrate usefulness of carbon fiber. It is evident that the PET fiber does not enhance the biological oxidation of Mn(II) ions. Accordingly, it is presumed that the geometry of fiber (thickness, for example) is not important but the material (carbon in the present case) exhibits functionality.
Figure 1. Oxidation curves of Mn(II) ions by fungus in the PY medium at pH 7.3 and 25°C 692
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Figure 2(a) and (b) are SEM images of Mn-precipitates formed after the complete oxidation of Mn(II) with fibers. The fungus with Mn-precipitates is entangled on the carbon fiber as shown in Fig. 2(a). In contrast to the carbon fiber, the fungus with Mnprecipitates was isolated from the PET fiber as shown in Fig. 2(b). The adhesion of the fungus and precipitates on the carbon fiber is totally different from the adsorption on porous materials such as activated carbons, since the carbon fiber has the smooth surface and very small specific surface area. It is likely that the surface of carbon fiber has an affinity for extracellular substances from the fungus.
Figure 2. SEM images of Mn precipitates with fungus in the presence of (a) carbon filter and (b) PET fiber. Bars are 10 µm
The products after the 200 h biological oxidation were examined by XPS. Figure 3 shows Mn 3s spectra for the products in the presence and absence of carbon fiber. It is reported that Mn species can be identified from the splitting between a satellite peak (higher binding energy side) and the Mn 3s main peak [6]. The splitting was 4.5 eV for both spectra, indicating that Mn(IV) species are predominant. The Mn-precipitates were also characterized by XRD as shown in Fig. 4. Though the signal to noise ratios are poor, diffraction peaks correspond to ramsdellite (γ-MnO2, JCPDS 42-1316). It is known that ramsdellite often occurs in relatively low temperature environments [7]. From these results, the oxidation products were considered to be a mixture of ramsdellite and amorphous precipitates of Mn(IV) oxides. Mandenack et al. (1995) have reported that spores of marine Bacillus strain, SG-1, directly oxidized Mn(II) to Mn(IV). In many investigations, such as by Schlosser et al. [8], however, it is reported that first Mn(III) complexes are formed by biological oxidation, and then the Mn(III) species disproportionate into Mn(II) and Mn(IV) species eventually. During the fungal oxidation, the precipitates were initially brownish, and then turned blackish. This suggests that the formation of Mn(IV) species in the present work may be effected by the reported oxidation scheme.
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Figure 3. XPS spectra of Mn 3s for the precipitates after 300h oxidation
Figure 4. XRD patterns for the precipitates after 300h oxidation
Several pathways have been proposed for Mn oxidation by microorganisms: they can be operationally classified as the indirect oxidation and the direct oxidation. Some enzymes responsible for Mn oxidation have been identified, purified and characterized. For the Mn oxidation in white rot and litter-decaying fungi, two extracellular enzymes are involved: laccase and Mn-peroxidase. In the reaction catalyzed by laccase and Mnperoxidase, dissolved oxygen and hydrogen peroxide are the terminal electron acceptors, respectively [8]. Figure 5 shows the time-variations of Mn(II) and H2O2 concentration during the biological oxidation of Mn(II) ions by the present fungus. The data points are the averages of two runs. The H2O2 concentration increased quickly within the initial 10 h, and after that H2O2 was consumed with accomplishment of the oxidation of Mn(II) ions. In order to confirm that Mn(II) ions were not chemically oxidized by H2O2, control experiments were carried out with the PY medium containing 50 ppm of Mn(II) and 23 ppm of H2O2 in the presence and absence of the carbon fiber. After 100 h, Mn(II) concentration did not change at all. The H2O2 concentration slightly decreased but it was in a level due to spontaneous decomposition. Consequently, it is concluded that the 694
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oxidation of Mn(II) ions is totally biological. The concentration profiles in Fig. 5 imply that the concentrations of Mn(II) and H2O2 follow a similar trend, that is, the Mn(II) and H2O2 concentrations decrease more quickly in the presence of carbon fiber than in the absence of carbon fiber. This favors the oxidation scheme by Mn-peroxidase but it is not appropriate to conclude it at present. Additional experiments are required to determine which oxidant, dissolved oxygen or H2O2, is mainly involved in the oxidation of Mn(II) ions.
Figure 5. Changes in Mn(II) ions (left) and H2O2 (right) concentrations during biological oxidation of Mn(II) ions by fungus 5.
CONCLUSIONS It was found that the isolated fungus has a high activity of Mn-oxidation, and that the carbon fiber specifically enhanced the fungal oxidation of Mn(II) ions. The enhancement of fungal oxidation of Mn(II) ions by carbon fiber is likely to be effected by the adhesion of fungus on the fiber, which may affect the release of Mn-oxidizing enzymes or the catalytic activity of them. More experiments are necessary to understand such functions of the carbon fiber. ACKNOWLEDGMENTS This work was supported by the Grant-in-Aid for Scientific Research (B) (No. 13555275) from Japan Society for the Promotion of Science. REFERENCES
1. Ehrlich, H. L. (1997) Geomicrobiology of manganese. “Geomicrobiology” (edited by Ehrlich, H. L.), pp.389-489. 2. Nealson Y. M., Lion, L. W., Shuler, M. L., Ghiorse, W. C. (1999) Lead binding to metal oxide and organic phases of natural biofilms. Limnol. Oceanogr., 44, 17151729. 3. Kojima, A., Matsumoto, H., Nagashima, A., Hirano, N. and Otani, S. (1998) Water purification of reservoir with carbon fiber. Proc. Intl. Symp. Carbon, Tokyo, 354-355. 4. Matsubara, C., Takamura, K. (1980) A new spectrophotometric method for the determination of trace of hydrogen peroxide by the titanium (IV)-4-(2-pyridylazo)resorcinol reagent; Application to the assay for hydrogen peroxide as a food additive. Bunsekikagaku, 29, 759-764. 695
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5. Tebo, B. M., Ghiorse, W. C., Waasbergen, L. G., Siering, P. L. and Caspi, R. (1997) Bacterial mediated mineral formation: insights into manganese(II) oxidation from molecular genetic and biochemical studies. "Geomicrobiology: Interactions between microbes and Minerals” (edited by Banfield, J. F. and Nealson, K. H.), pp. 225-266. 6. Junta, J. J. and Hochella, M. F. Jr. (1994) Manganese(II) oxidation at mineral surfaces: A microscopic and spectroscopic study. Geochim. Cosmochim. Acta, 58, 4985-4999. 7. Fritsh S, Post, J. E., Navrotsky, A. (1997) Energetics of low-temperature polymorphs of manganese dioxide and oxyhydroxide. Geochim. Cosmochim. Acta, 61, 2613-2616. 8. D. Schlosser and C. Hofer, Laccase-catalyzed oxidation of Mn2+ in the presence of natural Mn3+ chelators as a novel source of extracellular H2O2 production and its impact on manganese peroxidase. Appl. Environ. Microbiol., 68(7) (2002) 3514.
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Simultaneous removal of oil and heavy metals from wastewaters by means of a permeable reactive barrier V.I. Groudevaa, S.N. Groudevb and A.S. Doychevaa a
Department of Microbiology, Faculty of Biology, University of Sofia, Sofia 1421, Bulgaria b Department of Engineering Geoecology, University of Mining and Geology, Sofia 1700, Bulgaria Abstract In the Dolni Dubnik oil deposit, Northern Bulgaria, waters polluted with crude oil and heavy metals (manganese, zinc, cadmium, lead, copper) were treated by means of a permeable reactive barrier. The waters had a pH in the range of about 5.0-6.4 and contained about 1-4 mg/l oil. The concentrations of heavy metals usually were about 2-3 times higher than the relevant permissible levels for waters intended for use in the agriculture and industry. The barrier was constructed in a ravine, which collected a portion of the polluted waters. The barrier consisted of a mixture of soil, gravel, sand and biodegradable organic substrates (plant and spent mushroom compost, cow manure, sawdust). It had the shape of a turned upside down truncated pyramid and was 1.4 m high, 5.4 m wide at its topside, 3.8 m wide at its bottom side, and about 9 m thick from the front to the back side. The surface of the barrier was covered by abundant vegetation in which Phragmites australis, Scirpus communis and species related to the genera Juncus, Carex and Poa were the most numerous. A consortium of microorganisms related to different physiological groups, including oil-degrading bacteria and fungi as well as sulphatereducing bacteria and other metabolically interdependent microorganisms inhabited the barrier. The water flow varied in the range of about 0.1-0.4 l/s. The treatment of the polluted waters by means of the above-mentioned barrier was very efficient during the different climatic seasons. The oil content in the barrier effluents in most cases was decreased to less than 0.2 mg/l, and the concentrations of heavy metals were decreased below the relevant permissible levels. The removal of oil was connected with its microbial degradation. The removal of the heavy metals was connected mainly with the processes of microbial dissimilatory sulphate reduction and sorption on the organic matter and clay minerals present in the barrier. Keywords: wastewater, permeable reactive barrier, oil, microbial sulphate reduction, sorption of metals 1.
INTRODUCTION In the oil deposit Dolni Dubnik, Northern Bulgaria, oil is recovered through numerous wells, which produce fountains of fluid containing brine and oil. The oil content in the fluid recovered from the different wells varies in the range of about 5-10%. The oil is 697
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light, with a specific gravity of 0.8 g/cm3, rich in paraffines and with a very low content of asphaltene-resinous substances. The brine is characterized by a slightly acidic pH and contains chloride, sulphate, sodium and magnesium ions as main components. However, some toxic heavy metals such as iron, manganese, zinc, cadmium, lead and copper are present in the brine recovered from some wells. The concentrations of these heavy metals in some cases are higher than the relevant permissible levels for waters intended for use in the agriculture and/or industry. The fluid from each well is collected in an individual vessel where the oil is separated from the brine as a result of their different specific gravities. The aqueous phase is removed by siphon from the relevant vessel and is discharged into several ravines located in the deposit. In most cases some oil escapes together with the water from the vessels. Rainwaters are also collected in these ravines where they are mixed with the brine. Natural wetlands characterized by an abundant water and emergent vegetation are located in the ravines. In most cases an efficient removal of the pollutants is carried out during the water flow through the wetlands. The water clean up during the cold winter months at temperatures close to 0°C is not so efficient, especially at the high water flow rates. To avoid, at least partially, the negative temperature effect, in a small ravine, which collected a portion of the abovementioned polluted waters, a permeable reactive barrier was constructed and used to remove the pollutants from these waters. Some data about this clean up process are shown in this paper. 2.
MATERIALS AND METHODS The polluted waters had a pH in the range of about 5.0-6.4 and contained about 1-4 mg/l oil. The concentrations of heavy metals usually were about 2-3 times higher than the relevant permissible levels for waters intended for use in the agriculture and industry. The barrier was constructed perpendicularly to the direction of the water flow in the ravine. The barrier consisted of a mixture of gravel, soil, sand and biodegradable solid organic substrates (plant and spent mushroom compost, cow manure, sawdust). It had the shape of a turned upside down truncated pyramid and was 1.4 m high, 5.4 m wide at its topside, 3.8 m wide at its botton side, and about 9 m thick from the front to the back side. The initial total porosity of the barrier was about 59% but the pore distribution was not uniform. The average filtration coefficient initially was about 10-2 m/s. The barrier was located on the bottom of the ravine, which consisted of an intrusive rock with a very low permeability. The filtration coefficient of this rock was about 7x10–8 m/s. The surface of the barrier was covered by an abundant vegetation in which Phragmites australis, Scirpus communis and species related to the genera Juncus, Carex and Poa were the most numerous. A consortium of microorganisms related to different physiological groups, including oil-degrading bacteria and fungi as well as sulphatereducing bacteria and other metabolically interdepenant microorganisms inhabited the barrier (Table 1). The water flow varied in the range of about 0.1 - 0.4 l/s. The quality of the barrier influents and effluents was monitored at least twice per month in the period June 1996June 2000. The parameters measured in situ included: pH, Eh, dissolved oxygen, total dissolved solids, and temperature. Elemental analysis was done by atomic absorption spectrophotometry and induced coupled plasma spectrophotometry in the laboratory. 698
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Table 1. Microflora of the waters before and during their treatment in the permeable reactive barrier Microorganisms
Before treatment
Aerobic heterotrophic bacteria Aerobic cellulose-degrading microorganisms Aerobic oil-degrading microorganisms S2O32- -oxidizing chemolithotrophs (at pH 7) Fe2+-oxidizing heterotrophs (at pH 7) Mn2+-oxidizing heterotrophs (at pH 7) Anaerobic heterotrophic bacteria Anaerobic cellulose-degrading microorganisms Anaerobic oil-degrading microorganisms Sulphate-reducing bacteria
103-105 101 -102 101-104 101-103 101-104 101-104 102-104 < 101 101-102 101-103
During the treatment in the permeable barrier Cells/ml 103-107 103-106 103-106 101-103 102-106 101-106 104-107 102-105 103-105 104-107
The hydrocarbons were determined by direct extraction from the relevant water sample with 1,1,2-trichlorotrifluoroethane and IR determination. The mobility of the pollutants was determined by the sequential extraction procedure (Tessier et al., 1979). The isolation, identification and enumeration of microorganisms were carried out by methods described elsewhere (Karavaiko et al., 1988; Widdel and Hansen, 1991; Widdel and Bak, 1991; Groudeva et al., 1993). 3.
RESULTS AND DISCUSSION The treatment of the polluted waters by means of the permeable reactive barrier was very efficient during the warmer months of the year (from March to November) when the temperatures inside the barrier exceeded 10°C and during the summer (June-August) usually were higher than 20°C (Tables 2 and 3). The oil content in the barrier effluents was decreased to less than 0.2 mg/l. The removal of oil was connected with its microbial degradation. The biodegradation of different alkanes was very efficient and even the polyaromatic hydrocarbons were removed to a significant extent (Table 4). The number of oil-degrading microorganisms in the barrier exceeded 108 cells/g solid matter (plant compost or soil). Bacteria related to the genera Pseudomonas, Corynebacterium and Arthrobacter were the most active oil-degrading microorganisms. The removal of heavy metals was also very efficient. The removal of non-ferrous metals (zinc, cadmium, lead and copper) was connected mainly with the process of the microbial dissimilatory sulphate reduction. This conclusion was based on several data about the efficient removal of sulphates and the steady increase in the content of the relevant sulphides of the above-mentioned metals in the barrier. The low redox potentials (lower than minus 100 mV) and the presence of suitable electron donors (dissolved organic monomers) within the barrier facilitated the growth of the sulphate-reducing bacteria. These bacteria were a quite numerous and diverse population in this system (Table 5). Most of them were firmly attached to the solid surfaces in the barrier and exceeded 108 cells/g solids. A part of the iron was also removed as the relevant sulphide (FeS).
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Table 2. Data about the waters before and after their treatment by the permeable reactive barrier Parameters
Before treatment
After treatment
Temperature, °C pH Eh, mV Dissolved O2, mg/l Total dissolved solids, mg/l Oil, mg/l Sulphates, mg/l Iron, mg/l Manganese, mg/l Zinc, mg/l Cadmium, mg/l Lead, mg/l Copper, mg/l
(+0.1) - (+22.1) 5.0 - 6.4 (+95) - (+215) 1.2 - 2.8 1070 - 2453 0.9 - 3.8 415 - 871 9.5 - 25 0.71 - 2.95 8.2 - 21.5 0.02 - 0.08 0.20 - 0.91 0.14 - 1.25
(+0.8) - (+24.8) 7.1 - 7.5 (+35) - (+140) 0.5 - 1.0 640 - 1405 < 0.2 -1.0 204 - 374 0.2 - 1.2 < 0.5 0.15 - 1.0 < 0.01 < 0.1 < 0.1
Permissible levels for waters used in agriculture and industry 6-9 2 1500 0.2 400 5 0.8 10 0.02 0.2 0.5
Table 3. Removal of pollutants in the permeable reactive barrier Pollutant Oil Iron Manganese Zinc Cadmium Lead Copper
Pollutant removed, g/24h During the warmer months During the cold winter months (at 0-5°C) 41 - 99 15 - 37 170 - 347 107 - 198 9 - 28 5 - 17 109 - 304 71 - 181 0.2 - 0.9 0.1 - 0.6 2.3 - 9.5 1.4 - 4.6 1.7 - 11.3 1.0 - 6.0
Table 4. Biodegradation of the different oil hydrocarbons during the water treatment in the permeable reactive barrier Hydrocarbons Alkanes Iso-alkanes Cycloalkanes Monoaromatics Polyaromatics Total
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Content in the oil, % Before treatment After treatment 8.2 - 12.0 0.7 - 1.4 57.2 - 64.0 43.1 - 52.3 10.9 - 13.4 6.8 - 12.7 7.0 - 9.1 7.7 - 12.5 7.1 - 12.5 24.4 - 35.0 100 100
Hydrocarbon degradation, % > 98 73 - 92 75 - 93 64 - 87 23 - 68 73 - 92
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Table 5. Sulphate-reducing bacteria in the permeable reactive barrier Sulphate-reducing bacteria Desulfovibrio (D. desulfuricans, D. vulgaris, D. saprovorans) Desufobulbus (D. elongatus, D. propionicus) Desulfobacter (D. multivorans) Desulfobacterium (D. autotrophicum, D. vacuolatum) Desulfococcus (D. postgatei) Desulfotomaculum (D. nigrificans, D. orientis) Desulfosarcina (D. variabilis)
Cell/ml 105 – 107 102 – 105 103 – 106 102 – 104 101 – 103 101 – 102 101 – 104
The removal of manganese was connected with the microbial oxidation of the bivalent manganese to the tetravalent state. The Mn4+ was then precipitated mainly as MnO2. The bacteria able to oxidize the bivalent manganese were present mainly in the relatively richin-oxygen zones in the barrier (near to the surface and in the root systems of the plants grown on the barrier). In some effluent samples the number of these bacteria exceeded 106 cells/ml. Most of these bacteria were able to oxidize also the ferrous iron to the ferric state. The Fe3+ was then precipitated as Fe(OH)3. The prevalent microorganisms in this physiological group were related to the genera Metallogenium, Sphaerotilus and Crenothrix. It was found that portions of the heavy metals were removed by their sorption on the plant and microbial biomass and on the clay minerals in the barrier. However, the relative portions of easily leachable fractions (exchangeable and carbonate) of these metals were small and together were always less than 20% from the total content of the respective metal. Some of the living plants grown on the barrier were able to accumulate heavy metals, mainly in their roots. The living vegetation on the barrier facilitated the water clean-up process also by secreting soluble organic compounds and oxygen into the barrier. The dead plant biomass, apart from its sorption capacity towards the heavy metals, was used as a source of soluble organic compounds released as a result of the activity of the indigenous cellulose-degrading microorganisms. The removal of pollutants markedly depended on the temperature but was efficient even during the cold winter months (December-February) when the temperatures inside the barriers were often close to 0°C (but were markedly higher than the temperatures outside the barrier). Under such conditions, the microbial activity was negligible and the sorption of heavy metals on the dead plant biomass and the clays in the barrier was the main mechanism involved in the water clean up. The total content of non-ferrous metals in the dead plant biomass varied in the range of 1070-3740 mg/kg dry biomass, and in some clay samples exceeded 10 g/kg dry clay. These non-ferrous metals were present mainly as the easily leachable fractions (exchangeable and carbonate) of the relevant metal. Oil also was retained in the barrier. The residence times during the treatment varied from about 15 to 60 hours depending mainly on the water flow rate and the temperature. Longer residence times were needed during the cold months reflecting the lower clean up potential of the barrier. In any case, the real residence times for the oil degradation were longer than those evaluated on the basis of the measured water flow rates because a significant part of the oil was retained in the barrier and then was degraded by the microorganisms. The filtration coefficient of the barrier steadily decreased during the treatment. The content of dissolved organic substances released from the biodegradable solid substrates (dead plant biomass) also decreased in the course of time. For that reason, the barrier 701
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material was replaced by fresh material of the same type twice during the experimental period (in May 1997 and April 1998). The permeable reactive barrier used in this study was very similar to the well-known anoxic sulphate-reducing and alkalinity-producing cells (Cambridge, 1995; Groudeva and Groudev, 1998). However, this barrier possessed also an efficient aerobic zone located in the top layers near to the surface. The roots of the plants growing on this surface increased the oxygen content in this zone. The barrier can be used in combinations with other units of the type of the passive systems, such as natural and constructed wetlands and/or rock filters. Under certain conditions, especially at relatively higher water flow rates, the abundant vegetation on the barrier surface can act as a typical wetland and in this way to increase the efficiency of the water clean up. ACKNOWLEDGMENTS The National Science Fund (Research Contract CC-904/99) funded a part of this work. REFERENCES
1. Cambrige, M. Use of passive systems for the treatment of mine outflows and seepage, Minerals Industry International, May (1995) 35-42 2. Groudeva, V.I. and Groudev, S.N. Cleaning of acid mine drainage waters from a uranium mine by means of a passive treatment system, Mineral Processing and Extractive Metallurgy Review, 19 (1998) 89-95. 3. Groudeva, V.I., Ivanova I.A., Groudev, S.N. and Uzunov, G.C. Enhanced oil recovery by stimulating the activity of the indigenous microflora of oil reservoirs. In: A.E.Torma, M.L.Apel and C.L.Brierley (Eds), Biohydrometallurgical Technologies, vol. II, 349 - 356, TMS, The Minerals, Metals & Materials Society, Warrendale, PA. 1993. 4. Karavaiko, G.I., Rossi, G., Agate, A.D., Groudev, S.N., Avakyan, Z.A. Biogeotechnology of Metals. Manual, Center for International Projects GKNT, Moscow. 1988. 5. Tessier, A., Campbell, P.G.C. and Bisson, M. Sequential extraction procedure for speciation of particulate trace metals, Analytical Chemistry, 51 (7) (1979) 844-851. 6. Widdel, F. and Bak, E. Gram-negative mesophilic sulphate-reducing bacteria. In: A.Balows, H. G.Truper, M.Dworkin, W.Harder and K. H.Schleifer (Eds), The Prokaryotes, vol. IV, 3352 - 3378, Springer, New York, NY. 1991. 7. Widdel, F. and Hansen, T.A. The dissimilatory sulphate and sulfur-reducing bacteria. In: A.Balows, H. -G.Truper, M.Dworkin, W.Harder and K. -H.Schleifer (Eds.) The Prokaryotes, vol. I, 583 - 624, Springer, New York, NY. 1991.
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
The exploitation of sulphate-reducing bacteria for the reclamation of calcium sulphate sludges Alena Luptakovaa*, Maria Kusnierovaa, Maria Bezovskaa and Peter Feckob a
Department of Mineral Biotechnology, Institute of Geotechnics of Slovak Academy of Sciences, Watsonova 45, 043 53 Kosice, Slovak Republic b Institute Environmental Engineering, VSB-Technical University Ostrava, Mining and Geological Faculty, 17. listopadu 15, 708 33 Ostrava-Poruba, Czech Republic Abstract Combustion of fossil fuels containing sulphur releases sulphur oxides to the atmosphere. If combustion gases are desulphurized by lime or limestone scrubbing, calcium sulphates sludges are generated and these must be disposed of. Under appropriate conditions these sulphate can be converted to sulphide by bacterial sulphate reduction as hydrogen sulphide gas, which can be oxidized chemically or biologically to elemental sulphur. Because bacterial sulphate reduction by sulphate-reducing bacteria has widespread environmental effects and this process has potential as a treatment process for calcium sulphate sludge and acid mine drainage, the aim of this work was to study the gypsum biodegradation from the final stored product – "stabilizate", from deposit thermal power plant Vojany in Slovakia. The process may consist of three stages: anaerobic sulphate reduction to hydrogen sulphide by sulphate-reducing bacteria, metals precipitation from acid mine drainage by continuous stripping of hydrogen sulphide and the excess hydrogen sulphide oxidation to elemental sulphur. Metal sulphides, calcium carbonate and elemental sulphur are final products and these may by recycled in chemical or hydrometallurgical processes. 1.
INTRODUCTION The populations of different species of microorganisms (MO) are able to transform organic substances into inorganic ones and vice versa supporting the circulation of elements in nature. In principle it is the course of basic metabolic procedures of MO by which they obtain energy and nutrients they need for their growth, movement and reproduction. They use various types of oxidation reactions which are always connected with the reduction ones. Under certain conditions in nature it is possible to find MO which reduce substances formed by oxidation procured by other MO species. A good example of this is the circulation of sulphur and its compounds in the biosphere which is considered as one of the basic biological systems on the Earth and points out the environmental
* Corresponding author, E-mail:
[email protected] The authors are grateful to the Slovak Grant Agency for Science (Grant No. 2210622) for the financial support of this work.
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significance of sulphate-reducing bacteria (SRB) which have irreplaceable role from the point of view of life existence on the Earth. The biological circulation of sulphur in nature consists of assimilation and dissimilation. SRB participate in the dissimilation part of the cycle and the microbial population of the given part is called sulphuretum [1]. The activity of sulphuretum is a basis of many processes in nature (both positive and negative) and it finds its application in industry. The SRB represent a group of chemoorganotrophic and strictly anaerobic bacteria that may be divided into four groups based on rRNA sequence analysis [2]: Gram-negative mesophilic SRB, Gram-positive spore forming SRB, thermophilic bacterial SRB and thermophilic archaeal SRB. They include genera like Desulfovibrio, Desulfomicrobium, Desulfobacter, Desulfosarcina, Desulfotomacullum, Thermodesulfobacterium, Archaeoglobus, etc. The basic metabolic process of SRB is the anaerobic reduction of sulphates in which organic substrate (lactate, malate, etc.) or gaseous hydrogen is the electron donor and sulphate is the electron acceptor. Considering the inorganic or organic character of energy source of SRB there are two types of anaerobic respiration of sulphates [1]: 1. Authotrophic reduction of sulphates – the energy source is a gaseous hydrogen, the reaction proceeds in several stages and the whole process can be summarily expressed by equation (1) : SRB
4 H2 + SO42-⎯⎯⎯→ S2- + 4 H2O (1) 2. Heterotrophic reduction of sulphates – the energy sources are simple organic substances (lactate, fumarate, pyruvate, some alcohols and the like). Depending on the final product of organic substrate oxidation we know: - incomplete heterotrophic oxidation of organic substrate in which the final product is acetate (equation 2): SRB
2CH3CHOHCOO- + SO42- ⎯⎯⎯→ 2CH3COO- + 2HCO3- + H2S (2) - complete heterotrophic oxidation of organic substrate in which the final products are CO2 and H2O (equation 3): SRB
4CH3COCOONa + 5MgSO4 ⎯⎯⎯→ 5MgCO3 + 2Na2CO3 + 5H2S + 5CO2 + H2O (3) In the process of anaerobic respiration of sulphates SRB produce a considerable quantity of gaseous hydrogen sulphide (H2S) which reacts easily in the water medium with heavy metal cations forming not easily dissoluble sulphides of the given metals (equation 4): (4) Me2+ + H2S ⎯⎯⎯→ MeS + 2H+ (Me2+ - metal cation) The excess hydrogen sulphide can be further used to produce elemental sulphur as expressed by equations (5) and (6): 2Fe3+ + H2S⎯⎯⎯⎯→2Fe2+ + S° + 2H+ (chemical oxidation of hydrogen sulphide) (5) Chromatium vinosum
2H2S + CO2⎯⎯⎯⎯→2S° + C + 2H2O (biological oxidation of hydrogen sulphide) (6) light
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The issue of formation and processing of liquid or solid wastes containing high concentrations of sulphates is constantly actual. The typical representatives of these wastes are e.g. acid mine drainage (AMD); coalmine drainage; sulphate waste waters or sludges produced by chemical, textile, pharmaceutical, paper or metallurgic industries. The above-listed equations (1) to(4) show the possibility of using SRB to solve this issue, namely: • in the removal of sulphates and heavy metals from AMD [3, 4], • in the removal of sulphates (desulphatization) from industrial waste waters [5], • in the production of sulphur and sulphuric acid from waste waters [6]. From the study of natural conditions of SRB existence it is known that one of sulphate sources important for SRB growth is gypsum – CaSO2.2H2O. The industrial technologies for the desulphurization of combustion products produced during the generation of electric energy by combustion of fossil fuels use limestone (CaCO3) as an absorption agent. The desulphurization of combustion products proceeds in the absorber in several stages. This process results in the formation of gypsum suspension which is incorporated into the final stored product – "stabilizate" [7] - after being treated together with other wastes (ash, burnt lime, desulphurization waste water, etc.). The objective of our study was to verify experimentally the possibility of using gypsum contained in the above-mentioned „stabilizate“ as the source of sulphate for the cultivation of SRB with the prospect of: • recycling of desulphurization agent – limestone, • production of elemental sulphur from hydrogen sulphide. 2.
MATERIALS AND METHODS
2.1 Microorganisms A culture of SRB (genera Desulfovibrio and Desulfotomaculum) was obtained from drinking mineral water Gajdovka (locality Kosice-north, Slovak Republic). For the isolation and cultivation of these bacteria a selective nutrient medium (DSM-63 – Postgate's C medium) [8] was used. The growth of SRB was detected by the formation of black precipitates at the bottom of the flasks and flask walls. Bergey's Manual of Determinative Bacteriology was used for identification of these bacteria. 2.2 Feed solution The feed solution was prepared by dissolving analytical grade salts such as: K2HPO4 0.5 g/l, NH4Cl 1 g/l, CaCl2.6H2O 0.1 g/l, MgCl2.6H2O 0.3 g/l, C3H5O3Na 2.0 g/l, FeSO4.7H2O 0.05 g/l, C2H3O2SNa 0.1 g/l and C6H8O6 0.1 g/l in distilled water. 2.3 Solid phase The sample of "stabilizate" from Vojany power plant (Slovak Republic) was used in the experiments. The chemical composition was: CaSO4 40.84%, SiO2 22.70%, Al2O3 10.70%, Fe2O3 4.26%, CaO 3.00%, loss by ignition 18.50%. 2.4 Analytical procedures A turbidimetric method was used to measure the concentration of soluble sulphate ion concentrations [9] in the liquid phase. Sulphates form an insoluble precipitate with barium (BaCl2) under acidic conditions. The absorbance of the sample was measured at a
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wavelength of 420 nm using Spectromom195 (Hungary) instrument. Samples were centrifuged for 10 minutes at 10 000 rpm before performing the analysis. The glass pH electrode combined with the reference Ag/AgCl electrode and a platinum redox plus Ag/AgCl reference electrode was used to measure pH. Digital pHmeter GPRT 144 AGL (Germany) was used. The presence of hydrogen sulphide was determined using an orientation test [10]. The principle of this test is the reaction Cu2+ with hydrogen sulphide under acidic conditions according to the following equation (7): (7) Cu2+ + H2S ⎯⎯→ CuS + 2H+ The intensive brown colour is proportional to the hydrogen sulphide volume. Qualitative changes of "stabilizate" were performed by the qualitative X-ray diffraction analysis using Dron-2 instrument. 2.5 Biological utilization of gypsum from "stabilizate"
Series of anaerobic tests were studied in a fed batch reactor in the thermostat at 30°C. Samples of "stabilizate" were kept in static conditions for a period of 40 days at pH 7.5. The weight of "stabilizate" was 20g. The stock culture of SRB was used as an inoculum (10% v/v). The total volume of feed solution consisted of 200 ml distilled water and 300 ml selective nutrient medium for SRB (DSM-63 – Postgate's C medium without sulphates). The abiotic control was carried out without the SRB application at the same conditions. After 40 days the solid phase was filtered, dried and analysed using the qualitative X-ray analysis. 3.
RESULTS AND DISCUSION The black precipitates creation, positive results of the orientation test for the determination of the presence hydrogen sulphide and the sensorial detection of classical strong H2S smell were observed after 3 – 4 days from beginning of the process. These remarks were not detected in the abiotic control until the end of the experiment. The sulphate concentration and pH values at the beginning and end of tests are shown in Table 1: Table 1. Changes the sulphate concentration and pH values
"Stabilizate" with SRB "Stabilizate" without SRB (abiotic control)
SO42- (g/l) Start End 0.08 0.40 0.08 2.36
pH Start 7.50 7.50
End 7.44 8.58
The results of qualitative X-ray analysis of original "stabilizate", bacterially treated "stabilizate" and "stabilizate" of abiotic control are shown in Figures 1-3. They show the significant qualitative changes in the "stabilizate" initiated by SRB. The sulphates from the original major component of "stabilizate" – CaSO4.2H2O (Figure 1) - were reduced according to equations (1), (2) and (3) by SRB forming hydrogen sulphide as proved by the visual change in the color of liquid phase and orientation test on the formation of hydrogen sulphide. This is indirectly confirmed by Figure 2, which proves the extinction of CaSO4 and formation of CaCO3 according to equation (8): SRB
2C3H5O3Na+CaSO4 ⎯⎯⎯⎯⎯→ 2CH3COONa+CO2+CaCO3+H2S+H2O 706
(8)
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The above-mentioned changes have not been observed in abiotic control as documented by the comparison of qualitative X-ray analysis of "stabilizate" before applying SRB (Figure 1) with the qualitative X-ray analysis of "stabilizate" in abiotic control (Figure 3).
Figure 1. Qualitative X-ray analysis of "stabilizate" before application of sulphatereducing bacteria
Figure 2. Qualitative X-ray analysis of solid phase formed by the effect of SRB on "stabilizate"
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Figure 3. Qualitative X-ray analysis of solid phase in abiotic control 4.
CONCLUSION The results confirmed the theoretical assumptions on the use of gypsum, which forms the substantial component of "stabilizate", as the source of sulphate for sulphate-reducing bacteria, which produce hydrogen sulphide in the process of bacterial reduction of sulphates. They also showed the possibility of recycling desulphurization agent-limestone, as well as the realistic alternative of using "stabilizate" in the production of elemental sulphur which still represents an important raw material needed in chemical, paper or other industries. On the basis of the complex evaluation of theoretical and practical knowledge obtained up till now and published in the previous studies, it is possible to propose the following method of processing the „stabilizate“ with the objective of producing elemental sulphur (Fig. 4):
Figure 4. Proposal of "stabilizate" processing method using sulphate-reducing bacteria, 1 – organic substrate and mineral substances (nutrient medium) 2 – inoculation of sulphate-reducing bacteria
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REFERENCES
1. J.M. Odom and J.R. Rivers Singleton (eds.), The sulphate-reducing bacteria Contemporary Perspectives, Springer-Verlag, New York, 1993. 2. H.F. Castro, N.H. Williams and A. Ogram, FEMS Microbiology, 31 (2000) 1-9. 3. A. Kontopoulos, In: S.H. Castro, F. Vergara and M.A. Sánchez (eds.) Effluent Treatment in the Mining Industry University of Conception, Chile, 1988, 57-112. 4. S. Foucher, F. Battaglia-Brunet, I. Ignatiadis and D. Morin, Chemical Engineering Science, Vol. 56, Issue 4 (2001) 1639-1645. 5. D. J. Cork and M. A. Cusanovich, Develop. Industr. Microbiol., 20 (1979), 1024. 6. J. Postgate (ed.), Microbes and Man, England, Middlesex, 1975. 7. M. Jeleňová, Diploma work, TU – BERG, Košice, 2000. 8. G.I. Karavaiko, G. Rossi, A.D. Agate, S.N. Groudev, Z.A. Avakyan (eds.), Biotechnology of metals, Centre of projects GKNT, Moscow, 1988, 59-61. 9. APHA, Standard Methods for the Examination of Water and Wastewater, 17th edition, American Public Health Association, USA, Washington D. C., 1989. 10. A.Luptáková, PhD. Thesis, IG SAS, Košice, 2002.
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
The role of metal-organic complexes in the treatment of chromium containing effluents in biological reactors E. Remoudaki, A. Hatzikioseyian, F. Kaltsa and M. Tsezos School of Mining and Metallurgical Engineering, Heroon Polytechniou 9, 157 80 Zografou, Athens, Greece Tel: +30 210 772 2271 / + 30 210 772 2172 Fax: +30 210 772 2173 e-mail:
[email protected],
[email protected],
[email protected] Abstract Industrial effluents containing Cr(VI) introduce significant toxicity in the environment. Treatment purposes are (i) the reduction of toxic chromium (VI) to the less toxic Chromium(III) and (ii) the precipitation of trivalent chromium. Metabolically mediated reduction of hexavalent chromium by microbial biomass is now well documented and, if applied successfully in an effluent treatment scheme, offers an efficient and low cost alternative avoiding the consumption of chemicals and energy. Positive results of complete biological reduction of hexavalent chromium are reported in the literature from the successful operation of pilot scale biological reactors. Organic compounds are often simultaneously present with chromium, in the solution under treatment, having three possible origins: (i) organic compounds co-existing with chromium in the effluent (e.g. leather tanning), (ii) organic compounds metabolically produced by the microbial biomass used for chromium (VI) reduction (EPS and other organic molecules), (iii) excess of nutrients added in the system to support microbial growth during treatment. The formation of complexes between these organic compounds and trivalent chromium alters the solubility behavior of Cr(III) inhibiting the precipitation of chromium. In this paper, the results of a systematic experimental study of the solubility of trivalent chromium in the presence of selected organic compounds, representative of the above three possible origins (organic acids, amino-acids, proteins and biomass growth nutrients) are presented. The organic compound concentration threshold, above which the solubility behavior of trivalent chromium becomes different from that reported from simple chromium (III) aqueous solutions, was experimentally determined for each one of the compounds tested. Keywords: metal-organic complexes, chromium, solubility curves 1.
INTRODUCTION Treatment purposes of industrial effluents containing Cr(VI) are (i) the reduction of toxic chromium (VI) to the less toxic Chromium(III) and (ii) the precipitation of trivalent chromium. Biological reduction of Cr(VI) is an attractive alternative as it can be efficient and low cost avoiding the consumption of chemicals and energy. 711
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Positive results of complete biological reduction of hexavalent chromium are reported in the literature [1-6], from the successful operation of pilot scale biological reactors by various microbial strains. Hexavalent chromium feed in such biological reactors, is quantitatively reduced to Cr(III) by microbial biomass, used as alternative electron terminal acceptor or by reductive enzyme activity induced by the cells [5]. Usually, the pH of the effluent is well within the range of the minimum Cr(III) solubility (i.e. from 6.5 to 9.5) and precipitation of insoluble Cr(III) species would be expected. However, the precipitation of trivalent chromium is not always observed [1]. This is probably due to the fact that the composition of the solution induces significant modifications in the classic Cr(III) solubility behavior. Organic compounds are often simultaneously present with chromium, in the solution under treatment, having three possible origins: (i) organic compounds co-existing with chromium in the effluent (e.g. leather tanning [7,8]), (ii) organic compounds metabolically produced by the microbial biomass used for chromium (VI) reduction (EPS and other organic molecules), (iii) excess of nutrients added in the system to support microbial growth during treatment. Three possible situations of chromium/organic content could be met: i. Cr (III) concentration in solution is comparable or in excess of organic compounds [2] ii. Organic compounds concentration is in excess of Cr(III) although the microbial activity [1] iii. Organic compounds are in excess of Cr(III), but if microbial degradation of the organic molecules occurs, the situation is progressively shifted to case (i) This paper investigates the effect of selected organic molecules of biological origin / importance in the precipitation characteristics of trivalent chromium. Selected cases are presented explaining the solubility behavior of chromium. 2.
MATERIALS AND METHODS Stock standard solutions of trivalent chromium were prepared at a concentration of 1000 mg/l by dissolving the appropriate amount of Cr(NO3)3.9H2O in ultra pure water. The solutions were acidified with pure HNO3 to avoid metal precipitation. All working solutions of trivalent chromium were at concentration of 10 mg/l. 2.1 Determination of Chromium (III) solubility as a function of pH The Chromium solubility study was carried out by preparing solutions of chromium at different initial pH values ranging between 3 and 12. The solutions pH was adjusted by dropwise addition of 0.1N HNO3 or NaOH appropriately. The solutions were left to reach equilibrium for 24 hours under continuous agitation in a rotary shaker. At the end of that period the solutions were vacuum filtered through a 0.45 µm Milipore membrane. The pH of the filtrate was measured and reported as pH equilibrium value in the Cr solubility diagrams. The concentration of Chromium in the filtrate was determined by atomic absorption spectroscopy (AAS) and represents the fraction of Chromium, which remains soluble at the corresponding pH. This experimental procedure was followed in all the solubility experiments presented in this paper.
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2.2 Determination of Chromium (III) solubility as a function of pH in the presence of organic compounds The solubility curves of Chromium (III) in the presence of three groups of pure organic compounds and a mixture of microbial nutrients were experimentally determined at different pH values. Table 1 presents the organic compounds selected. Pure organic compounds of analytical grade with known molecular formula weight (organic acid, amino acid) were used in a Cr(III) to organic compound molecular ratio of 1/1, 1/10 and 1/100. In this way, sets of three experimental solubility curves were determined for each one of the compounds tested. A commercial protein reagent casein and a microbial nutrient nutrient broth which is a mixture of organic compounds with have not well determined composition and defined molecular weight. In these cases, Cr(III) to organic compound mass ratio of 1/1, 1/10 and 1/100 was used. All the solutions of trivalent chromium were at initial concentration of 10 mg/l. This concentration was choosen as a typical value for a medium strength Cr(III) containing wastewater. In addition if Cr(III) becomes from a complete microbial reduction of Cr(VI), the concentration of 10 mg/l Cr(VI) is tolerable by the microbial biomass in a bioreactor [1]. Table 1. The organic compounds and the microbial nutrient used for the experimental determination of Cr(III) solubility curves
3.
Group of organic compounds
Compound
Organic acids Amino acids Proteins Microbial nutrient media
Acetate Alanine Casein Nutrient broth
RESULTS AND DISCUSSION
3.1 Determination of Chromium (III) solubility as a function of pH Figure 1 presents the experimental data of Cr(III) solubility in aqueous solutions at initial concentration of 10 mg/l, as a function of pH. The shape and the position of the curve match well to those proposed in the literature, [9].
Figure 1. Experimental determination of Cr(III) solubility as a function of pH 713
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As expected, in acidic pH values up to pH 4, Cr(III) remains quantitatively soluble. From pH 6.5 to 9.5 Cr(III) forms insoluble hydroxide species. In highly alkaline solutions Cr(III) is resolubilized. 3.2 Determination of Chromium (III) solubility as a function of pH in the presence of organic molecules and nutrient media Figure 2 presents the experimental determination of Cr(III) solubility as a function of pH in the presence of acetate. The solubility curves have been determined for three different acetate to chromium mole ratios, 1/1, 10/1 and 100/1. From this figure it is apparent that: i. For equimolar chromium and acetate solution concentrations, the solubility curve is similar to that presented in Figure 1. This means that at acetate concentration of 0.2 mM the solubility of Cr(III) remains unaffected. ii. When acetate concentration increases 10 fold, the solubility curve of Cr(III) is slightly shifted to the right, indicating that Cr(III) remains soluble for a pH increase of about 0.5 pH unit. iii. At acetate concentrations 20mM, (100 fold the initial concentration), a significant shift of the solubility curve is observed for 2 pH units i.e. from pH 6 to 8. The results presented indicate the effect of the progressive increase of acetate concentration on chromium solubility. It seems that there is a threshold of acetate concentration between 2 and 20mM above which a significant increase in Cr(III) solubility is observed towards more alkaline pH values.
11 10 9 8
Cr(III) mg/L
7 6 5 4 3 Acetate/Cr: 1/1 mole ratio Acetate/Cr: 10/1 mole ratio Acetate/Cr: 100/1 mole ratin
2 1 0 0
1
2
3
4
5
6
7
8
9
10
pH
Figure 2. Experimental determination of Cr(III) solubility as a function of pH in the presence of acetate. Three different acetate to chromium mole ratios have been experimentally examined
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Figure 3 presents the experimental determination of Cr(III) solubility as a function of pH in the presence of alanine. The solubility curves have been determined for three different alanine to chromium mole ratios, 1/1, 10/1 and 100/1. From this figure it is apparent that: i. For equimolar chromium and alanine solution concentrations, the solubility curve is shifted about 1 pH unit to more alkaline values compared to that presented in Figure 1. This means that at alanine concentration of 0.2 mM Cr(III) is quantitatively precipitated above pH 7.5. ii. When alanine concentration increases 10 fold, the solubility curve becomes wider. This means that more Chromium remains soluble in a wider range of pH values. In addition, the minimum of the curve is positioned at around 0.5- 1ppm of soluble Cr(III), indicating that a percentage of 5-10% of total Cr(III) remains soluble in the presence of alanine even at pH 8 – 9.5. iii. At alanine concentrations 20mM, (100 fold the initial concentration), the solubility of chromium becomes even higher and about 50% of total Cr(III) remains soluble. The results presented in Figure 3 show a more pronounced effect of alanine in chromium solubility than that observed for chromium solubility in the presence of acetate (Figure 2). There is also a threshold of alanine concentration between 2 and 20mM above which a significant increase in Cr(III) solubility is observed. 11 10 9 8
Cr(III) mg/L
7 6 5 4 3 Cr/Alanine: 1/1 mole ratio Cr/Alanine: 1/10 mole ratio Cr/Alanine: 1/100 mole ratio
2 1 0 0
1
2
3
4
5
6
7
8
9
10
pH
Figure 3. Experimental determination of Cr(III) solubility as a function of pH in the presence of alanine. Three different chromium to alanine mole ratios have been experimentally examined
A similar behavior to that of alanine-chromium solutions is observed for caseinchromium solutions as is shown in Figure 4. In this case weight ratios have been used. Figure 5 shows that in presence of nutrient broth, the increase in chromium solubility is very pronounced, reaching a minimum value of 80% soluble chromium for nutrient broth concentration of 1000 mg/l. This behavior could be attributed to a mixture of organic molecules present in the nutrient broth having strong complexing capability. 715
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11 10 9 8
Cr(III) mg/L
7 6 5 4 3 2
Cr/Casein 1/1 weight ratio Cr/Casein 1/10 weight ratio Cr/Casein 1/100 weight ratio
1 0 0
1
2
3
4
5
6
7
8
9
10
pH
Figure 4. Experimental determination of Cr(III) solubility as a function of pH in the presence of casein. Three different chromium to casein weight ratios have been experimentally examined 11 10 9 8
Cr(III) mg/L
7 6 5 4 3 2
Cr/Nutrient Broth 1/1 weight ratio Cr/Nutrient Broth 1/10 weight ratio Cr/Nutrient Broth 1/100 weight ratio
1 0 0
1
2
3
4
5
6
7
8
9
10
pH
Figure 5. Experimental determination of Cr(III) solubility as a function of pH in the presence of nutrient broth. Three different chromium to nutrient broth weight ratios have been experimentally examined
The results presented above have shown that representative organic molecules of biological significance such as organic acids, amino acids, proteins and mixtures of them, significantly enhance the solubility of trivalent chromium at pH values where, in the 716
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absence of organic molecules, chromium would form insoluble species. For the compounds experimentally tested, as well as for nutrient broth, a common microbiological growth medium, when the organic is present at an excess of a hundred fold molar or weight concentration (i.e. 20mM or 1000mg/l), the solubility of chromium is significantly enhanced, hindering the formation and precipitation of insoluble chromium species. The observed progressive increase of chromium solubility in this work when the concentration of organic molecules increases can be explained by considering the chemical coordination and stereochemical characteristics of trivalent chromium. Trivalent chromium forms complexes with virtually any species capable of donating an electron pair. These complexes may be anionic, cationic, or neutral and, with hardly any exceptions, are hexacoordinate and octahedral, [10]. The various complexes of Cr(III) are notable for their kinetic inertness, due to the octahedral d3 electron configuration. Ligand substitution and rearrangement reactions are slow (half times are of the order of hours), [10]. Moreover, the preference of Cr(III) to form hexacoordinate and octahedral complexes explains the need of excess of the ligand confirming the results obtained in this work. Evident consequences of the above to the biological treatment of wastewater containing hexavalent chromium are expected. Biological reduction of hexavalent chromium in biological reactors has been successfully achieved and is documented in the literature, [1,2]. In the reactor environment, organic molecules such as those presented above are very common and usually present in much higher concentrations than those examined in this work and in various combinations. For example, nutrient broth is prepared by dissolving 15g/l of the commercial reagent in forming common microbiological nutrient medium. Moreover, the content of spent medium in organic molecules such as organic acids, amino acids, proteins etc. is much higher than the highest concentration of organic molecules tested, [8]. When Cr(III) concentration in solution is comparable or is in excess of organic compounds, according to our results, no modification of its solubility behavior is expected. When organic compounds concentration is in excess, the formation of Cr(III) complexes is observed, and precipitation is hindered. It is apparent that the concentrations of 20mM of a pure compound and that of 1000mg/l of a mixture of organic molecules, represent a low limit of organic content, below which chromium precipitation could be initiated at concentration levels of 10 mg/l Cr(III). Although biological reduction of Cr(VI) to Cr(III) is feasible, the precipitation of Cr(III) depends on the relative abundance between chromium and the organic molecules present in the bioreactor environment. It can be concluded that effluent standards of 0.5-2 ppm Cr(III) for treated waste water, which is requested from the European legislation, is difficult to be achieved by standard alkaline precipitation treatment when organic molecules are present in excess. 4.
1.
CONCLUSIONS The following conclusions can be drawn based on the results of this study: The solubility curve of pure trivalent chromium solutions, determined experimentally in this work, is in good agreement to that reported in the literature and is used as a reference curve.
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2.
3.
4.
The solubility of Cr(III) in the presence of characteristic organic molecules, likely present in biological reactors, has been quantitatively determined by drawing experimental solubility curves for three organic molecules and a commercial microbial nutrient (mixture of organic compounds). For each experimental set, three solubility curves were obtained corresponding to chromium to organic molecule concentration ranging from 1/1, 1/10 and 1/100 in mole or weight ratios. For the compounds tested as well as for the microbial nutrient, a progressive increase of trivalent chromium solubility was observed systematically when the concentration of organic molecules in solution was increased, and is in excess compared to chromium concentration. This result matches well with the chemical coordination and stereochemical characteristics of trivalent chromium. In practice in most biological reactors organic content is higher than this examined in the present work. This means that if this organic content is also in excess of Cr(III), the precipitation of Cr(III) is hindered due to Cr(III)-organic complex formation.
REFERENCES
1.
A. Hatzikioseyian, E. Remoudaki, M. Tsezos, Proceedings of International Biohydrometallurgy Symposium 2001, Part B, Editors V.S.T. Ciminelli, O.Garcia Jr., Elsevier, (2001), 265-277. 2. Fujie K., Tsuchiko T., Urano K., and Ohtake H., Wat. Sci. Tech., vol 30, No 3, pp. 235-243, 1994. 3. Fujie K., Hu H.Y., Huang X., Tanaka Y., Urano K., and Ohtake H., Wat. Sci. Tech., vol 33, no. 5-6, pp 173-182, 1996. 4. Yamamoto K., Kato J., Yano T., Ohtake H., Biotechnology and Bioengineering, vol 41, pp. 129-133, 1993. 5. P.C. Wang, T. Mori, K. Toda and H. Ohtake, Journal of bacteriology, 172, 16701672, 1990 6. K. Komori, P.Wang, K. Toda and H. Ohtake, Applied microbiology and biotechnology, vol 33, 117-119, 1990. 7. A.D. Covington, Chromium Review, No 5, pp. 2-9, 1985. 8. Y. Zao-Yan and F. Zhen-San, Wat. Sci. Tech. Vol 22, No. 1/2, pp. 119-126, 1990. 9. C.F. Baes & R.E. Mesmer, The Hydrolysis of Cations. Wiley, New York, 1976. 10. Greenwood N.N., Earnshaw A., Chemistry of the elements, Pergamon Press First Edition 1984.
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15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
The selective precipitation of heavy metals by sulphatereducing bacteria A. Luptakovaa*, M. Kusnierovaa, M. Bezovskaa and P. Feckob a
Department of Mineral biotechnology, Institute of Geotechnics of Slovak Academy of Sciences, Watsonova 45, 043 53 Košice, Slovak Republic b Institute Environmental Engineering, VSB-Technical University Ostrava, Mining and Geological Faculty, 17. listopadu 15, 708 33 Ostrava-Poruba, Czech Republic Abstract The aim of this work is to study the possibility of using sulphate-reducing bacteria for a purging process of the Acid Mine Drainage, which is considered to be the major environmental problem associated with mining activities. Tests were conducted to determine if biogenic hydrogen sulphide could be used to eliminate soluble heavy metals as insoluble sulphides. This method involves three stages such as: hydrogen sulphide production by sulphate-reducing bacteria, metals precipitation by biologically produced hydrogen sulphide and metal sulphides filtration. The second stage allowed the selective recovery of heavy metals. 1.
INTRODUCTION Heavy metals pollution is a major environmental problem caused by the industry and has important consequences for human and animal health. Industrial wastewaters are their main source. These waters are produced by metal processing, surface treatment companies as well as the mining industry. One of the most important problems affecting mining companies around the world is the treatment of acid mine drainage (AMD). AMD is the result of the natural oxidation of sulphide minerals contained in rocks when exposed to the combined action of sulphide and atmospheric oxygen. It is particularly related to coal, lignite and polymetallic sulphide mining. Sources of AMD are underground and open pit mining works, over burden and waste rock dumps, flotation tailings and concentrate stockpiles. Most of these sources remain active for decades or centuries after mine closure [1]. Oxidation of the sulphide minerals takes place through a complex series of reactions involving direct, indirect and microbially assisted mechanisms (genera of Acidithiobacillus and Leptospirillum). AMD contains sulphuric acid, dissolved heavy metals (Cu, Cd, Zn, Pb, Ni, Co, etc.), sulphates, solid iron precipitates and their pH is very low about 1.5 – 2.0. This water presents a harsh and extreme environment for biological activity and must be contained and treated before it can be discharged since it can have severe impact on the environment. *
Corresponding author, E-mail:
[email protected] The authors are grateful to the Slovak Grant Agency for Science (Grant No. 2210622) for the financial support of this work.
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Several methods for treatment of AMD exist but only few of them have been applied under commercial-scale conditions. The most common methods are chemical methods e.g. the neutralization using lime or other alkaline components. This results in the precipitation of sulphates and heavy metals as gypsum and metal hydroxides respectively, which have to be discharged. The operating costs of this process are high whereas sulphate and metals removal efficiencies are relatively low [2]. In addition, all valuable metals are lost in the sludge. A possible alternative to the chemical treatment of these effluents is bioremediation using anaerobic sulphate-reducing bacteria (SRB). The treatment of AMD by SRB is based on the ability of SRB to reduce sulphates to hydrogen sulphide, which binds readily with most metals to form an insoluble precipitate. The metals are therefore removed from solution in a stable form. The metabolism of SRB also generates alkalinity, which contributes towards neutralizing the acidity of the AMD [3]. The following reactions (1), (2) and (3) represent the transformations of the principal constituents of AMD by SRB: SRB
4H2 + SO42- + H+ ⎯⎯⎯→ HS- + 4H2O 2-
SRB
Organic matter + SO4 ⎯⎯⎯→HS- + HCO32+
-
(1)
+
(2)
2+
(3) Me + HS ⎯⎯→MeS (↓) + H (Me - the cation of particular metal) The SRB represent a group of chemoorganotrophic and strictly anaerobic bacteria that may be divided into four groups based on rRNA sequence analysis: gram-negative mesophilic SRB, gram-positive spore forming SRB, thermophilic bacterial SRB and thermophilic archaeal SRB [4]. These groups of bacteria include representatives of the Desulfovibrio, Desulfomicrobium, Desulfobacter, Desulfosarcina, genera Desulfotomacullum, Thermodesulfobacterium, Archaeoglobus, etc. The closing and flooding of mines in Slovakia (e.g. Smolník, Pezinok, Sobov) generates large volumes of AMD. The old metal processing sites lead to major problems of this metal contaminated groundwater. Therefore we studied the possibility of using SRB for a bioremediation process for AMD. In this article, we present our initial results, which led us to construct a laboratory reactor of a simple design to evaluate the ability of SRB for treating AMD. This method involves three stages such as: the production of hydrogen sulphide by sulphate-reducing bacteria, the precipitation of metals by the biologically produced hydrogen sulphide and the filtration of metal sulphides. The experiments were conducted with synthetic solutions containing copper and cadmium. The second stage allowed the selective recovery of these heavy metals in the form of CuS and CdS. 2.
MATERIALS AND METHODS
2.1 Microorganisms A culture of SRB (genera Desulfovibrio and Desulfotomaculum) was obtained from the wastewater collection tank used for washing machinery in the metallurgical plant. For the isolation and cultivation of aforementioned bacteria a selective nutrient medium (DSM-63 - Postgate′s C medium) was used [5]. The growth of SRB was detected by the formation of black precipitates at the bottom of the flasks and flask walls. Bergey's Manual of Determinative Bacteriology was used for identification of these bacteria. 720
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2.2 Model solutions The experiments were conducted with synthetic solutions containing Cu2+ and Cd2+. These solutions were prepared by dissolving the analytical grade sulphate salts of copper (CuSO4.5H2O) or cadmium (CdSO4.5H2O) in distilled water. These solutions contained: 1 – Cu 10 mg/l, 2 - Cd 10 mg/l, 3 - Cu 10 mg/l + Cd 10 mg/l. The pH of the solutions was adjusted to the desired value using sulphuric acid and sodium hydroxide at need. 2.3 Analytical procedures A turbidimetric method was used to measure the concentration of sulphate ion concentrations [6]. Sulphates form an insoluble precipitate with barium (BaCl2) under acidic conditions. The absorbance of the sample was measured at a wavelength of 420 nm using a Spectromom195 (Hungary) instrument. Samples were centrifuged for 10 minutes at 10000 rpm before performing the analysis. The concentration of metals in the samples taken in the model solution was determined by atomic absorption spectrometry (AAS) using Spectrometer AA – 30 Varian (Australia) instrument. Samples for metal analysis were acidified after collection and centrifuged for 10 minutes at 10000 rpm. A glass pH electrode combined with the reference Ag/AgCl electrode and a platinum redox plus Ag/AgCl reference electrode were used to measure pH and redox potential (Eh) respectively. Digital pH- meter GPRT 144 AGL (Germany) and mV-meter Ion-activitymeter MS 20 (Czech Republic) were used. The qualitative analysis of precipitates obtained by biologically produced hydrogen sulphide was done by energy dispersive spectrometry (EDS) analysis using instruments, which consisted of a scanning electron microscope BS 300 (Tesla, Czechoslovakia) and an X-ray microanalyser EDAX 9100/60 (Philips, Holland). Samples of precipitates were dried and coated with gold before the EDS analysis. 2.4 Elimination of copper and cadmium from model solutions Figure 1 shows a schematic diagram of the experimental sequence used in the study of elimination of heavy metals (Cu and Cd) from model solutions.
Figure 1. The schematic diagram of the experimental sequence
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The experiments were performed in two bioreactors with a capacity 1000 ml (the first bioreactor) and 250 ml (the second bioreactor). This method contains several process steps and can be divided in to three stages: 1. Biological hydrogen sulphide production – in this stage SRB using organic matter (lactate) and hydrogen as electron donor, convert the sulphate to hydrogen sulphide (according to reactions (1) and (2)). This part was carried out under following conditions: • a discontinuous hermetically closed reactor (a first reactor) • statically • temperature 30°C • selective nutrient medium for D. desulfuricans (DSM-63 – Postgate's medium), pH 7.5 • anaerobic conditions ( nitrogen gas) 2. Metals precipitation by the biologically produce hydrogen sulphide - the fundamental is the recovery of heavy metals from solution in the form of sulphides due to by reaction between the biologically produced hydrogen sulphide and heavy metals (according to reaction (3)). This stage followed after the indication of sulphate reduction (blackening of the medium by production of FeS in the first bioreactor). It was carried out under following conditions: • continuous stripping of a gas mixture (H2, CO2 and H2S) by N2 from the first bioreactor into a second reactor, which was filled the model solution of heavy metals • continuous precipitation of Cu at pH 2.8 (next Cd at pH 3.5) by the biologically produced H2S 3. Heavy metal sulphides separation - in this stage precipitates of the heavy metals sulphides (CuS and CdS) were removed from the model solution by filtration. The precipitates were dried and analysed using the qualitative EDS analysis. 3.
RESULTS AND DISCUSSION
3.1 Biological hydrogen sulphide production Figures 2 and Figure 3 show typical changes of sulphate concentration, pH and Eh during the discontinuous cultivation of SRB in the first reactor i.e. during the biological hydrogen sulphide production. The sulphate concentration decreased slowly at the beginning of the process and rapid decreasing was observed after 20-25 hours. The pH increased gradually from 7.6 to 8.9. The increase of pH after 20 hours may be attributed to the reduction of sulphate by SRB according to reaction (4) : (4) 2 H+ + SO42- + 2Corg ⎯⎯⎯→ H2S + 2CO2 The oxidation-reduction potentials (Eh) decreased from –140mV to –260mV and remained at the negative values until the end. The decrease of sulphate concentration, the increase of pH values, the decrease of Eh values, the formation of black precipitates (generation of FeS according to reaction (3)) and the sensorial detection of classical strong H2S smell were not observed in the abiotic control. It confirms that aforesaid changes were caused by the bacterial metabolism of SRB.
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Figure 2. Changes of sulphates concentration and pH values in the first bioreactor during the discontinuous cultivation of sulphate-reducing bacteria. Sulphates I. and pH I. – abiotic control; sulphates II. and pH II. – with sulphate-reducing bacteria
Figure 3. Changes of pH and Eh values during the discontinuous cultivation of sulphate-reducing bacteria 3.2 Metals precipitation by the biologically produce hydrogen sulphide The results of metals precipitation (the second stage of this method) are summarized in Figures 4-7. They demonstrate that Cu or Cd was effectively recovered from the solution using biologically produced H2S. Figure 4 shows the situation when precipitation of Cu from the solution was studied. In this case the starting concentration of Cu was 10 mg/l and the range of suitable pH values for precipitation of Cu was 2.5-2.8. Cu was removed from the model solution completely in the course of 4 hours. Figure 5 demonstrates the study of Cd2+ precipitation from the model solution with the initial concentration of Cd 10 mg/l. The range of suitable pH values for precipitation of Cd was 2.5 – 3.5. Cd was removal from the model solution completely in the course of 2 hours.
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Figure 4. Precipitation of Cu2+ by biologically produced H2S from the solution at different pH values. Starting concentration of Cu - 10 mg/l.
Figure 5. Precipitation of Cd2+ by biologically produced H2S from the solution at of different pH values. Starting concentration of Cd - 10 mg/l
The possibility of selective recovery of heavy metals was investigated, too. The selective recovery of metals is based on the different solubility of metal sulphides at different pH values [1]. By adjusting the pH in the precipitation reactor (the second reactor) to a certain value, it is possible to form a specific metal sulphide precipitate. This principle was used to remove Cu and Cd from the model solution when the starting concentration of Cu was 10 mg/l and of Cd 10 mg/l. Figure 6 demonstrates that copper was effectively and selectively recovered in the presence of Cd from the model solution at pH 2.8 using the biologically produced H2S provided from the first reactor. The total elimination of Cd was achieved during 3 hours. The concentration of Cd remained without changes. After that the suspension of precipitates was filtered and the pH value of the filtrate was adjusted at 3.5 using sulphuric acid. The filtrate was returned into the second reactor and again was submitted to the effect of biologically produced H2S (the subsequent 724
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precipitation). A decrease of the concentration of Cd was observed. The total elimination of Cd was achieved during 45 minutes. A slight precipitation of Cu was registered, too (Figure 7).
Figure 6. Decrease of Cu concentration during selective precipitation of Cu from the model solution (10mgCu/l + 10mg Cd/l) at pH 2.8 by biologically produced H2S
Figure 7. Decrease of Cd concentration during selective precipitation of Cd at pH 3.5 by biologically produced H2S from the filtrate after selective precipitation of Cu
3.3 Heavy metal sulphides separation Figures 8 and 9 indicate that probably Cu and Cd were precipitated in the form sulphides CuS and CdS. The element composition of originated precipitates correspond with this fact.
Figure 8. EDS qualitative analysis of precipitates originating from selective precipitation of Cu from model solution (10mgCu/l + 10mgCd/l) at pH 2.8 by biologically produced H2S 725
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Figure 9. EDS qualitative analysis of precipitates originating from selective precipitation of Cd at pH 3.5 by biologically produced H2S from the filtrate after selective precipitation of Cu 4.
CONCLUSIONS Experimental studies confirm that the sulphate-reducing bacteria can be used for the removal of heavy metals from model solution into three stages such as: hydrogen sulphide production by sulphate-reducing bacteria, metals precipitation by biologically produced hydrogen sulphide and metal sulphides filtration. The results of the first stage demonstrate that the isolated sulphate-reducing bacteria were active, because the decrease of the sulphate concentration, the increase of pH values, the decreased of Eh values, the black precipitates creation and the sensorial detection of classical strong H2S smell were not observed in the abiotic control. The finding of the second stage documented that:
− Cu and Cd are effectively recovered from individual model solutions using biologically produced hydrogen sulphide at pH 2.5-2.8 (for Cu) and pH 2.5-3.5 (for Cd) − this stage allowed the selective recovery of these heavy metals − Cu and Cd was selective recovered at pH 2.8 and pH 3.5, respectively The results of the third stage confirm that heavy metals precipitated in the sulphides form. In further experiments this method will be studied for the treatment of real solutions – Acid Mine Drainage from the closed mines in Slovakia (e.g. Smolnik, Pezinok, Sobov). REFERENCES
1.
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A. Kontopoulos, In: S.H. Castro, F. Vergara and M.A. Sánchez (eds.), Effluent Treatment in the Mining Industry, University of Conception, Chile, 1988, 57-112.
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2. 3. 4. 5. 6. 7.
J. Boonstra, R. van Lier, G. Janssen, H. Dijkman and C.J.N. Buisman, In: R. Amils, A. Ballester (eds.), Biohydrometallurgy and the Environment toward the mining of the 21st century, Elsevier, Amsterdam, 1999, 559-566. D. Lyew and J. Sheppard, Wat. Res. Vol. 35, No. 8 (2001) 2081-2086. H.F. Castro, N.H. Williams and A. Ogram, FEMS Microbiology, 31 (2000) 1-9. G.I. Karavaiko, G. Rossi, A.D. Agate, S.N. Groudev, Z.A. Avakyan (eds.), Biotechnology of metals, Centre of projects GKNT, Moscow, 1988, 59-61. APHA, Standard Methods for the Examination of Water and Wastewater, 17th edition, American Public Health Association, USA, Washington D. C., 1989. J.M. Odom and J.R. Rivers Singleton (eds.), The sulphate-reducing bacteria Contemporary Perspectives, Springer-Verlag, New York, 1993.
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A PPENDIX
15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Author index
A Abin L., 1127 Acevedo F., 185, 219, 227 Acosta M., 1287 Adamov E.V., 85 Adryanczyk-Perrier G., 1147 Agate A.D., 101 Agatzini-Leonardou S., 41 Ageeva S.N., 1379 Ahmad W.A., 513, 859 Ahring B.K., 567 Alvarez-Rosales E., 419 Amann R., 1325 Amils R., 653, 1325 Andrade M.C., 835
Beskoski V.P., 345 Bevilaqua D., 1023 Bezovska M., 703, 719 Bipinraj N.K., 491 Blazquez M.L., 783 Bond P., 1337 Bosecker K., 585, 1127 Bowen A., 1337 Boyer A., 925 Brandl H., 457 Braun J.J., 1037 Breed A.W., 1215 Brown N.L., 1119, 1137 Bruhn D.F., 1195 Bunyok M.A., 513
B Baillet F., 925 Baker-Austin C., 1337 Ballester A., 783 Banderas A., 1361 Bandhapadhyay T., 843 Banerjee P.C., 291, 1297 Banghui G., 35 Barreto M., 1271 Bastiaens L., 671 Bath M., 387 Battaglia-Brunet F., 549, 639, 1313 Bauer-Cuya J., 947 Baxter-Plant V.S., 1147 Benedetti A.V., 1023 Beolchini F., 731
C Canales C., 185 Cara J., 653 Carballo M.T., 653 Carcamo C., 277 Carvalho N., 75 Casamayor E.O., 1325 Castillo A., 243 Chandraprabha M.N., 1099 Chassary P., 935 Chatterjee B.P., 869 Chatterjee S., 843, 869 Chen J.P., 807, 817 Chevard S., 549 Chiacchiarini P., 117 Chida T., 663 A-3
Author index
Cho K.S., 613, 631 Chockalingam E., 1037 Choi M. S., 631 Ciminelli V.S.T., 965 Clarens M., 1313 Clark T.R., 497 Cohen R.H., 497 Cossich E.S, 919 Costa M.C., 75 Coto O., 1127 Coupland K., 639, 1057 Cruz F., 1085 Curutchet G., 595, 681 D d’Hugues P., 147, 1313 d’Souza S.F., 975 da Costa W.C., 793 Dahrazma B., 253 Dave S.R., 211, 1279 Davodi M., 261 de Azevedo Gomes H., 793 de Carvalho R.J., 301 de Carvalho R.P., 741, 835 de J. Cerino-Cordova F., 1109 de Kock S.H., 319 de Sousa A-M.G., 741 de Vargas Parody I., 935 Deane S.M., 1249 Dementin S., 1147 Dempers C.J.N., 1215 Denamur A., 549 Deseure J., 337 Dianzuo W., 399 Dictor M.C., 639 Diels L., 671 Dixon D.G., 65 Donati E., 117, 595, 681 Dopson M., 1337 A-4
Doycheva A.S., 697 du Plessis C.A., 319 Duarte G., 965 Dymov I., 377 E Eidan L.S., 919 El Korchi-Buchert D., 997 Endo M., 689 Englert G.E., 1085 Escobar B., 1091 Essa A.M.M., 1119 F Farah C., 1361 Faramarzi M.A., 457 Fecko P., 703, 719 Fehmida F., 771 Ferrini M., 957 Ferron C.J., 377 Filho O.B., 301 Fischer H., 575 Fomina M.A., 749 Forster C.F., 525 Fouad M.Q., 771 Foucher S., 549 Fraguela P., 783 Freitas J.R., 741 Frenay J., 285 Frizan V., 117 Fugivara C.S., 1023 Fujii M., 1305 G Galan-Wong L.J., 419 Galvez-Cloutier R., 175 Garcia Frutos F.J., 235, 653 Garcia Jr. O., 559, 793, 911, 1023 Gardner M.N., 1249 Garrido F., 639
Author index
Geets J., 671 Gentina J.C., 185, 219, 227 Ghosh S., 1297 Giaveno A., 117 Glendinning K.J., 1137 Glombitza F., 575 Golan M., 1237 Goldschmidt G.K., 1249 Gomez F., 653 Gondrexon N., 337, 1109 Gonzalez F., 783 Gonzalez-Toril E., 1325 Groudev S.N., 533, 697 Groudeva V.I., 697 Guha A.K., 843, 869 Guibal E., 621, 899, 935, 947 Guiliani N., 1287, 1361 Gutierrez D., 1127 H Hallberg K.B., 639, 1057, 1163, 1179 Hansford G.S., 1215, 1227 Harneit K., 1369 Harrison S.T.L., 359 Harvey T.J., 387 Hatzikoseyan A., 711 Hevia M.J., 1091 Holmes D.S., 989, 1187, 1271 Hong J. H., 631 Hongxu L., 399 Horak G., 799 Humphries A.C., 525 Hurtado J.E., 1353 I Iglesias N., 75 Ignatiadis I., 549 Imanakunov B., 465 Inoue C., 663
J Jaapar J., 513 Janneck E., 575 Jedlicki E., 989, 1187, 1271 Jensen P.E., 567 Jerez C.A., 1287, 1361 Johnson D.B., 165, 639, 1057,1163, 1179, 1195 Joshi N.R., 491 K Kalman E., 1003 Kaltsa F., 711 Kamimura K., 449, 1261, 1305 Karamanev D., 647, 1067 Karapantsios T.D., 849 Karavaiko G.I., 85, 1389 Karnatz F., 575 Kasatkina T.P., 749 Kazy S.K., 975 Khijniak T., 483 Kim D.J., 631 Kim D.S., 631 Kimura S., 1057, 1179 Kinnunen P.H.-M., 193 Kinzler K., 1003 Klauber C., 1011 Kolousek D., 157 Komnitsas K., 533 Kondrateva T.F., 1379 Konno H., 689 Krambrock K., 741 Krylova N., 85 Kuncoro E.K., 621 Kusnierova M., 703, 719 Kwon H.K., 541 L Lamaignere V., 359 Lazic M.L., 1077 A-5
Author index
Lehtonen T., 621 Lengauer C., 157 Lilova K., 647, 1067 Llobet-Brossa E., 1325 Loi G., 429 Lopez J., 243 Lopez-Juarez A., 203 Loukidou M.X., 849 Lozovaya O.G., 749 Lugg H., 1205 Luptakova A., 703, 719 Ly Arrascue M., 947
Monhemius A.J., 409 Moon H.S., 613 Moran A., 653 Moreno L., 311 Morin D., 147, 549, 1313 Morra C., 337 Mugabi M., 147 Mukherjee A., 25 Müller I.L., 1085 Mulligan C.N., 175, 253 Munoz J.A., 783 Muntyan L.N., 1379
M Macaskie L.E., 525, 935, 1119, 1147, 1155, 1205 Maeda T., 449, 1305 Magnin J.P., 337, 925, 1109 Mahapatra N.R., 1297 Maliszewska I., 439 Mandal S.K., 291 Marquis P.M., 1205 Marrero J., 1127 Martinez L., 595 Massacci P., 957 Matic V., 345 Matis K.A., 849 Matlakowska R., 265, 1237 Medrano-Roldan H., 419 Medvedeva-Lyalikova N.N., 483 Mengel-Jung G., 585 Meruane G., 277 Mesquita H.M., 919 Metodiev B., 1067 Meyer G., 1369 Migliavacca E., 957 Mikheenko I.P., 1147 Miller H., 1163 Mitchell D., 1369 Modak J.M., 25, 1099
N Nakbanpote W., 891 Naldrett K., 319 Nancucheo I., 219 Nandi S., 1297 Natarajan K.A., 25, 759, 1037, 1099 Ndlovu S., 409 Nebera V.P., 107 Negishi A., 449, 1305 Nemati M., 359 Norris P.R., 1347 Nuzhat A., 771
A-6
O Ohimain E.I., 475 Okibe N., 165 Oliazadeh M., 261, 423, 507 Ottosen L.M., 567 Ouattara A., 175 Ozil P., 337, 925, 1109 P Pagnanelli F., 731, 825 Paknikar K.M., 491, 877 Palencia I., 75 Palmieri M., 911 Panin V.V., 85 Papassiopi N., 603
Author index
Park D., 883 Park J.M., 541, 883 Park K.H., 631 Parker A., 1011 Paspaliaris I., 533, 603 Paterson-Beedle M., 1155 Patino E., 285 Paulo P.S., 243 Pearce S.J.A., 359 Pehkonen S.O., 193 Peirano Blondet F., 947 Penfold D.W., 525 Pereyra-Alferez B., 419 Pertsov N., 127, 465 Petersen J., 65 Phillips W., 377 Pimentel P.F., 835 Pinka J., 575 Pivato D., 243 Pivovarova T.A., 85, 1379 Podgorsky V.S., 749 Podolska V., 127, 465 Polowczyk I., 439 Pompe W., 799 Puhakka J.A., 193 Q Quatrini R., 989 R Raichur A.M., 25 Raja S., 359 Ramirez P., 1287 Ramirez-Rodriguez D.G., 419 Ranville J.F., 497 Rawlings D.E., 1249 Remoudaki E., 711 Renman R., 399 Rivas M., 1271 Rivera-Santillán R.E., 203
Roberto F.F., 1195 Rohi R., 261 Rohwerder T., 997, 1171 Rojas-Chapana J., 1047 Romera E., 783 Rossi G., 3, 429 Rousset M., 1147 Roussy J., 621, 947 Rubinger C.P.L., 741 Rubio A., 235 Ruiz M., 935 Ryu H.W., 613, 631 S Sadowski Z., 439 Saha T., 843 Salazar M.F.M., 419 Salo V.L.A., 193 Sammons R.L., 1205 Sampaio C., 1085 Sand W., 997, 1003, 1171, 1369 Sandoval R., 285 Santos M.H., 835 Sar P., 975 Sasaki K., 689 Sastre A., 935 Savari E.Ε., 91 Savic D.S., 1077 Sayed Baghery S.A., 423, 507 Schippers A., 55, 585 Searby G.E., 1227 Sedelnikova G.V., 91 Sen A.M., 1179 Shahverdi A.R., 261, 423 Sheng P.X., 817 Shishikado T., 1261 Sidborn M., 311 Silva E.A., 919 Simmons S., 1347 Simonoff M., 483 A-7
Author index
Sissing A., 359 Sklodowska A., 265, 1237 Smith K.S., 497 Sohn J.S., 631 Sohor S.H., 513 Solis-Soto A., 419 Solozhenkin P.M., 107 Som Majumdar S., 843, 869 Songkroah C., 891 Spanelova M., 899 Spasic S., 345 Spasova I.I., 533 Stackebrandt E., 1369 Stagars M., 457 Stott M., 1011 Styriak I., 157, 1029 Styriakova I., 157, 1029 Subramanian S., 759, 1037 Sugio T., 449, 1261, 1305 Sung J.Y., 541 Suto K., 663 Svecova L., 899 T Takahashi Y., 663 Takano K., 689 Takeuchi F., 449, 1305 Tan L.H., 807 Tavares C.R.G., 919 Teixeira M.C., 965 Teledgi J., 1003 Ten W.K., 137 Thiravetyan P., 891 Thyagarajan H., 759 Tillmanns E., 157 Ting Y.P., 137, 329, 807, 817 Tipre D.R., 211, 1279 Toro L., 731, 825 Tributsch H., 1047 Trois P., 429 A-8
Tsezos M., 711 Tupe S., 877 U Ubaldini S., 825 Ukermann S., 1163 Ulberg Z., 127, 465 Uzma B., 771 V Valdes J., 1187 Valencia P., 227 Valenzuela L., 1287 van Broekhoven K., 671 van Bronswijk W., 1011 van Zyl L.J., 1249 Vargas T., 277, 1091 Vaxevanidou K., 603 Veglio F., 731, 825 Veljkovic V.B., 1077 Veloso F., 989 Viera M., 681 Vigliocco A., 243 Villar L.D., 559 Vora S.B., 211 Voronin D.Y., 85 Vos J., 671 Vrvic M.M., 345 W Walton-Day K., 497 Watling H.R., 1011 Welzel A., 1085 Werner P., 799 Willscher S., 497, 799 Woo S.H., 541 X Xianwan Y., 35 Xu T.J., 329
Author index
Y Yakubenko L., 127, 465 Yamakado M., 1261 Yermolenko A., 127. Yong P., 1205 Yoo N.Y., 613 Yuehua H., 399 Yun Y.-S., 883 Yun Z., 35
Yuxia G., 35 Z Zafiratos J.G., 41 Zahari M.A.K.M., 513 Zakaria Z.A., 859 Zeballos F., 301 Zouboulis A.I., 849
A-9
15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas "Biohydrometallurgy: a sustainable technology in evolution"
Subject index 16S rDNA, 1313 16S rRNA, 1325, 1369 A Accessory genes, 1249 Acid Mine Drainage (AMD), 497, 585, 639, 997, 1057, 1163, 1179 Acid production potential, 1037 streamers, 1057 sulphate soils, 475 Acidification, 475 Acidihalobacter, 1347 Acidiphilium, 1171 Acidithiobacillus, 85, 117, 423, 1171, 1353 caldus, 1249 ferrooxidans, 337, 429, 449, 595, 631, 647, 925, 989, 1003, 1023, 1037, 1047, 1067, 1077, 1099, 1109, 1187, 1261, 1271, 1287, 1305, 1361, 1369 thiooxidans, 613, 681, 1369 Acidocella strain, 1297 Acidophile heterotrophs, 497, 585 Acidophiles, 1179, 1337, 1347 Acidophilic bacteria, 1195 Activated carbon cloth, 1109 Activation energy, 291 Adaptation, 1099, 1379 Adsorption, 439, 891 Aeromonas, 859 Agar-plate screening, 749 Air-lift, 211
Algae, 783 Algal biomass, 947 Analytical aspects, 345 Antimony, 107 Archaea, 219, 319, 1337 Arsenic, 603, 639, 799 precipitation, 377 Arsenite, 965 Arsenopyrite / Pyrite, 377 Arsenopyrite, 377, 1353 Aspergillus niger, 137, 291, 329, 899, 1127 Atomic Force Microscopy (AFM), 1003 Attached cells, 1091 B Bacillus polymyxa, 759 sp., 157, 1029, 1137 Bacteria modified, 399 Bacterial activity, 277 attachment, 409 leaching, 75, 311 Barium, 793 Batch processing, 1215 Bench scale, 449 Beneficiate, 101 Bimetallic systems, 783 Bioaccumulation, 771 Biodegradation, 491, 541 Biodiversity, 1057, 1313 Biofilm, 541, 1109, 1361 formation, 1271 A-11
Subject Index
Biohydrometallurgy, 91, 107 Bioleaching, 25, 41, 85, 117, 127, 137, 147, 157, 203, 219, 235, 277, 285, 301, 319, 329, 359, 399, 429, 507, 559, 613, 1067, 1091, 1127, 1227, 1313 Biolixiviation, 175 Biological leaching, 377 Biomass estimation, 1091 Biomineralization, 1205 Biooxidation, 65, 219, 387, 513, 891, 1099 Bioreactors, 541, 1163, 1215 Bioreduction, 759 Bioremediation, 759, 965, 1163, 1179 Biosorption, 567, 731, 741, 749, 759, 771, 783, 793, 799, 807, 817, 825, 835, 843, 849, 859, 877, 883, 911, 919, 947 Biosurfactant, 439 Biotechnology, 35 Brown coal, 35 C Cadmium, 783, 849 Caesium, 1155 Calcium phosphate, 1205 sulphate, 703 Carbon dioxide, 319 fiber, 689 Catalytic effect, 203 Cellulose, 741 Chalcocite, 203 Chalcopyrite, 193, 203, 235, 399, 419, 1011, 1023, 1037, 1091 Chalk, 439 Chelation, 621 Chemoautotrophs, 1215 Chemolithotrophic microorganisms, 1379 A-12
Chitin, 891 Chitosan, 621, 869, 935 Chromate reduction, 525 Chromium, 711, 759, 807, 859, 883, 919 biosorption, 859, 919 removal, 681 Chromobacterium, 457 Chromosomal DNA, 1379 integration, 1297 Citric acid, 1127 Clay tiles, 595 Coal biodesulphurization, 653 Cobalt, 1155 Complex sulfides concentrates, 429 Complexing organic matter, 497 Computational fluid dynamics, 1067 Concentrates, 107 Continuous culture, 1227 Continuous electrochemical regeneration, 337 Converter slag, 507 Copper, 75, 253, 507, 631, 647, 741, 783, 799, 869 bioleaching, 311, 419 fixation, 925 heap, 423 mine, 1279 sulphides, 301 Corrosion patterns, 409 Crystal orientation, 409 CSTR, 185 Cyanide, 491, 541 Cyanogenic microrganisms, 457 Cysteine biosynthesis, 1187 D Deferritization, 1029 Design Of Experiment (DOE), 925 Desorption, 835 Desulfotomaculum nigrificans, 1037
Subject Index
Desulfovibrio, 681 fructosovorans, 1147 vulgaris, 525 DGGE, 1369 Dibenzothiophene, 345 Diversity, 1279 Dolomite, 439 Dredging, 475 E E. coli, 1297 EDTA, 603 Electric field, 465 Electrochemistry, 399 Electron paramagnetic resonance (EPR), 741 Electronic scrap materials, 137 Elemental sulfur, 91 Enargite, 219 Environmental risk potential, 585 Equilibrium, 731, 799, 825, 975 EXAFS, 965 Exopolysaccharide, 1337 Extracellular polysaccharides, 1271 Extremely thermophilic culture, 235 F FeoB, 989 Ferric leaching, 193 Ferrooxidans, 301 Ferrous iron catalysis, 235 oxidation, 277 Fixed-bed columns, 919 Flotation tailings, 85 Fluorescence in situ hybridization, 653 Fluorescent in situ hybridization (FISH), 1057 Fluvial tailings deposit, 497 Fly ash, 329 Forced aeration, 311
Fossil fuels, 345 FTIR, 741, 835 Fungal biomass, 947 G Galactose, 1271 Galena, 127 Galvanic couple, 127 Ge, 35 Genome analysis, 989, 1187, 1271 GeoBiotics process, 261 Geochemistry, 175 GEOCOAT®, 387 Gold recovery, 377 Goldbearing ores, 107 sulfide concentrates, 91 Gram-negative bacteria, 1279 Growth inhibition, 65 kinetics, 65 H Haloalkaliphilic bacteria, 483 Halomonas, 483 HCN, 457 Heap leaching, 65 Heavy metals, 175, 613, 671, 731, 771, 799, 817, 825, 883, 957, 1119 Hematite, 439 Heterotrophic microorganisms, 1127 High-cell density, 337 Horizontal gene pool, 1249 Hydrogen generation, 525, 663 peroxide, 689 sulphide, 663 uranyl phosphate, 1155 Hydroxyapatite, 1205 I Immobilization, 595 A-13
Subject Index
Immobilized cells, 525 Indigenous thiobacilli, 559 Indirect action, 277 Inhibition, 1099 Iron leaching, 291 oxidation, 193, 653, 1227 oxide, 55 regulation, 989 solubilisation, 227 uptake, 989 Ironoxidising bacteria, 265 oxidizing acidophiles, 585 reducing bacteria, 603 ISAfe1, 1353 Isotherms, 935, 947 K Kaolin, 439 Kinetic parameters, 1215 Kinetics, 935, 947, 1227 Klebsiella pneumoniae, 1119 L Lactuca Sativa, 835 Lanthanum, 911 Laterite ore, 1127 Lead, 843, 899, 947 Leptospirillum, 165 ferrooxidans, 429, 1003, 1047, 1369 Lignin, 741 Low-grade ores, 41 M Magnetite, 127 Maintenance requirements, 1215 Manganese, 41, 1163 oxide, 55 Mangrove, 475 Marine A-14
algal, 807 isolate, 25 Mechanism, 399 Membrane proteins, 1261 Mercury resistance, 1137 Mercury, 449, 621, 835, 877, 899, 1085, 1305 Mesophilic, 203 bacteria, 1279 Metabolic activity, 359 Metal, 749 binding, 975 complexed cyanide, 465 cyanides, 457, 541 recovery, 1179 removal, 849 resistance, 1297, 1337 sorption, 697 sulfide oxidation, 55 Metal-organic complexes, 711 Metals, 559 Microbial cell damage, 359 degradation, 465 leaching, 243 monitoring, 653 recovery, 631 sulphate reduction, 697 Microcalorimetry, 997 Mine drainage, 533 Mineral suspension, 439 Mining tailings, 253, 513 waste, 585, 997 Mixed cultures, 165, 513, 1195 Mn-oxidizing fungus, 689 Modelling, 301, 311, 731, 825, 849, 919, 925, 1099, 1215 Moderate acidophiles, 639 Molecular ecology, 1325 Monitoring, 997
Subject Index
Mucor rouxii, 843 Mud, 1085 N Neodymium, 911 Nickel, 783 Niger delta, 475 Non-metallics, 1029 Numerical modelling, 311 O Oil, 697 Open pit mining, 997 Optimal conditions, 227 Optimization, 329 Organic acids, 175, 1127 carbon source, 689 Oxalic acid, 291 Oxidation, 399, 647 Oxide ore, 1127 Oxygen, 319 P Packed bed reactor, 653 Palladium, 935, 1147 Passivation, 193 Penicillium simplicissimum, 1127 Permeable reactive barrier, 697 Pertechnetate reduction, 483 pH, 419, 483 stress, 1195 effect, 731 Phosphatic ores, 101 Photographic waste, 891 Piloting, 377 Plant biomass, 741 Plasmids, 1249, 1297, 1379 Platinum, 935 Polymetallic concentrate, 211 Polysulfide, 663
mechanism, 55 Powder microelectrode, 399 Precipitation of heavy metals, 719, 1119 Pressure drop, 621 oxidation, 377 Printed circuit board, 631 Process control, 147 Process flowsheet, 85 Proteomics, 1287, 1337 Pseudomonas aeruginosa, 439 putida, 491 sp., 975 Pulp density, 85 Pyrite, 147, 165, 203, 261, 387, 409, 475, 1003, 1037,1047 biooxidation, 185 oxidation, 55 Q Quantitative ecology, 585 Quorum sensing, 1361 R Radium, 793 Reaction rate, 291 Reactor configuration, 185 Refractory gold concentrate, 185 gold, 377 ore, 387 Regeneration, 663, 957 Rehabilitation, 475 Remediation, 567, 671 Removal, 345, 1155 Response surface, 227 Restoration, 475 Retention efficiency, 621 Revolving drum bioreactor, 429 Rhamnolipids, 253 A-15
Subject Index
Rhodanese, 1287 S Salt tolerance, 1347 Sargassum sp., 793, 835, 919 Seaweeds, 817 Sediments, 253 SEM, 265 Semi-continuous, 211 Serratia sp., 1155, 1205 Settling, 439 Sewage sludge, 265, 559, 613 Shrinking-core, 301 Silicate bacteria, 157 minerals, 1029 Silver, 891 Soil remediation, 603 Solids loading, 359 Solubility curves, 711 Sorption, 835, 899, 957, 965, 975 sites, 741 Spectroscopy, 1011 Sphalerite, 85 Spirogyra insignis, 783 SSCP, 1313 Stability, 439 Starvation stress, 1195 Static mixers, 337 Stirred tank, 211 Strain polymorphism, 1379 Strontium, 1155 Sulfate Reducing Bacteria (SRB), 549, 663, 703, 719, 935, 1179 Sulfide, 387 concentrate, 147 ores, 117, 423 oxidation, 1171 Sulfidic dredge spoils / sediment, 475 Sulfolobus metallicus, 277, 319, 1091 A-16
Sulfur assimilation, 1187 cycle, 663 dioxygenase, 1171 metabolism, 1287 oxidation, 285, 1171 regulation, 1187 Sulfuroxidizing acidophiles, 585 oxidising bacteria, 265 oxidizing neutrophiles, 585 Surface treatment effluents, 549 Swarming, 1353 Synergism, 165 T Tailing, 75 Tc(III), Tc(IV), Tc(VII), 483 TDS, 65 TEM, 265 Temperature, 419 Textile, 859 Thallium, 1237 Thermophiles, 165, 219, 1227 Thermophilic, 319, 1137 archaeon, 227 Thiobacillus, 41 ferrooxidans, 101, 261, 345, 399 Thiomonas sp., 639 Thiosulfate mechanism, 55 Thorium, 975 TonB dependent receptors, 989 Total solids, 559 U Ultrafiltration, 621 Uranium, 243, 533, 575, 975 Ureibacillus sp., 1137 W Waste biomass, 899
Subject Index
Wastewater, 465, 697, 849 Wetlands, 533, 575 X X-Ray photoelectron spectroscopy, 1011 Y Yeast, 749
Yield, 65 Z Zea Mays, 835 Zeolite, 157 Zeta-potential, 127, 439 Zinc, 75, 117, 783
A-17
