Julkaisu 489 Publication 489

Anna Kaksonen The Performance, Kinetics and Microbiology of Sulfidogenic Fluidized-Bed Reactors Treating Acidic Metal- and Sulfate-Containing Wastewater

Tampere 2004

Tampereen teknillinen yliopisto. Julkaisu 489 Tampere University of Technology. Publication 489

Anna Kaksonen

The Performance, Kinetics and Microbiology of Sulfidogenic Fluidized-Bed Reactors Treating Acidic Metal- and Sulfate-Containing Wastewater

Thesis for the degree of Doctor of Technology to be presented with due permission for public examination and criticism in Festia Building, Small Auditorium 1, at Tampere University of Technology, on the 17th of September 2004, at 12 noon.

Tampereen teknillinen yliopisto - Tampere University of Technology Tampere 2004

ISBN 952-15-1234-2 (printed) ISBN 952-15-1404-3 (PDF) ISSN 1459-2045

Anna Kaksonen

The performance, kinetics and microbiology of sulfidogenic fluidized-bed reactors treating acidic metal- and sulfate-containing wastewater

ERRATA

Summary 2- - Page 50, equation 14: “HCO3 “should read “HCO3 “ Pages 56-59, Tables 18 and 19, 1st column 1st row: “rector” should read “reactor”

Paper I: Page 255, abstract, line 3, “250 and 350” should read 350 and 250” Page 257, Figure 2, part A, legend, “yeast extractt” should read “yeast extract” Page 258, right paragraph, 2nd last row of paragraph 3.1, “FBR3” should read “FBR2” Page 260, Figure 3, part D, “Fe load an precipitation” should read “Fe load and precipitation” Page 261, 1st paragraph, line 4: “79-87%” should read “79-86%” Page 261, 1st paragraph, line 6: “14-21%” should read “15-21%” Page 262, Table 4, caption, “FBR3” should read “FBR2” Page 263, Table 5, 1st row in column FBRs: “77-83” should read “77-85” Page 263, Table 6, last column: The order of the four first values should be: 78; 70; 55; 20

Paper II: Abstract, row 9: “2.7-3.5 mg l-1” should read “2.5-3.5 mg l-1” Page 209, Table 1, row 7: “Thioglycolic acid” should read “Sodium thioglycollate” Page 213, 1st column 6th last row: “2.7-3.5 mg l-1” should read “2.5-3.5 mg l-1” Page 214, Table 3, row 4, last column: “Widdel (1988” should read “Widdel (1988)” Page 214, Table 3, row 7, 3rd column: “5.9” should read “5.7”

Paper IV: Page 280, right column, 6th last row: “Thioglycolic acid” should read “Sodium thioglycollate” Reference [36]: The publication year should be 1995.

ABSTRACT

In this work, a sulfidogenic fluidized-bed reactor (FBR) process was developed for treating acidic metal- and sulfate-containing wastewater, the process design parameters were determined and the bacterial diversity of the FBR was described. The process is based on sulfate reduction by sulfate-reducing bacteria (SRB), precipitation of metals as sulfides with the biogenic H2S, and neutralization of the water with biologically produced bicarbonate alkalinity. Lactate and ethanol were used as electron donors for sulfate reduction.

The FBR process performance was evaluated under various operational conditions with continuous flow experiments. In terms of sulfate reduction, sulfide production and effluent alkalinity, the start-up of the FBR with 10 % fluidization rate was superior to the FBRs with 20-30 % fluidization rates. However, the FBRs with 20-30 % recycling rates performed better at the highest loading rates. The substrate feed of the FBR could be changed from lactate to ethanol without significant change in the FBR performance.

The robustness of the bioprocess was studied by increasing stepwise the zinc, sulfate and substrate feed concentrations and decreasing the feed pH, as well as decreasing hydraulic retention time (HRT) until process failures occurred. Lactate- and ethanol-utilizing FBRs precipitated 350-360 mg Zn l-1 d-1 and 86 mg Fe l-1 d-1 from acidic wastewater containing 230-240 mg Zn l-1 and 57-58 mg Fe l-1 at a HRT of 16 h. Percentual metal precipitation was over 99.8 % resulting in effluent soluble Zn and Fe concentrations below 0.1 mg l-1. The alkalinity produced in the sulfidogenic lactate and ethanol oxidation increased the wastewater pH from 2.5 to 7.5-8.5.

Maximum metal precipitation rates were obtained with ethanol-fed FBR at a HRT of 6.5 h from acidic wastewater containing 176 mg Zn l-1 and 87 mg Fe l-1. The rates were over 600 mg Zn l-1 d-1 and 300 mg Fe l-1 d-1, percentual metal precipitation over 99.9 % and effluent metal concentrations below 0.08 mg l-1. The alkalinity produced in ethanol oxidation increased the initial pH of 3, resulting in effluent pH of 7.9-8.0. At the HRT of 6.5 h, ethanol oxidation was incomplete and acetate accumulated in the FBR. The operation of the FBR at HRTs below 6.5 h was limited by partial acetate oxidation, since alkalinity production by acetate oxidation is necessary for wastewater neutralization.

After a performance failure caused by intentional overloading, the FBR process was recovered in a few days by adjusting the FBR liquid pH to 7 and interrupting the feed pumping for one day. Biomass yield in the FBR processes was 0.049-0.054 g dry biomass / g of lactate oxidized or 0.053-0.074 g dry biomass / g of sulfate reduced. Zinc and iron precipitated predominantly as ZnS, FeS2 and FeS. The metal precipitates did not cause clogging of the reactors during the continuous operation of over 1200 days.

Substrate utilization and sulfide inhibition kinetics were determined with batch FBR-kinetic experiments. Acetate oxidation was the rate-limiting step in ethanol oxidation. Michaelis- -1 -1 Menten constants (Km) were 4.3-7.1 mg l and 2.5-3.5 mg l for ethanol and acetate -1 oxidation, respectively. The maximum oxidation velocities (Vmax) were 0.19-0.22 mg gVS min-1 and 0.033-0.035 mg gVS-1 min-1, for ethanol and acetate, respectively. Noncompetitive inhibition model described the sulfide inhibition of the sulfate-reducing enrichment culture. -1 Dissolved sulfide inhibition constants (Ki) for ethanol and acetate oxidation were 248 mg S l -1 -1 and 356 mg S l , respectively, and the corresponding Ki values for H2S were 84 mg S l and 124 mg S l-1, for ethanol and acetate, respectively.

Bacterial communities of the FBRs were characterized using culture-independent molecular methods (clone library and denaturing gradient gel electrophoresis, DGGE), phenotypic i methods (phospholipids fatty acid profiling, PLFA) as well as bacterial isolations. The bacteria were identified based on 16S rRNA gene sequences. The FBR communities were diverse given that they were enriched and long-term maintained with simple electron donors. The FBR communities, originally enriched from methanogenic granular sludge and mine sediments, contained SRB related to members of the genera Desulfovibrio, Desulfobulbus, Desulfotomaculum, Desulfobacca and Desulforhabdus, and also many species that do not reduce sulfate. The FBR communities contained many previously undescribed bacteria. The clone libraries showed a significant difference in the communities of the lactate- and ethanol- utilizing FBRs, and DGGE indicated some changes in the communities over time. The PLFA analyses showed some signatures consistent with sulfate-reducing communities, but not any substantial difference in the microbial communities of the reactors. Many of the strains isolated from the FBRs were only distantly related to previously described species and, thus, may represent novel species or genera. The combination of molecular and culturing methods allowed the detection of more species in the sulfate-reducing FBR communities than either approach alone.

This study demonstrated the feasibility of the sulfidogenic FBR for concomitant removal of acidity, metals and sulfate from wastewaters. Process design parameters for the FBR process were determined. The study also contributed to the understanding of the wide microbial diversity in the sulfate-reducing FBRs fed with simple electron donors only. Further, several bacterial strains possibly representing novel species or genera were isolated and characterized.

ii TIIVISTELMÄ

Tässä työssä kehitettiin rikkivetyä tuottava leijupetireaktori (FBR) prosessi happaman metalli- ja sulfaattipitoisen jäteveden käsittelyyn, määritettiin prosessin suunnitteluparametrit sekä kuvattiin FBR:n bakteerilajistoa. Prosessi perustuu biologiseen sulfaatin pelkistykseen sulfaatinpelkistäjäbakteerien (SRB) avulla, metallien saostamiseen sulfideina biologisesti tuotetulla H2S:llä, ja veden neutralointiin SRB:ien tuottamalla bikarbonaattialkaliniteetilla. Sulfaatin pelkistyksen elektronidonorina käytettiin laktaattia ja etanolia.

Jatkuvatoimisen FBR-prosessin toimintaa tutkittiin erilaisissa käyttöolosuhteissa. Sulfaatin pelkistyksen, rikkivedyn tuoton ja käsitellyn veden alkaliniteetin perusteella 10 % leijutuksella toimiva FBR oli reaktorin käynnistysvaiheessa tehokkaampi kuin 20-30 % leijutuksella toimivat FBR:t. Korkeilla kuormituksilla 20-30 % leijutuksella toimivat FBR:t olivat kuitenkin tehokkaampia. FBR:iin syötetyn substraatin vaihto laktaatista etanoliksi ei merkittävästi vaikuttanut FBR:n toimintaan.

Bioprosessin kuormitettavuutta tutkittiin nostamalla asteittain syöttöveden sinkki-, sulfaatti- ja substraattipitoisuuksia tai laskemalla hydraulista viipymää (HRT), kunnes prosessihäiriöitä ilmaantui. FBR:t saostivat 350-360 mg Zn l-1 d-1 ja 86 mg Fe l-1 d-1 16 h HRT:llä happamasta jätevedestä, jossa oli 230-240 mg Zn l-1 ja 57-58 mg Fe l-1. Prosentuaalinen metallien saostus oli yli 99.8 % ja käsitellyn veden liukoisen Zn ja Fe pitoisuudet alle 0.1 mg l-1. Laktaatin ja etanolin hapetuksesssa muodostunut alkaliniteetti nosti jäteveden pH:n arvosta 2.5 arvoon 7.5-8.5.

Suurin metallien saostusnopeus saavutettiin etanolia hapettavassa FBR:ssä 6.5 h HRT:llä jätevedestä, jossa oli 176 mg Zn l-1 ja 87 mg Fe l-1. Saostusnopeudet olivat yli 600 mg Zn l-1 d-1 ja 300 mg Fe l-1 d-1, prosentuaalinen metallien saostus yli 99.9 % ja käsitellyn veden liukoiset metallipitoisuudet alle 0.08 mg l-1. Etanolin hapetuksessa tuotettu alkaliniteetti nosti jäteveden pH:n arvosta 3 arvoon 7.9-8.0. Etanolin hapetus oli epätäydellistä 6.5 h HRT:llä ja asetaattia kertyi FBR:iin. FBR:n toimivuutta alle 6.5 h HRT:llä rajoitti asetaatin osittainen hapettuminen, sillä asetaatin hapetuksesta saatavaa alkaliniteettia tarvitaan jäteveden neutralointiin.

Tarkoituksella aikaansaadun ylikuormituksen aiheuttaman toimintahäiriön jälkeen FBR- prosessi toipui muutamassa päivässä, kun reaktorissa olevan veden pH säädettiin 7:ään ja jäteveden syöttövirta katkaistiin vuorokaudeksi. FBR-prosessin biomassasaanto oli 0.049- 0.054 g kuivaa biomassaa / g hapetettua laktaattia tai 0.053-0.074 g kuivaa biomassaa / g pelkistettyä sulfaattia. Sinkki- ja rauta saostumien pääasiallinen koostumus oli ZnS, FeS2 ja FeS. Metallisaostumat eivät aiheuttaneet reaktorien tukkeutumista yli 1200 päivää kestäneiden jatkuvatoimisten kokeiden aikana.

Substraatin hapetuksen ja sulfidi-inhibition kinetiikkaa tutkittiin panoskokeina FBR:ssa. Asetaatin hapetus oli etanolin hapetuksen hitain vaihe. Etanolin ja asetaatin hapetuksen -1 -1 Michaelis-Menten vakiot (Km) olivat 4.3-7.1 mg l ja 2.5-3.5 mg l . Etanolin ja asetaatin -1 -1 maksimaaliset hapetusnopeudet (Vmax) olivat 0.19-0.22 mg gVS min ja 0.033-0.035 mg gVS-1 min-1. Ei-kilpaileva inhibitiomalli (noncompetitive inhibition model) kuvasi sulfaattia pelkistävän rikastusviljelmän sulfidi-inhibitiota. Liukoisen sulfidin inhibitiovakiot (Ki) -1 -1 etanolin ja asetaatin hapetukselle olivat 248 mg S l ja 356 mg S l , ja vastaavat Ki arvot -1 -1 H2S:lle 84 mg S l ja 124 mg S l .

iii Laktaatin ja etanolin avulla H2S tuottavien FBR:ien bakteeriyhteisöjä tutkittiin bakteerien viljelystä riippumattomien molekulaaristen menetelmien (kloonikirjasto, denaturoiva gradienttigeelielektroforeesi = denaturing gradient gel electrophoresis, DGGE), fenotyyppisten menetelmien (fosfolipidirasvahapot = phospholipid fatty acids, PLFA) sekä bakteerieristysten avulla. Bakteerit tunnistettiin 16S rRNA geenin sekvenssin perusteella. FBR:ien bakteeriyhteisöt olivat monimuotoiset ottaen huomioon, että ne oli rikastettu sekä pitkän aikaa ylläpidetty käyttäen yksinkertaisia elektronidonoreita. Alun perin metanogeenisestä granulalietteestä ja kaivossedimenteistä rikastetuissa FBR-yhteisöissä oli seuraavia SRB-sukuja: Desulfovibrio, Desulfobulbus, Desulfotomaculum, Desulfobacca ja Desulforhabdus sekä monia lajeja, jotka eivät pelkistä sulfaattia. FBR-yhteisöissä oli monia aiemmin kuvaamattomia lajeja. Kloonikirjastojen perusteella laktaattia- ja etanolia- hapettavien FBR:ien yhteisöt olivat keskenään hyvin erilaiset. DGGE analyysi osoitti bakteeriyhteisöissä tapahtuneen joitakin muutoksia kokeiden aikana. PLFA profiileissa oli sulfaatinpelkistäjäyhteisöille ominaisia rasvahappoja, mutta laktaattia- ja etanolia-hapettavien FBR-yhteisöjen profiileissa ei havaittu suurta eroa. Monet FBR:sta eristetyt bakteerikannat olivat kaukaista sukua aiemmin kuvatuille lajeille ja siten ne saattavat edustaa aiemmin kuvaamattomia lajeja tai sukuja. Molekyylibiologisten menetelmien ja bakteerien eristämisen yhdistelmä antoi monipuolisemman kuvan FBR:ien bakteeriyhteisöjen koostumuksesta kuin kumpikaan lähestymistapa yksinään.

Tämä tutkimus osoitti, että sulfidia tuottavat FBR:t soveltuvat happamuuden, metallien ja sulfaatin yhtäaikaiseen poistoon jätevesistä. Prosessin suunnitteluparametrit määritettiin. Tutkimus lisäsi myös ymmärtämystä yksinkertaisilla elektronidonoreilla syötettyjen sulfaattia pelkistävien FBR-yhteisöjen mikrobidiversiteetistä. Lisäksi työssä eristettiin useita bakteereita, jotka saattavat olla edustaa uusia lajeja tai sukuja.

iv PREFACE AND ACKNOWLEDGEMENTS

This thesis is based on the work carried out at the Institute of Environmental Engineering and Biotechnology, Tampere University of Technology (TUT), Tampere, Finland and at the Land and Water Division, Commonwealth Scientific and Industrial Research Organization (CSIRO), Floreat, Australia. The work was conducted as a part of Mining Biotechnology project in cooperation with Outokumpu Research Oy and the Biotechnology Unit of Technical Research Center of Finland.

I am grateful to my supervisor professor Jaakko Puhakka for introducing me into the world of mining biotechnology, providing excellent research facilities and an inspiring research environment. I sincerely thank Dr. Peter Franzmann for giving me the opportunity to do part of this work at CSIRO and for advising me with the microbiological work. I am grateful to Dr. Jason Plumb, Dr. Wendy Robertson, Dr. Amanda Tilbury and Dr. John Gibson for the help with the microbial techniques. I thank Dr. Steve Fisher and Lic. Marja Riekkola- Vanhanen, Outokumpu Research Oy, for the advice and support with the chemical analysis. Further, I would like to thank Dr. Olli Hyvärinen and Lic. Mikko Ruonala, Outokumpu Research Oy, for the advice and inspiring discussions. I am grateful to Dr. Piet Lens, Wageningen University and professor Olli Tuovinen, Ohio State University, for pre- reviewing this thesis and for their valuable comments and suggestions.

I greatly acknowledge Dr. Päivi Kinnunen, who shared the experiences in the Mining Biotechnology project. Tarja Karjalainen, Sook Leng Thong, Maija Laurikkala and Annukka Hämäläinen are thanked for their excellent assistance in the lab. I would like to express my gratitude to my friends and colleagues for their support and friendship. My warmest thanks are dedicated to my dear husband Teemu and my family for all the patience, encouragement and love.

National Technology Agency of Finland, Outokumpu Oyj, Finnisch Graduate School of Environmental Science and Technology, the Academy of Finland, European Commission, the Scientific Foundation of the City of Tampere, Land and Water Technology Foundation, The Foundation of Technology, and Konkordia Foundation are gratefully acknowledged for their financial support.

v

TABLE OF CONTENTS

ABSTRACT...... i

TIIVISTELMÄ...... iii

PREFACE AND ACKNOWLEDGEMENTS...... v

TABLE OF CONTENTS...... vi

LIST OF ORIGINAL PAPERS ...... viii

THE AUTHOR’S CONTRIBUTION...... ix

ABBREVIATIONS...... x

1 INTRODUCTION...... 1

2 MINING AND MINERAL PROCESSING IN FINLAND...... 3

3 SULFIDE MINERALS...... 6

4 WASTEWATERS FROM MINING AND MINERAL PROCESSING ...... 7 4.1 Formation and chemical characteristics of wastewaters ...... 7 4.2 Environmental impacts of wastewaters...... 12

5 SOURCE AND MIGRATION CONTROL OF ACID MINE DRAINAGE...... 14 5.1 Isolation of pyritic materials ...... 14 5.2 Alkaline addition...... 16 5.3 Immobilization...... 16 5.4 Phosphate application...... 16 5.5 Sacrificial anodes...... 16 5.6 Microbial control...... 17

6 WASTEWATER TREATMENT...... 18 6.1 Abiotic processes...... 18 6.1.1 Passive limestone neutralization...... 18 6.1.2 Chemical precipitation...... 18 6.1.3 Alternative physico-chemical treatment techniques...... 21 6.2 Biological processes...... 24 6.2.1 Passive treatment methods ...... 25 6.2.2 Bioreactor processes...... 29

7 SULFATE-REDUCING PROKARYOTES...... 39 7.1 Taxonomy ...... 39 7.2 Physiology...... 40 7.3 Ecological and economical importance ...... 44

8 HYPOTHESES AND AIMS OF THE PRESENT WORK ...... 46

9 MATERIALS AND METHODS...... 48 9.1 Bioreactors ...... 48 9.2 Continuous flow experiments ...... 50 9.3 Batch kinetic experiments...... 50 vi 9.4 Kinetic calculations...... 51 9.5 Physico-chemical analyses...... 52 9.6 Microbiological analyses...... 52

10 RESULTS AND DISCUSSION ...... 54 10.1 Start-up and process performance...... 54 10.1.1 Loading capacity...... 54 10.1.2 The effect of hydraulic retention time on fluidized-bed reactor performance ....60 10.1.3 Biomass yield...... 62 10.1.4 Composition and fate of metal precipitates...... 62 10.1.5 Molar stoichiometry of the reactions in the fluidized-bed reactors...... 64 10.2 Electron donor oxidation and sulfide inhibition kinetics ...... 66 10.2.1 Ethanol and acetate oxidation kinetics ...... 66 10.2.2 Sulfide inhibition kinetics...... 68 10.3 Microbiology of the fluidized-bed reactors ...... 71 10.3.1 Clone libraries...... 71 10.3.2 Denaturing gradient gel electrophoresis...... 72 10.3.3 Bacterial isolates...... 72 10.3.4 Phospholipid fatty acid profiling...... 76

11 CONCLUSIONS...... 79

12 RECOMMENDATIONS FOR FUTURE RESEARCH...... 81

13 REFERENCES...... 83

vii LIST OF ORIGINAL PAPERS

This thesis is based on the following original papers referred to, in this thesis, by the roman numerals:

I Kaksonen, A.H., Riekkola-Vanhanen, M.-L. and Puhakka, J.A. 2003. Optimization of metal sulphide precipitation in fluidized-bed treatment of acidic wastewater. Water Research 37: 255-266.

II Kaksonen, A.H., Franzmann, P.D. and Puhakka, J.A. 2003. Performance and ethanol oxidation kinetics of a sulfate-reducing fluidized-bed reactor treating acidic metal- containing wastewater. Biodegradation 14: 207-217.

III Kaksonen, A.H., Franzmann, P.D. and Puhakka, J.A. 2004. Effects of hydraulic retention time and sulfide toxicity on ethanol and acetate oxidation in sulfate- reducing metal-precipitating fluidized-bed reactor. Biotechnology and Bioengineering 86: 332-343.

IV Kaksonen, A.H., Plumb, J.J., Franzmann, P.D. and Puhakka, J.A. 2004. Simple organic electron donors support diverse sulfate-reducing communities in fluidized- bed reactors treating acidic metal- and sulfate-containing wastewater. FEMS Microbiology Ecology 47: 279-289.

V Kaksonen, A.H., Plumb J.J., Robertson, W.J., Franzmann, P.D., Gibson, J.A.E. and Puhakka, J.A. 2004. Culturable diversity and community fatty acid profiling of sulfate-reducing fluidized-bed reactors treating acidic metal-containing wastewater. Geomicrobiology Journal (in press).

Paper I reprinted with kind permission of Elsevier Science Ltd. Paper II reprinted with kind permission of Kluwer Academic Publishers. Paper III reprinted with kind permission of the Wiley Periodicals, Inc. Paper IV reprinted with kind permission of Elsevier B.V. Paper V reprinted with kind permission of Taylor & Francis, Inc.

viii THE AUTHOR’S CONTRIBUTION

Paper I: Anna Kaksonen wrote the paper and is the corresponding author. She planned and performed the experimental work and interpreted the results except for the elemental and mineralogical analysis of the metal precipitates, which was done by Lic. M.-L. Riekkola-Vanhanen.

Paper II: Anna Kaksonen wrote the paper and is the corresponding author. She planned and performed the experimental work and interpreted the results.

Paper III: Anna Kaksonen wrote the paper and is the corresponding author. She planned and performed the experimental work and interpreted the results.

Paper IV: Anna Kaksonen wrote the paper and is the corresponding author. She planned and performed the experimental work and interpreted the results. Dr. J.J. Plumb advised the Denaturing gradient gel electrophoresis and the DNA sequence analysis.

Paper V: Anna Kaksonen wrote the paper and is the corresponding author. She planned and performed the experimental work and interpreted the results except for the analysis of phospholipid fatty acids, which was done by Dr. J.A.E. Gibson. Dr. J.J. Plumb and Dr. W.J. Robertson advised the anaerobic culture work.

The experimental work was carried out under the supervision of Prof. J.A. Puhakka (papers I- V) and Dr. P.D. Franzmann (Papers II-V) who also provided the research facilities.

ix ABBREVIATIONS

ABR Anaerobic baffled reactor ACP Anaerobic contact process ADP Adenosine diphosphate AFR Anaerobic filter reactor AHR Anaerobic hybrid reactor ALD Anoxic limestone drain AMD Acid mine drainage AMP Adenosine monophosphate APS Adenosine 5'-phosphosulfate ARD Acid rock drainage ATP Adenosine triphosphate CFB Cytophaga-Flexibacter-Bacteroides CSTR Continuously stirred tank reactor DGGE Denaturing gradient gel electrophoresis DNA Deoxyribonucleic acid DIC Dissolved inorganic carbon DOC Dissolved organic carbon DS Dissolved sulfide EDS Energy dispersive x-ray spectrometer FBR Fluidized-bed reactor FISH Fluorescent in situ hybridization GC % Guanine plus Cytosine mol % of DNA HRT Hydraulic retention time I Inhibitor concentration Ki Inhibition constant Km Michaelis-Menten constant OLC Open limestone channel OTU Operational taxonomic unit PAP Phosphoadenosine 5'-phosphate PAPS Phosphoadenosine 5'-phosphosulfate PCR Polymerase chain reaction pKa -logKa PLFA Phospholipid fatty acid PPi Pyrophosphate rRNA Ribosomal RNA S Initial substrate concentration SEM Scanning electron microscope SFS Finnish Standards Association SRA Sulfate-reducing archaea SRB Sulfate-reducing bacteria SRP Sulfate-reducing prokaryotes SRT Sludge retention time T Temperature TS Total solids TSS Total suspended solids UASB Upflow anaerobic sludge blanket reactor ν Oxidation velocity Vmax Maximum oxidation velocity VS Volatile solids VSS Volatile suspended solids XRD X-ray diffraction x 1 INTRODUCTION

The exploitation of sulfide minerals in mining and metallurgy results in the chemical and biological oxidation of iron and sulfur, and thus in the formation of acidic metal- and sulfate- containing wastewaters, often called as acid mine drainage (AMD) or acid rock drainage (ARD) (Dugan 1975; van Houten and Lettinga 1995; Foucher et al. 2001; Johnson and Hallberg 2002). The environmental impacts of AMD are severe and widespread in many countries (Jarvis and Younger 2000). Pollution control of AMD can be achieved by preventing AMD formation and migration as well as collecting and treating contaminated waters (Geldenhuis and Bell 1998; Johnson 2000). Numerous physico-chemical and biological techniques are available for removing acidity, metals and sulfate from wastewaters (for reviews, see Lanouette 1977; Peters et al. 1985; Johnson 2000; Hao 2000).

The most widely used active treatment process for AMD is based on chemical neutralization and hydroxide precipitation of metals (Lanouette 1977; Peters et al. 1985; Veeken and Rulkens 2003). The disadvantages of the traditional chemical treatment include high cost of the chemical reagents, inefficient removal of sulfate, and production of a bulky sludge, which must be disposed of (Tichý et al. 1998; García et al. 2001). Sulfide precipitation of metals has been demonstrated to have several benefits over the hydroxide precipitation, such as lower effluent metal concentrations, better thickening characteristics of the metal sludge and the possibility to recover valuable metals (Whang 1982; Peters et al. 1985; Boonstra et al. 1999; Jalali and Baldwin 2000; Veeken and Rulkens 2003). Chemical sulfide precipitation has not been widely used for AMD treatment, probably due to the high cost of chemicals (Lanouette 1977; van Houten and Lettinga 1995).

Efforts have been made to develop biological alternatives for AMD treatment. Recently, the interest in the application of sulfate reduction as the dominant step of wastewater treatment has been growing (for reviews, see Hulshoff Pol et al. 2001; Lens et al. 2002). The process is based on biological hydrogen sulfide production by sulfate-reducing bacteria (SRB), metal sulfide precipitation and neutralization of the water with the bicarbonate alkalinity produced by the microbial metabolism (Dvorak et al. 1992; Christensen et al. 1996). Biological sulfate reduction can be enhanced by providing a carbon and electron source for the SRB. SRB- utilizing wastewater treatment systems range from wetlands (Eger 1994) and reactive barriers (Herbert Jr. et al. 1998; Waybrant et al. 1998; Benner et al. 1999; Amos and Younger 2003) to bioreactor applications (for review, see Johnson 2000). Among the bioreactor applications, various immobilized biomass systems appear to be most promising, since they allow high loading rates (Speece 1983). Much research has been done on the use of anaerobic filter reactors (AFR) and upflow anaerobic sludge blanket reactors (UASB) for AMD treatment, whereas the potential of fluidized-bed reactors (FBR) for concomitant removal of acidity, metals and sulfate has not been previously extensively studied.

Anaerobic FBRs have become a well established technology to treat concentrated industrial effluents containing organic contaminants (for a review, see Saravanane and Murthy 1999). The FBR process has various potential advantages over other high rate anaerobic reactors, such as AFR and UASB. These are: no channeling or clogging of reactors, good biomass retention, high biomass activity, high treatment efficiency and least chance for shock loads (Somlev and Tishkov 1992; Saravanane and Murthy 1999). Due to the high biomass retention and the dilution of the feed water with the recycling flow, it was hypothesized that sulfidogenic FBRs would be efficient in treating AMD.

1 The present study focuses on the treatment of acidic metal- and sulfate-containing wastewaters using sulfidogenic FBRs. In the experimental part of this work a) the FBR process performance was evaluated under various operational conditions with continuous flow experiments, b) substrate utilization and sulfide inhibition kinetics was determined with batch FBR experiments, and c) bacterial communities of the FBRs were characterized using culture-independent molecular methods and phenotypic methods as well as bacterial isolations. The summary part of the work reviews literature on a) mining and mineral processing in Finland, b) sulfide minerals, c) the formation, properties and environmental impacts of AMD, d) various methods for AMD pollution control with the emphasis on sulfate-reducing bioreactors and e) sulfate-reducing prokaryotes.

2 2 MINING AND MINERAL PROCESSING IN FINLAND

Finland has a long history of mining activity, which dates back to 1540s when iron ore mining commenced. Between years 1533 and 2001 at least 418 metal ore mines (Figures 1 and 2), and 616 industrial mineral and carbonate stone mines have been in operation (Figure 2) (Puustinen 2003). The exploitation of metal ores has included Au, Ag, Pt, Cu, Pb, Zn, Sn, Fe, Ni, Cr, Co, MoS2, WO3, TiO2, V2O5, U and As deposits, and provided the raw material base for the country's metal industry (Puustinen 2003). At present, five metal ore mines are in operation (Figure 1). The Figure 3 shows a photo of Pyhäsalmi mine, an open pit and underground Zn and Cu mine in North Ostrobothnia, Finland (Anonymous 2004a). The major industrial minerals mined in Finland are limestone, apatite, talc, wollastonite, feldspar and quartz (Anonymous 2004b). The total extend of mining has been at least 1027 million tons of stone, of which 698 million tons has been ore. The total extent of mining in metal ore mines has been over 481 million tons of stone, of which over 271 million tons has been ore (Figure 4) (Puustinen 2003). About 66 % of the metal ores has been sulfide and 34 % oxide ores (Anonymous 2004b). The number of metallurgical mills in Finland has been at least 134 (Figure 5), and currently mills exist for Cu, Ni, Zn, Co, Cr and Fe ores (Puustinen 2003; Anonymous 2004b).

Figure 1. Metal ore mines in Finland and former Finnish areas (Puustinen 2003). Operating mines are named on the map (Anonymous 2004c).

3 60 Metal ore mines 50 All mines 40

30

20 Number of mines 10

0 1500 1550 1600 1650 1700 1750 1800 1850 1900 1950 2000 Year

Figure 2. Number of metal ore mines and all mines in Finland between 1530 and 2002 (data obtained from Puustinen 2003).

Figure 3. An open pit and underground Pyhäsalmi mine in North Ostrobothnia, Finland, view to the North-West (Anonymous 2004a) (Printed with kind permission of Outokumpu Oyj.). Sediments obtained from the Pyhäsalmi mine were used as a source of sulfate-reducing bacteria in the present study.

The extent of mining expanded rapidly in the 1970s, and still the Nordic countries (Finland, Sweden and Norway) are in European terms the leading producer of iron, chromite and gold, and the second biggest producer of zinc (Ellis 2003). The ancient Fennoscandian Shield of Northwestern Europe has long been regarded as prime exploration territory, particularly by the established mining companies in the region. Alone in Finland, 41 million Euros were invested in exploration in 2002. An increasing number of overseas companies, from countries such as Australia, Canada and the UK have also become interested in Scandinavian ore deposits. The objects of exploration have been e.g. gold, platinum, cobalt and base metal deposits as well as potential diamond-bearing minerals (Ellis 2003).

4 40 Metal ore mines, total mining including wasterock 35 Metal ore mines, ore mining 30 All mines, total mining including wasterock

s 25 All mines, ore mining 20 15 Million ton 10 5 0 1500 1550 1600 1650 1700 1750 1800 1850 1900 1950 2000 Year

Figure 4. Mining in Finland between years 1530 and 2002 (data obtained from Puustinen 2003). The major industrial minerals mined at non-metallic mines include limestone, apatite, talc, wollastonite, feldspar and quartz (Anonymous 2004b).

Figure 5. Metallurgical mills in Finland and former Finnish areas (Puustinen 2003). Major operating mills are named on the map (Anonymous 2004b; Anonymous 2004d).

5 3 SULFIDE MINERALS

Many economically important metals are present in nature combined with sulfur as metal sulfides (Table 1) (Woods 2004). The general formula of sulfides is given as MmSn in which M represents the metallic elements and S sulfur (Klein and Hurlbut 1985). Pyrite (FeS2) is ubiquitous in most metal sulfide and coal deposits and may exist together with other chalcophile elements (Banks et al. 1997). Other common sulfide minerals in Finnish deposits are chalcopyrite (CuFeS2), sphalerite (ZnS), galena (PbS), arsenopyrite (FeAsS), pentlandite ((Fe,Ni)9S8)), chalcocite (Cu2S), cubanite (CuFe2S3), cobaltite (CoAsS), millerite (NiS) and molybdenite (MoS2) (Siivola 1986; Anonymous 2004e).

Table 1. Metals and metalloids extracted from the mineral groups of the major ore minerals (adapted from Craig 1989). Mineral group Principal metals and metalloids extracted Elemental metals Au, Ag, Pt, Bi, Cu Sulfides Cu, Pb, Zn, Hg, Co, Mo, Ni, Ag, Sb, Cd, Ga, Ge, In, Re, Tl, Au Oxides Fe, Mn, Mo, Ta, Al, Ti, Sn, W, U, Nb Silicates Be, Li, Zn, Sc, U Arsenides As, Pt, Co, Ni Phosphates Rare earth metals Carbonates Rare earth metals Tellurides Au, Te

Mineral deposits are developed by various processes of formation in igneous, sedimentary and metamorphic rocks (Carr and Herz 1989). Igneous rocks are formed by crystallization or precipitation from molten magma and its dissolved fluids, either below or on the surface of the earth (Jensen 1989a). Most metal sulfides, including those of commercial interest, are of igneous origin (Ehrlich 2002). Sedimentary rocks are formed by deposition in seas or lakes, or on the surface of the earth (Jensen 1989a). Sulfides in sedimentary rocks are thought to be of biogenic origin (Shimazaki et al. 1985). Most biogenic metal sulfides are associated with bacterial sulfate reduction in anaerobic environments, such as organic-rich sediments (Shimazaki et al. 1985; Ehrlich 2002). Metamorphic rocks are formed from pre-existing rocks in environments of high temperature and pressure below the surface of the earth (Jensen 1989a). The majority of mineral deposits formed by metamorphism are nonmetallic and sulfidic, e.g. asbestos, talc and graphite (Jensen 1989b).

When sulfide ore deposits are exposed to weathering processes, they develop a zonal arrangement of different mineral associations, which reflect different degrees of oxidation. The electrochemical interaction between ore-bodies and ground water involves anodic oxidation of the sulfide and the simultaneous cathodic reduction of oxygen. Since the sulfide ore-body is conducting, the anodic and cathodic processes can occur at different places. The electrons will flow through the ore-body to the surface and as a result the top layer of the ore has a more positive potential than that at greater depth. The alteration of the composition of the ore-mineral will be greatest closer to the surface. For example, the alteration of copper ore-deposits containing the major economic copper mineral, the copper iron sulfide (chalcopyrite), often occur to form an overlying copper sulfide (chalcocite) zone and a cap containing oxidized copper minerals and elemental copper (Woods 2004).

6 4 WASTEWATERS FROM MINING AND MINERAL PROCESSING

4.1 Formation and chemical characteristics of wastewaters

Mining results in the introduction of oxygen and water to the deep geological environment and thus the oxidation of minerals, which are in a reduced state (Banks et al. 1997; Christensen et al. 1996). Oxidation also occurs when reduced minerals are brought to the surface and deposited in heaps. The most common family of reduced minerals are the sulfides (Banks et al. 1997). Sulfide minerals are stable under dry anoxic conditions, such as undisturbed ore deposit (Robb 1994). The oxidation of the sulfides of the type MS2 leads to production of protons (Banks et al. 1997). The oxidation of pyrite (FeS2) can be described with the following reaction (Banks et al. 1997):

2+ 2- + 2 FeS2 + 7 O2 + 2 H2O Æ 2 Fe + 4 SO4 + 4 H (1)

The oxidation of ferrous to ferric iron consumes protons according to the following reaction (Banks et al. 1997):

2+ + 3+ 4 Fe + 4 H + O2 Æ 4 Fe + 2 H2O (2)

Ferric iron may act as an electron acceptor for further pyrite oxidation, or hydrolysis may occur, both processes releasing further protons (Banks et al. 1997):

3+ 2+ 2- + FeS2 + 14 Fe + 8 H2O Æ 15 Fe + 2 SO4 + 16 H (3)

3+ + Fe + 3 H2O Æ Fe(OH)3 + 3 H (4)

The overall sequence of reactions is acid-producing (Banks et al. 1997):

2- + 4 FeS2 + 14 H2O + 15 O2 Æ 4 Fe(OH)3 + 8 SO4 + 16 H (5)

Other sulfide minerals are oxidized in similar way as pyrite, releasing metals and sulfate in solution. However, the oxidation of sulfides of the type MS does not release acid, e.g. sphalerite oxidation (Banks et al. 1997):

2+ 2- ZnS + 2 O2 Æ Zn + SO4 (6)

Acidic conditions result in further dissolution of heavy metals from metal oxides and carbonates (van Houten and Lettinga 1995). The oxidation rate of the sulfide minerals is dependent on many factors, such as mineral composition, particle size (Robb 1994), Fe3+ concentration and the availability of oxygen (Clarke 1995). The oxidation of Fe2+ (Reaction 2) appears to be the rate limiting step in chemical oxidation pyrite (Singer and Stumm 1970). Abiotic Fe2+ oxidation increases as pH increases (Evangelou et al. 1998). In low pH environments, bacteria, such as Acidithiobacillus ferrooxidans, can accelerate the Fe2+ oxidation even 106 fold as compared to chemical oxidation (Robb 1994). Other iron oxidizing bacteria include species from genera Leptospirillum and Sulfobacillus. In high temperatures, the catalysis of iron oxidation may be provided by thermophilic Archaea, such as members of the genera Ferroplasma, Sulfolobus, Metallosphaera and Acidianus (for a review, see Rawlings 2002). The microbial activity is dependent on species, temperature, pH, nutrients, and O2 and CO2 availability, mineral substrate, and the presence of heavy metals, surfactants and organic compounds (for review, see Bosecker 1997).

7 The formation of acidic wastewaters can continue even tens and hundreds of years after mine closure, if the conditions remain favourable (Béchard et al. 1994; Szczepanska and Twardowska 1999). In Finland, the responsibility for remediating mine sites and compensating for environmental damage remains with the mining company indefinitely (Anonymous 1994; Anonymous 2000). The quality of the mine waters may vary depending on the local conditions, and over time (Clarke 1995). Examples of the chemical characteristics of waters at Finnish mine sites are presented in Table 2. For comparison, Table 2 also shows the median values for Finnish stream- and groundwaters as well as national drinking water limits and recommendations for household wells and small water cooperatives.

The acidity generated by pyrite oxidation results in a significant decrease in pH of the water, only when it exceeds the alkalinity available as bicarbonate in the water, or the alkalinity in the form of mineral phases (Banks et al. 1997). These are not, however, always present in significant quantities (Banks et al. 1997). The acidity of the mine waters results from proton and mineral acidity (soluble metals) (Clarke 1995). Most significant metals causing mineral acidity are Fe, Al and to some extend Mn. The acidity of metals is due to their hydrolysis (Reactions 4, 7 and 8), which in the case of Fe2+ and Mn2+ is preceded by the oxidation (Johnson 2000).

3+ + Al + 3 H2O Æ Al(OH)3 + 3 H (7)

4+ + Mn + 2 H2O Æ MnOOH + 3 H (8)

The concentration of dissolved inorganic carbon (DIC) in acidic (pH < 3.5) mine waters is usually very low (< 0.5 mg/l) due to the dissociation of carboxylic acid and low solubility of carbon dioxide. Also the concentration of dissolved organic carbon (DOC) is often low, usually < 10 mg l-1 and comparable to neutral oligotrophic watercourses. Part of the organic carbon may come from the degradation of old wooden pillars in the mines (Johnson 2000). In coal mines, trace contamination by organic compounds derived from coal is also possible (Banks et al. 1997).

Ammonium is usually the dominant form of nitrogen in mine waters (Johnson 2000). Considerable amounts of nitrate may be present, if N-based explosives have been used for mining (Banks et al. 1997). Phosphorous is the most common nutrient limiting primary production in mining ponds (Johnson 2000). Occasionally, mine waters may have high salinity. This may be caused by deep mining, which encroach upon bodies of stagnant, or even fossil, groundwater of high salinity. Dewatering of coastal mines may also cause the intrusion of seawater with high chlorine content (Banks et al. 1997).

In addition to metal mining, AMD formation may also occur in coal mines (Figure 6). Coal contains both inorganic sulfur (mostly pyrite and sulfate) and organic sulfur (mostly sulfides, tiofenes, sulfoxides and sulfones). The sulfur-content of coal is usually 1-10 % (Johnson 2000). The Figure 6 shows AMD formed at a South-African coal mine. The red colour of the water comes from the high iron concentration. Acidic metal-containing wastewaters are formed not only in mine shafts, heaps and tailings, but also at metallurgical mills processing sulfide minerals. In the present work the term AMD is used for all of these wastewaters. The Figure 7 shows wastewater streams from the Norilsk mining and smelting complex, which is located above the Arctic Circle on the Taimyr Peninsula of northern Siberia. The Norilsk mining and smelting complex (Polar Division of Norilsk Nickel Group) is the largest nickel and palladium producer in the world and has a substantial market share in platinum, copper, and cobalt production (Karnachuk et al. submitted). Figure 8 shows the annual wastewater

8 volumes and wastewater metal loads from Finnish mines and metal processing mills between years 1995 and 2002.

Figure 6. Acid mine drainage waters from a South African coal mine (Photos: Anna Kaksonen). The red colour of the water is due to the presence of iron.

Figure 7. Wastewater streams from the Norilsk mining and smelting complex, Russia (Photo: Jaakko Puhakka).

9

Table 2. Examples of the chemical characteristicsa of waters at Finnish mine sites. Median values for Finnish stream- and groundwaters and national drinking water limits and recommendations for household wells and small water cooperatives are presented for comparison.

Surface waters Groundwater Finnish reference values Groundwater Drinking water Paroistenjärvi Hitura Ruostesuo Kangasjärvi Otravaara Stream Hitura mine from drilled limits and mine mine mine mine mine waters wells recommendations pH 3.1-5.8 6.3-6.6 3.0 3.7 2.2-2.3 5.4-8.3 5.9 6.6 6.5-9.5 (Rb) 2- SO4 168-872 1400-3680 4885 900 6260-9440 3900-17250 3.5 11.3 < 250 (R) Cl 1-6.7 450-840 NRc NR NR 23-1760 1.4 12.3 < 100 (R) Fe 0.5-74 0.02-0.11 233 20 2130-2650 0.02-4.6 0.68 0.1 < 0.2 (R) Al NR 0.01 NR NR 267-395 0.01 0.095 NR < 0.2 (R) As 0.01-0.39 NR NR NR NR NR 0.00036 NR < 0.01 (Ld) Cd NR 0.09 NR NR NR 0.0002-0.0003 0.00002 0.0005 < 0.005 (L) Co NR 0.15-0.29 NR NR 0.99-1.16 0.03-2610 0.00017 NR NR Cu 0.5-20 0.01-0.02 2 2.5 2.8-3.3 0.01-0.21 0.00064 0.009 < 2.0 (L) Mn 0.3-7.6 0.78-2.67 55 6 8.5-14 1.7-5.4 0.029 0.05 < 0.05 (R) Ni NR 1.57-1.88 NR NR 1.3-1.6 0.02-2860 0.00052 0.002 < 0.02 (L) Zn 0.1-3.2 0.1-0.38 557 45 9.2-10.7 0.05-760 0.0036 0.1 NR Reference Carlson et al. Heikkinen Mustikkamäki Mustikkamäki Räisänen et Heikkinen et al. Lahermo et Lahermo et al. Anonymous 2001 2002 et al. 2002 2000 2000 al. 2001 2002 al. 1996 1990 aAll units (except for pH) in mg l-1, bR = recommendation, cNR = Not reported, dL = limit.

10 a 300 25 f 8 100

d Cromium 250 d 3 20 6 80 200 15 60

150 Tons 4 Wastewater 10 40 the total Million m Million 100 the total % share of % share of industrial loa

2 industrial loa 50 5 20 0 0 0 0 g b 300 100 4 100 d 250 Iron d 80 3 80 200 60 60 150 2 Tons Tons

40 40 total the 100 total the % share of % share of

1 Cobolt industrial loa 50 20 industrial loa 20 0 0 0 0

c 12 100 h 1.5 80 d 10 Nickel d 80 60 8 1 60 6 40 Tons Tons 40 the total total the 4 total the 0.5 % share of % share of

20 industrial loa 20 loa industrial 2 Arsenic 0 0 0 0

d 12 100 i 0.2 120 d 10 d Cadmium 100 80 0.15 8 80 60 6 0.1 60 Tons the total Tons Copper 40 total the 4 40 % share of % share of industrial loa

industrial loa industrial 0.05 2 20 20 0 0 0 0

e 12 30 j 0.02 70 10 25 Mercury 60 0.015 8 20 50 40 6 15 0.01 load Zinc 30 load Tons 4 10 Tons 0.005 20 total industrial industrial total total industrial % share of the 2 5 10 % share of the 0 0 0 0 1995 1996 1997 1998 1999 2000 2001 2002 1995 1996 1997 1998 1999 2000 2001 2002 Year Year

Figure 8. Annual (a) wastewater volumes and wastewater (b) iron, (c) nickel, (d) copper, (e) zinc, (f) chromium, (g) cobolt, (h) arsenic, (i) cadmium and (j) mercury loads from Finnish mines (circles) and metal processing mills (triangles) between years 1995 and 2002 (Data obtained from Anonymous 2004f). Filled symbols show the quantitative wastewater volumes and metal loads, and open symbols the mine or metal processing mill percent shares of the total industrial volumes and metal loads.

11 4.2 Environmental impacts of wastewaters

The environmental impacts of AMD are severe and widespread in many countries. In UK alone, over 700 km of the streams and rivers are affected by the mining wastewaters (Jarvis and Younger 2000). The major impact areas are rivers, lakes, estuaries and costal waters (Gray 1997), although metallurgical waters can also contaminate groundwater resources (Barnes et al. 1991a; Barnes et al. 1991b) as well as arable land (Clemente et al. 2003), and damage manmade constructions (Jarvis and Younger 2000). The effects of AMD can be categorized as chemical, physical, biological, ecological and socioeconomic impacts (Table 3) (Gray 1997; Jarvis and Younger 2000).

Table 3. Major effects of acid mine drainage (adapted from Gray 1997; Jarvis and Younger 2000) Chemical Physical Biological Ecological Socioeconomic • Increased • Increased • Behavioural • Habitat • Flooding acidity turbidity • Respiratory modification • Subsidence • Destruction of • Decrease in • Reproduction • Niche loss • Corrosion bicarbonate light • Osmoregulation • Bioaccumulation • Aesthetic loss buffering penetration • Acute and in food chain • Decrease of • Increase in • Sedimentation chronic toxicity • Loss of food drinking, soluble and • Adsorption of • Death of source or prey agricultural, particulate metals into sensitive species • Reduction in industrial and metal sediments and • Acid-base primary recreational concentrations organisms balance failure productivity water quality • Salinization in organisms • Food chain • Decrease in • Migration or modification catch of fish avoidance • Health effects

The AMD affects aquatic ecosystems by a number of direct and indirect pathways (Gray 1997). The acidity of the wastewaters can cause direct toxic effects, e.g. harm fish gills, or influence indirectly by increasing the solubility of toxic metals (Johnson 2000). Iron-rich precipitates hamper fish spawning (Clarke 1995) and smother river sediments, impending oxygen diffusion and killing benthic organisms (Robb 1994; Johnson 2000). The precipitation of iron can also increase turbidity of the receiving stream, reduce light penetration and thus primary production (Robb 1994; Johnson 2000). This may have a dramatic effect on the food chain in impacted waters (Johnson 2000). AMD-impacted water courses tend to be devoid of fish and have limited biodiversity in planktonic and benthic organisms as compared to non- polluted waters (Johnson 2000). Also the microbial community structure and function in AMD impacted waters are very different from those of non-polluted waters (Mills and Mallory 1987).

Mine waters rising in deep mines following mine closure may result in flooding, subsidence, and corrosion of foundations (Jarvis and Younger 2000). Acidic waters are also corrosive to bridges, dams and plumbing (Dugan 1975). Deposition of bright orange, iron-rich precipitates on stream beds has a devastating aesthetic impact (Jarvis and Younger 2000). Severe mine water discharges may pollute water supplies (Jarvis and Younger 2000). Due to increased toxicity and hardness, the recipient waters of the AMD discharge may be unusable for agricultural, industrial as well as for recreational purposes (Dugan 1975; Clarke 1995). For example, fish farming had to be closed down in a Finnish Kangasjärvi-lake due to the acid Zn-containing waters flowing from near-by Kangasjärvi open pit mine in Keitele, Finland (Mustikkamäki 2000). In the Finnish village of Antskog, the pollution from historical Fe and Cu works resulted in a crop failure when sediment material dredged downstream of the metal works was spread out on a nearby farmland (Åström and Nyland 2000). Compared to the limit

12 values for contaminated soils in Finland, the concentrations of Cu, Pb and Zn in the dredged sediment were elevated by factor > 5 (Åström and Nyland 2000). Also high chloride concentrations are harmful to crops and cause corrosion in metal devices (Clarke 1995).

The most drastic environmental effects of wastewaters originating from mining and metallurgy have been caused by sudden accidents, such as the catastrophic spill of about 50 000 m3 of AMD from the former Wheal Jane tin mine, Cornwall, UK, in 1992 (Banks et al. 1997). The AMD was released into the Carnon river and left behind an iron-hydroxide coating on the floor of the Fal estuary (Hallberg and Johnson 2003). Another major accident took place in 1998, when a dam wall enclosing acidic pyritic mine sludge was broken at Aznalcóllar, Seville, Spain, and about 5 million m3 of a highly toxic pyrite waste spread along the Guadiamar river, covering 45 km2 of the surrounding arable land (Clemente et al. 2003). Although most of the sludge, together with the surface soil, was removed, heavy metal levels (especially Zn, Cd and Cu) of soils in some areas were still phytotoxic. Additionally the oxidation of the remaining sulfides to sulfuric acid can lead to further acidification of the soil (Clemente et al. 2003). A third recent accident happened in Baia Mare gold mine in Romania, where a wastewater dam was broken causing 100 000 m3 of cyanide- and heavy metal - containing wastewater to flow into Szamos river, a tributary river of Danube in 2000 (Järvinen 2000; Hallberg and Johnson 2003). The toxic plume killed tons of fish and the harmful effects were seen as far as Hungary and former Yugoslavia (Järvinen 2000).

13 5 SOURCE AND MIGRATION CONTROL OF ACID MINE DRAINAGE

Pollution control of AMD can be achieved by source, migration or release control measures (Geldenhuis and Bell 1998). Source control measures are designed to prevent or limit AMD formation. Migration control relies on the containment of AMD by restricting surface and rain water entry to the source of AMD (Geldenhuis and Bell 1998), and by restricting the movement of contaminated waters (Johnson 2000). Release control is based on collection and treatment of AMD (Geldenhuis and Bell 1998). In some cases, especially at operating mines, this is the only practical option available (Geldenhuis and Bell 1998). Various physical, chemical and biological techniques for source and migration control of acid mine drainage are shown in Table 4 (For a review, see Evangelou and Zhang 1995).

Table 4. Techniques available for source and migration control of acid mine drainage. Technique Principle Reference(s) Exclusion of air Flooding and sealing of underground mines and Evangelou and Zhang 1995; underwater storage of mine tailings Perry et al. 1998 Exclusion of water Storage of waste rock in heaps sealed with e.g. clay Banks et al. 1997; Perry et al. cover or plastic liners 1998 Revegetation Establishing vegetation on mined land for Tichý and Mejstřík 1996; Strock preventing erosion, and reducing the amount of 1998 water and oxygen entering the soil Alkaline addition Mixing of alkaline waste rock with pyritic material Sherlock et al. 1995; Perry et al. or injecting alkaline chemicals to neutralize acidity 1998; Smith and Brady 1998 and precipitate metals Immobilization Capturing the contaminants within the soil mass by Jang et al. 1998 using physico-chemical additives and cementation Rock phosphate Complexation of Fe2+ with rock phosphate in order Evangelou and Zhang 1995 addition to reduce Fe3+ formation

Microencapsulation H2O2-induced iron phosphate coating of pyrite Evangelou and Zhang 1995 surfaces Sacrificial anodes Scrap iron as sacrificial anode to suppress the Shelp et al. 1995 oxidation of metal sulfide ores Microbial inhibitors Application of anionic surfactants, organic acids or Dugan 1975; Evangelou and food preservatives that inhibit the activity of iron Zhang 1995; Kleinmann 1998; and sulfur oxidizing microbes Jozsa et al. 1999; Seidel et al. 2000 Grazing Application of acidophilic protozoa that graze on Johnson 1995 iron oxidizing bacteria

5.1 Isolation of pyritic materials

Most source control approaches aim to limit AMD formation by isolating potentially acid- forming materials from exposure to either oxygen or water (Johnson 2000). Flooding of underground mines is appropriate if they can be securely sealed and detailed knowledge of the mine shafts is available (Johnson 2000). Oxygen diffusion is greatly reduced after flooding because the diffusion coefficient of O2 through the covering water table is only a fraction of that through the atmosphere (Evangelou and Zhang 1995). Thus, the rate of O2 consuming reactions is decreased in flooded sites (Evangelou and Zhang 1995). Underwater storage of mine tailings may be effective, particularly when the tailings are covered with a layer of sediment to further limit O2 penetration and reduce surface perturbations (Johnson 2000). However, complete inhibition of pyrite oxidation by flooding may never be possible because of the availability of Fe3+ as an alternative oxidant (Evangelou and Zhang 1995). Another

14 challenge is to maintain complete and continuous water saturation despite of the fluctuations in the water table (Evangelou and Zhang 1995). Land-based waste heaps may be sealed with covers of clay, plastic liners or organic materials, both to reduce oxygen access and in case of plastic liners or clay, to limit water percolation (Johnson 2000). The feasibility of using isolation approaches depends on the size and nature of the closed mine or waste heap in question (Johnson 1995).

Replacing topsoil and establishing vegetation following mining is a very important reclamation practice in surface mine sites (Strock 1998). Vegetation stabilizes the soil surface against erosion and reduces the amount of water and oxygen entering the soil. An increase in vegetative cover increases the rate of evapotranspiration, ultimately decreasing the volume of mine drainage generated. Oxygen consumption by plants, soil biota and organic material decay provide potential means of limiting oxygen availability for pyrite oxidation (Strock 1998). Plant cover may also be applied to compacted mine waste banks as shown in Figures 9 and 10. Compacting of the waste not only reduces the amount of water and air entering the soil, but at coal mines also prevents spontaneous combustion.

Figure 9. Compacting coal mine waste for land reclamation at a South-African mine (Photo: Anna Kaksonen).

Figure 10. Reclamation and revegetation of coal mine waste at a South-African mine (Photo: Anna Kaksonen).

15 5.2 Alkaline addition

Acid mine drainage is the net result of acid-generating pyrite oxidation and dissolution of neutralizing minerals (Sherlock et al. 1995). Mixing of potentially acid-generating and - consuming materials has been used to produce environmentally benign composites (Johnson 2000). The most prolific mineral sources of alkalinity are carbonate minerals, calcite and dolomite. Other neutralizing minerals include oxides and hydroxides of calcium, magnesium, and aluminium, soluble, non-resistant silicate minerals, and phosphates, primarily apatite (Sherlock et al. 1995). Limestone has often been applied in active mines where AMD is a problem. Alkalinity derived from limestone acts as a pH buffer and AMD neutralizer. It also hydrolyzes most heavy metals, thus precipitating them as metal hydroxides (Evangelou and Zhang 1995). A more recent approach in controlling AMD production is the use of an alkaline recharge (e.g. NaOH, CaCO3). Highly water soluble neutralizers, such as NaOH can be easily moved with percolating water deep in the strata to sites where acid drainage is produced (Evangelou and Zhang 1995). However, the neutralizing effect lasts only as long as the chemicals are supplied (Evangelou and Zhang 1995).

5.3 Immobilization

Immobilization techniques involve the use of physico-chemical additives and cementation to capture metals within the contaminated soil mass (Jang et al. 1998). The immobilization can be attributed to sorption and ion exchange, precipitation, and inclusion of heavy metals into the reaction products. Jang et al. (1998) evaluated the potential of ten chemical additives (bentonite, kaolinite, sand, activated carbon, quicklime, calcium hydroxide, soluble starch xantane, ferrous sulfate, sodium sulfide, and aluminium oxide) to immobilize Cu, Ni and Pb in mine-impacted soils. Quicklime and sodium sulfide showed the best immobilization efficiency and capacity on tested metals. The immobilization was further enhanced when benonite was used in combination with quicklime and sodium sulfide (Jang et al. 1998).

5.4 Phosphate application

Rock phosphate can be used to complex Fe2+, and thus reduce the formation of Fe3+, which enhances pyrite oxidation (Evangelou and Zhang 1995). Rock phosphate does not coat pyrite, but complexes released Fe2+ from the oxidizing pyrite. Instead a coating with Fe2+ forms on the rock phosphate surface reducing rock phosphate dissolution. Therefore, the effectiveness of rock phosphate in controlling pyrite oxidation is short lived (Evangelou and Zhang 1995). A more recent technique is to form highly insoluble iron phosphate (FePO4) coating on pyrite surfaces. This microencapsulation involves leaching pyrite with a solution composed of low but critical concentrations of H2O2, KH2PO4, and pH buffer. H2O2 oxidizes pyrite and produces Fe3+ so that iron phosphate precipitates as a coating on pyrite surfaces. The purpose of the pH buffer is to eliminate the inhibitory effect of the protons on the precipitation of iron phosphate. Successful application of microencapsulation could mean a long-term solution to AMD formation (Evangelou and Zhang 1995).

5.5 Sacrificial anodes

Under natural conditions, the metals of an ore deposit may be oxidized selectively depending on their electrochemical properties. The mineral that is extensively oxidized is the most

16 hydrophobic, the lowest of an electrochemical series, or behaving as the anode of a galvanic cell (Harahuc et al. 2000). Shelp et al. (1995) showed at laboratory scale that the use of scrap iron as sacrificial anodes can suppress the oxidation of metal sulfide ores. The cathode of the electrochemical cell was a block of massive sulfide-graphite rock, and acid mine leachate (pH 3.0) was used as the electrolyte (Shelp et al. 1995). The cell was capable of generating sufficient voltage and current to raise and maintain the pH of the acid leachate at 5.6 and to significantly reduce the redox potential, thus inhibiting the oxidation of sulfide minerals (Shelp et al. 1995). Furthermore, iron sulfate precipitate formed, with a concomitant lowering of Al, Cd, Co, Cu and Ni solution concentrations (Shelp et al. 1995).

5.6 Microbial control

Anionic surfactants, organic acids and food preservatives have been shown to inhibit microorganisms that catalyze the oxidation of sulfides (Dugan 1975; Evangelou and Zhang 1995; Kleinmann 1998; Jozsa et al. 1999; Seidel et al. 2000). The alteration of the semipermeable properties of the cytoplasmic membrane, the following seepage of H+ into the cell and the following decrease in the activity of pH-sensitive enzymes is the most typical mode of inhibition (Evangelou and Zhang 1995; Kleinmann 1998). Anionic surfactants are commercially available (Evangelou and Zhang 1995) and their use has been shown to be cost effective when applied directly to highly pyritic material (Kleinmann 1998). However, the effectiveness of controlling AMD by applying microbial inhibitors has been reported to be highly variable and transient (Jozsa et al. 1999; Seidel et al. 2000; Johnson 2000). Anionic surfactants are very soluble and move with water, thus regular application of chemicals is required to maintain control (Johnson 1995). The surfactants may also biodegrade (Seidel et al. 2000) or adsorb on the mineral surfaces, which reduces their availability (Evangelou ja Zhang 1995; Seidel et al. 2000). Low or irregular dosage of surfactants may enhance pyrite oxidation (Johnson 1995). On the other hand, an excessive surfactant addition may results in water pollution by detergents themselves (Johnson 1995). Periodic or continuous application of surfactant solution can be appropriate at active mine waste piles where new pyritic material is always being added, but is not worthwhile after the pile is completed and revegetated (Kleinmann 1998). The surfactant may be adsorbed by the soil and never reach the pyritic material. If pyrite oxidation is allowed to proceed beneath the soil cover, the roots of the vegetation will be exposed to acidic water and may wither away. To counteract this problem, slow-release surfactant pellets that could be applied before the soil cover were developed. The release life-time is intended to allow time for vegetation and normal soil biota to become established in the topsoil layer, consuming the oxygen that would otherwise cause pyrite oxidation (Kleinmann 1998).

Another microbial control approach has been to use acidophilic protozoa that prey on iron- oxidizing microorganisms. Grazing of leaching bacteria by a variety of acidophilic protozoa has been shown to reduce numbers of metal-mobilizing bacteria oxidizing pyrite-rich coal in laboratory cultures (Johnson and Rang 1993). Moreover, the appearance of protozoa in a pilot-scale reactor has been found to coincide with a severe decline in bacterial numbers in the leaching liquor (Johnson 1995). However, because of the predator-prey relationship between protozoa and bacteria, the greatest level of control achievable by protozoan grazing would probably be to sustain lesser and fluctuating numbers of bacteria rather than eradicating them completely (Johnson 1995).

17 6 WASTEWATER TREATMENT

6.1 Abiotic processes

Numerous abiotic techniques are available for the treatment of acidic metal-containing wastewaters formed in mines and metallurgical operations (For reviews, see Lanouette 1977; Peters et al. 1985). The techniques include neutralization with alkaline materials, chemical metal precipitation, and various alternative physico-chemical processes. Most of the abiotic techniques aim at the removal of metals or acidity leaving sulfate in the solution (Tichý et al. 1998).

6.1.1 Passive limestone neutralization

Limestone (CaCO3) is a relatively inexpensive alkalinity source for passive neutralization of acidic mine water (Hedin et al. 1994; Evangelou and Zhang 1995). Neutralization can be achieved by passage of water through engineered systems, such as open limestone channels (OLC) or anoxic limestone drains (ALD) (Hedin et al. 1994; Robb and Robinson 1995; Evangelou and Zhang 1995; Ziemkiewicz et al. 1997). The mine water reacts with the surfaces of the limestone to dissolve carbonate which consumes hydrogen ions and adds alkalinity to the mine water (Robb and Robinson 1995). Limestone has limited solubility and a tendency to be coated with ferric hydroxide (Fe(OH)3) precipitates (Evangelou and Zhang 1995). When mine waters enriched with iron contact limestone in an oxidizing environment, such as OLC, the limestone is rapidly coated with ferric hydroxide precipitates (Evangelou and Zhang 1995; Robb and Robinson 1995). The dissolution of the coated limestone is inhibited, and the production of alkalinity is significantly diminished (Evangelou and Zhang 1995). Moreover, the formation of gelatinous aluminum hydroxide (Al(OH)3) can plug the flow paths (Gazea et al. 1996). In anoxic environments, limestone dissolution and subsequent production of alkalinity can proceed without inhibitory coating (Evangelou ja Zhang 1995). In ALDs, water containing little or no dissolved oxygen, Fe3+ and Al3+ is fed through a drain consisting of crushed limestone which is buried in the ground and sealed from air (Robb and Robinson 1995; Gazea et al. 1996; Johnson 2000). ALDs have often been used as a pre- treatment in combination with other treatment methods, such as wetlands (Hedin et al. 1994; Robb and Robinson 1995). Hedin et al. (1994) evaluated the performance of two ALDs receiving mine water contaminated with ferrous Fe and Mn. By significantly increasing alkalinity concentrations in the mine waters, both ALDs increased metal removal in downstream constructed wetland. While design parameters can be manipulated to effect the theoretical lifetime of ALDs, the alkalinity generation of an ALD is strongly influenced by chemical characteristics of the influent mine water (Hedin et al. 1994).

6.1.2 Chemical precipitation

The most widely used active process for heavy metal removal is chemical precipitation (Table 5) (Lanouette 1977; Peters et al. 1985; Veeken and Rulkens 2003). Hydroxide precipitation is the most commonly used precipitation technique due to its relative simplicity and ease of automatic pH-control (Peters et al. 1985; Veeken and Rulkens 2003). The chemicals used for hydroxide precipitation include quicklime (CaO), hydrated lime (Ca(OH)2), caustic soda (NaOH) and NH4OH (Lanouette 1977; Peters et al. 1985). Part of the sulfate is removed as gypsum (CaSO4·2H2O) when calcium-containing neutralizing chemicals are used (Johnson 2000). Most metal hydroxides have minimum solubility in the range of pH 7.5-11 (Conner 1990). However, some metals such as chromium which exhibit amphoteric behavior have

18 higher solubility at both low and high pH (Conner 1990). The optimum pH for hydroxide precipitation is different for each metal and, often, for different valence states of a single metal. It may also vary for a specific metal ion with the presence of other species in solution, with the redox potential, and with aging of the hydroxide (Conner 1990). Neutralization is often combined with aeration to allow the oxidation of metals such as iron and manganese (Evangelou and Zhang 1995). The treatment process may also include other unit operations, such as coagulant addition, flocculation, sludge settling and recycling part of the settled sludge (Figure 11) (Lanouette 1977). The sludge can be dewatered in centrifuges, vacuum or plate-and-frame filters or on sludge-drying beds (Lanouette 1977). Hydroxide precipitation of metals produces sludge that is usually gelatinous, thereby increasing the difficulty of dewatering (Lanouette 1977). Lime will produce considerably greater quantity of sludge than sodium hydroxide but is usually easier to dewater (Lanouette 1977).

Table 5. Techniques for the treatment of acid mine drainage by chemical precipitation. Technique Principle Reference(s) Hydroxide Precipitation of metals as hydroxides and Lanouette 1977; Peters et al. 1985; precipitation neutralization of acidity using NaOH, NH4OH, Robb and Robinson 1995; Banks et al. CaO or Ca(OH)2 (Alkaline dosing, aeration, 1997 flocculation and settling) Carbonate Precipitation of metals as carbonates using Peters et al. 1985 precipitation CaCO3, NaHCO3 or Na2CO3 Phosphate Precipitation of metals as phosphates Conner 1990 precipitation Sulfide Precipitation of metals as sulfides using soluble Lanouette 1977; Whang 1982; Peters et precipitation (Na2S, NaHS, CaS) or gaseous (H2S) sulfide al. 1985; Banks et al. 1997; Alhodali et al. 2002; Veeken and Rulkens 2003 Ferrous Dissolved metals react with added ferrous sulfide Lanouette 1977; Peters et al. 1985; sulfide resulting in replacement of the FeS by the Davis et al. 1994 precipitation appropriate metal sulfide Ferrite Controlled formation of ferrite from ferrous Morgan et al. 2003 process solutions using magnetite seed

pH- Neutralizing sensor chemicals Mixing

Acid Mixing Polymer mine drainage

Aeration and neutralization Treated Flocculation Sedimentation effluent

Recycled sludge

Sludge treatment

Figure 11. A flow chart of a heavy metal precipitation system using alkaline chemicals (adapted from Lanouette 1977).

19

With selected metals, carbonates are less soluble than their corresponding hydroxides (Table 6) (Conner 1990). Limestone (CaCO3) is the lowest cost alkali, but its reaction rate with sulfuric acid is slow due to the coating of limestone surfaces (van Houten and Lettinga 1995). Another often used chemical is soda ash (Na2CO3) (Lanouette 1977). A drawback in carbonate precipitation is that carbonates are solubilized at low pH. If CO2 is evolved, the reaction is irreversible and the final speciation of the metal will be as the hydroxide (Conner 1990).

Table 6. Solubility products of metals with hydroxide, carbonate, phosphate and sulfide ions at temperatures between 18 and 25 ºC (Dean 1999). Solubility products (mol l-1) Metal - 2- 3- 2- OH CO3 PO4 S Al3+ 1.3 x 10-33 9.84 x 10-21 2 x 10-7 Ag+ 2.0 x 10-8 8.46 x 10-12 8.89 x 10-17 6.3 x 10-50 As3+ 2.1 x 10-22 Ca2+ 5.5 x 10-6 2.8 x 10-9 2.07 x 10-29 Cd2+ 7.2 x 10-15 1.0 x 10-12 2.53 x 10-33 8.0 x 10-27 Co2+ 5.92 x 10-15 1.4 x 10-13 2.05 x 10-35 α-CoS 4.0 x 10-21 β-CoS 2.0 x 10-25 Cr2+ 2 x 10-16 3+ -31 -23 Cr 6.3 x 10 CrPO4·4H2O green 2.4 x 10 -17 CrPO4·4H2O violet 1.0 x 10 Cu2+ 2.2 x 10-20 1.4 x 10-10 1.40 x 10-37 6.3 x 10-36 Fe2+ 4.87 x 10-17 3.13 x 10-11 6.3 x 10-18 3+ -39 -16 Fe 2.79 x 10 FePO4·2H2O 9.91 x 10 Hg2+ 2.0 x 10-24 3.6 x 10-17 1.0 x 10-47 Mg2+ 5.61 x 10-12 6.82 x 10-6 1.04 x 10-24 Mn2+ 1.9 x 10-13 2.34 x 10-11 MnS amorphous 2.5 x 10-10 MnS crystalline 2.5 x 10-13 Ni2+ 5.48 x 10-16 1.42 x 10-7 4.74 x 10-32 α-NiS 3.2 x 10-19 β-NiS 1.0 x 10-24 γ-NiS 2.0 x 10-26 Pb2+ 1.43 x 10-15 7.4 x 10-14 8.0 x 10-43 8.0 x 10-28 Sn2+ 5.45 x 10-28 1.0 x 10-25 Zn2+ 3 x 10-17 1.46 x 10-10 9 x 10-33 α-ZnS 1.6 x 10-24 β-ZnS 2.5 x 10-22

The simple phosphate salts of toxic metals (Mn+2PnO3n+1, where n = 1) have low water solubility (Table 6), although they are soluble in acids (Conner 1990). However, phosphates with n > 1 have the potential to bind metals as water-soluble species (Conner 1990). Therefore, the presence of phosphates may be harmful or beneficial for metal removal, depending on the phosphate species (Conner 1990).

Sulfide precipitation has been demonstrated to have several benefits over the hydroxide precipitation (Peters et al. 1985). Attractive features of this process include: a) most divalent metal sulfides are less soluble than the corresponding hydroxides (Table 6) and their solubility is not as sensitive to changes in pH (Conner 1990), b) effluent concentrations are orders of magnitude lower (µg l-1 versus mg l-1), c) in interference of chelating agents in the wastewater is less problematic, d) possibility for selective metal precipitation, e) high reaction rates allow low hydraulic retention times, f) metal sulfides exhibit better thickening and dewatering characteristics than corresponding hydroxide sludge and g) sulfide precipitates can be processed by existing smelters for metal recovery (Whang 1982; Peters et al. 1985; Veeken and Rulkens 2003). Chemical sulfide precipitation has not been widely used for AMD

20 treatment, probably due to the high cost of chemicals (Lanouette 1977; van Houten and Lettinga 1995). One drawback of metal sulfides is that they can resolubilize in an acidic, oxidizing environment (Conner 1990). If the amount of sulfide reagent added to a solution differs from the stoichiometric amount required to precipitate metals, either excess metals or H2S will be present. The odor and toxicity of H2S may require process tanks to be closed and vacuum-evacuated (Lanouette 1977; Hammack et al. 1994b). Additionally excess sulfide requires removal before releasing the water into the environment (Lanouette 1977).

A process utilizing ferrous sulfide (FeS) as the principle source of sulfide appears to overcome the problem of generating excess H2S (Lanouette 1977; Peters et al. 1985; Davis et al. 1994). The sulfide is released from FeS only when other heavy metals with lower equilibrium constants for their sulfide form are present in solution. If pH can be maintained at 8.5-9, the liberated iron will form a hydroxide and precipitate out as well (Lanouette 1977). Disadvantages of the process include considerably larger than stoichiometric reagent consumption and large quantities of sludge due to the ferrous hydroxide formation (Peters et al. 1985).

Ferrite process is based on the controlled formation of partially oxidized iron oxide, magnetite 3+ 2+ Fe3O4 (i.e. Fe 2Fe O4), that has the capacity to substitute divalent and trivalent cations as 3+ 2+ part of the lattice to form ferrites (M1 2M2 O4, where M1 and M2 stand for any of a number of possible metals). The process involves a) the precipitation of hydroxyl-metals at pH 10.5 and their subsequent adsorption onto magnetite seed in a contact stabilization reactor, b) liquid-solid separation and treatment of solid fraction in an oxidizing reactor in which a fraction of ferrous species is oxidized to an intermediate ferric precipitate and c) incorporation of both ferrous and ferric species into stable magnetite. Crystallized end-products settle and dewater more efficiently than does the amorphous ferrihydrite (Fe(OH)3) formed in conventional hydroxide precipitation (Morgan et al. 2003). Additionally the magnetic properties of ferrite sludge can be used for sludge separation by applying a magnetic field (Peters et al. 1985; Barrado et al. 1999)

6.1.3 Alternative physico-chemical treatment techniques

Other active physico-chemical techniques which have been used for removal of metals from aqueous solutions include cementation, coagulation and flocculation, foam flotation, membrane filtration, electrolysis, electrodialysis, ion exchange, solvent extraction, adsorption and evaporation/distillation (Table 7). Performance characteristics of major abiotic heavy metal removal/recovery techniques are shown in Table 8.

Cementation is a metal replacement process that occurs when a solution containing a dissolved metallic ion comes into contact with a more active metal, such as iron. The less active metal is cemented onto scrap iron as the iron replaces it in solution. The less active metal is recoverable as a pure metal and the iron can be precipitated as an iron hydroxide. The process is particularly applicable as a means of removing copper and silver from solutions (Lanouette 1977).

Coagulation refers to the charge neutralization of particles. Flocculation involves slow mixing to promote the agglomeration of the destabilized particles. Foam flotation depends on the use of a surfactant that causes a nonsurface active material to become surface active. The product is removed by bubbling a gas through the bulk solution to form foam (Peters et al. 1985).

21 In recent years, there has been an increasing interest in the application of membranes for separation processes (Kozlowski and Walkowiak 2002). Membrane processes include high pressure reverse osmosis (500-1500 psi), low pressure reverse osmosis (200-500 psi) and ultrafiltration (20-100 psi) (Peters et al. 1985). In the processes, the wastewater flows under pressure through a porous membrane and the clear permeate product is withdrawn at atmospheric pressure (Peters et al. 1985). In ultrafiltration, the different chemical components are separated exclusively by molecular weight, whereas in reverse osmosis, both particle size and the chemical nature are important for the separation (Peters et al. 1985). Membrane processes are essentially concentration techniques which can be used for the recovery of metals (Lanouette 1977). Major limitations of membrane processes include membrane fouling and limited lifetime (Peters et al. 1985).

In electrolytic recovery of metals, a direct current is passed through an aqueous solution containing metal ions between cathode plates and insoluble anodes. The positively charged metallic ions adhere to the negatively charged cathodes leaving a metal deposit that can be stripped off and recovered (Peters et al. 1985). Electrodialysis involves the transport of ionic species through membranes by application of an electric potential (Peters et al. 1985). The membrane contains cation or anion exchange resins, selective for only cations or anions in the wastewater (Peters et al. 1985; Chartrand and Bunce 2003).

Table 7. Alternative physico-chemical techniques for the removal/recovery of metals from water. Technique Principle Reference(s) Cementation Metal replacement process in which an ionized Lanouette 1977; Peters et al. 1985 metal is recovered from solution by reduction to the elemental metallic state with subsequent oxidation of a sacrificial metal, such as scrap iron Coagulation- Coagulation: charge neutralization of particles Peters et al. 1985 flocculation using e.g. ferric sulfate, alum or sodium aluminate. Flocculation: slow mixing to promote the agglomeration of destabilized particles Foam Use of gas bubbles and surfactants to decrease the Peters et al. 1985; Rubio et al. 1996 flotation apparent density of aggregates which float to the liquid/air interfase Membrane Filtration through semipermeable membrane Lanouette 1977; Peters et al. 1985; filtration (reverse osmosis, micro/nano/ultrafiltration) Rubio et al. 1996; Banks et al. 1997; Kozlowski and Walkowiak 2002 Electrolytic Deposition of positively charged metal ions on Peters et al. 1985 recovery of negatively charged cathodes by applying a direct metals current through the metal containing aqueous solution (electrowinning) Electro- Use of electric potential to transport ionic species Peters et al. 1985; Chartrand and dialysis through cation or anion exchange membranes Bunce 2003 Ion exchange Exchange of ions on the resin for those in Lanouette 1977; Peters et al. 1985; wastewater Rubio et al. 1996; Chmielewski et al. 1997; Bosso and Enzweiler 2002 Liquid-liquid Extraction of metals using an organic solvent that Peters et al. 1985; Rubio et al. 1996; extraction flows countercurrent to the wastewater Chmielewski et al. 1997 Membrane Extraction of metals through membranes which Peters et al. 1985 extraction separate aqueous and organic solvent phases Adsorption Adsorption of metals on e.g. activated carbon Lanouette 1977; Peters et al. 1985; Rubio et al. 1996 Evaporation/ Recovery of metals by evaporation/distillation Peters et al. 1985 distillation

22 Table 8. Performance characteristics of major abiotic heavy metal removal/recovery techniques (adapted from Peters et al. 1985; Eccles 1995) Performance characteristics Working level for Technique Metal Influence of Tolerance of organic pH change appropriate metal End product selectivity suspended solids compounds (mg l-1) Hydroxide Limited tolerance Non-selective Tolerant Tolerant, complexing >10 Gelatinous sludge precipitation agents may have adverse effect Sulfide Limited tolerance Limited Tolerant Tolerant >10 Dense sludge precipitation selective, pH- dependent Electrochemical Tolerant Moderate Can be Can be >10 Metal deposit operations engineered to accommodated tolerate Ion exchange Limited tolerance Chelate-resins Fouled Can be poisoned <100 Concentrated liquid can be selective Membrane Limited tolerance Moderate Fouled Intolerant >10 Concentrated liquid processes Adsorption Limited tolerance Moderate Fouled Can be poisoned <10 Concentrated liquid or metal-containing adsorbent Liquid-liquid Limited tolerance Moderate Tolerant Tolerant >10 Concentrated liquid extraction Foam flotation Limited tolerance Non-selective Tolerant Tolerant >10 Dense sludge

23 Ion exchange effectively removes low concentrations of heavy metals from wastewater streams (Peters et al. 1985). The technique is based on a reversible chemical reaction, where the removal of heavy metals is accomplished by the exchange of ions on the resin for those in wastewater (Peters et al. 1985). When ion exchange resins are saturated, they must be regenerated with an acid or alkaline medium to remove metal ions from the resin bed. The regenerant solution is smaller in volume and higher in concentration than the wastewater (Lanouette 1977). Because of the regeneration step, the process is semicontinuous unless several ion exchange columns are used simultaneously allowing the regeneration of one column as the others are in use (Scheeren et al. 1991).

Liquid-liquid extraction involves a two phase system where an organic solvent is usually run countercurrent to an aqueous metal-containing wastewater. Acid-treating the solvent releases the metal in a concentrated water-soluble form (Peters et al. 1985). In membrane extraction, metals are removed from wastewater through membranes which separate aqueous and organic phases. The metals are removed from the organic solvent in a stripping module, and the regenerated solvent is recycled back to the extraction module (Peters et al. 1985). Water droplets in the organic phase can significantly deteriorate the performance of the system (Peters et al. 1985).

Activated carbon adsorbs hexavalent chromium, mercury and many metal compounds that have been complexed in organic forms (Lanouette 1977). The adsorption capacity of carbon depends on the pore size, solution pH and size and concentration of the target molecule. The adsorption capacity generally increases as the solution pH decreases. Granulated carbon is more expensive than powdered carbon, but it can be chemically regenerated and reused (Lanouette 1977).

6.2 Biological processes

Several biological processes can remove metals from wastewaters (Table 9) (for reviews, see Gadd 1992; White et al. 1995). Also a number of biological processes can generate alkalinity or consume acidity and which, therefore, have potential use in neutralizing AMD (Johnson 2000). These include photosynthesis (Robb and Robinson 1995; van Hille et al. 1999; Johnson 2000), denitrification (Kalin et al. 1991; Johnson 1995), ammonification, methanogenesis, and reduction of iron and sulfate (Kalin et al. 1991; Johnson 1995; White et al. 1997; Johnson 2000). Of these, iron (Reaction 9) and sulfate reduction (Reactions 10-11) are the obvious candidates for AMD bioremediation (Johnson 2000).

2+ - 4 Fe(OH)3 + CH2O Æ 4 Fe + H2CO3 + 2 H2O + 8 OH (9)

2- - SO4 + 2 CH2O Æ H2S + 2 HCO3 (10)

- + HCO3 + H Æ CO2 (g) + H2O (11)

Both processes consume electrons (CH2O is an electron donor in reactions 9 and 10) (Johnson 2+ 2000). Iron reduction results in the mobilization of soluble Fe , whereas the H2S produced in the sulfate reduction precipitates metals as low solubility sulfides (Reaction 12) (Dvorak et al. 1992).

2+ + 2+ 2+ H2S + M Æ MS(s) + 2 H where M = metal, such as Zn (12)

24 The sulfide precipitation of metal cations releases protons. Therefore, excess of sulfate reduction is needed compared to metal precipitation to produce net alkalinity. Due to the combined removal of acidity, metals and sulfate, sulfate-reduction appears to be the most promising process for AMD treatment. The potential utility of microbial sulfate reduction in mining applications was recognized in the late 1960s (Tuttle et al. 1969). Since then, the activity of sulfate-reducing microorganisms has been employed in various passive treatment systems and active bioreactors.

Table 9. Biological mechanisms for removal of metals from wastewaters. Mechanism Organisms/materials Reference(s) Biosorption Plants, algae, bacteria, fungi, Peters et al. 1985; Brierley et al. 1989; Rossi 1990; yeast, peat, agricultural Gadd 1992; Eccles 1995; Tsezos et al. 1995; Rubio et byproducts, biopolymers al. 1996; Scott and Karanjkar 1998; van Hille et al. 1999; Brown et al. 2000; Chen et al. 2000; Utgikar et al. 2000; Gazsó 2001; Mallick 2002; Ibeanusi et al. 2003 Intracellular uptake Plants, algae, bacteria, yeasts Brierley et al. 1989; Rossi 1990; Gadd 1992; Ghosh and accumulation and Bupp 1992; Brady et al. 1994; Eccles 1995; Evangelou and Zhang 1995; Scott and Karanjkar 1998; Gazsó 2001; Ibeanusi et al. 2003 Complexation Microorganisms, algae, peat, Brierley et al. 1989; Gadd 1992; Gazea et al. 1996; van immobilized siderophores Hille et al. 1999; Brown et al. 2000; Gazsó 2001; Ibeanusi et al. 2003 Oxidation Fe2+ and Mn2+ oxidizing Jianwei 1995; Gazea et al. 1996; Umita 1996; Diz and microorganisms Novak 1998; Gazsó 2001 Reduction Fe3+, Hg2+, Cr6+, Mn4+, U6+, Gadd 1992; Kashefi and Lovley 2000; Gazsó 2001; Tc7+ and Co3+ reducing Chardin et al. 2002 microorganisms Methylation and Hg2+, Cd2+, Pb2+, As, Se and Brierley et al. 1989; Rossi 1990; Gadd 1992; Gazsó volatilization Te methylating 2001 microorganisms, fungi Extracellular Phosphate liberating Rossi 1990; Macaskie et al. 1995; Roig et al. 1997; precipitation microorganisms, sulfate- and Thomas and Macaskie 1998; Ibeanusi and Wilde 1998; sulfur-reducers, pH Robb and Robinson 1995; White et al. 1997; Podda et increasing microorganisms al. 2000; Gazsó 2001; Boonstra et al. 2002

6.2.1 Passive treatment methods

Passive biological methods for AMD treatment include enhancement of microbial activity in groundwater aquifers through substrate injection (Groudev et al. 1998) or reactive barriers (Figure 12) (Herbert Jr. et al. 1998; Waybrant et al. 1998; Benner et al. 1999; Amos and Younger 2003), anoxic ponds, natural and constructed wetland systems, and aerobic rock filters (Figure 13) (for reviews, see Robb and Robinson 1995; Gazea et al. 1996).

AMD-impacted groundwater has been remediated in situ by enhancing the activity of sulfate- reducing bacteria (SRB) by placing or injecting substrates to the subsurface through boreholes (Groudev et al. 1998; Canty 1999) or by constructing permeable reactive barriers across the groundwater flow path (Herbert Jr. et al. 1998; Waybrant et al. 1998; Benner et al. 1999; Amos and Younger 2003). Groudev et al. (1998) reported an increase in numbers and activity of SRB in the contaminated groundwater after injecting an acetate-bearing waste product and ammonium phosphate to the subsurface near the Burgas Copper Mines, Bulgaria. The 6+ 4+ biogenic H2S reduced U to U and precipitated heavy metals (Cu, Zn Pb, Mo and Mn) as sulfides (Groudev et al. 1998). Also Canty (1999) reported high removal efficiencies for metals (Al, Cd, Co and Zn) and increased pH in mine water flowing through organic substrate

25 placed in mine shafts. However, higher flow rates during the spring and oxygenated surface water resulted in pH decrease and resolubilization of the metal precipitates (Canty 1999).

Permeable reactive barriers consist of zones of reactive material installed across the flow path of the plume of contaminated groundwater (Figure 12) (Richardson and Nicklow 2002). As the AMD-impacted groundwater flows through this zone, SRB reduce sulfate from the water by using the electron sources present in the barrier. This results in the generation of bicarbonate alkalinity and precipitation of metals as sulfides. Organic substrates tested in reactive barrier mixtures include municipal and leaf compost, sewage sludge, manure, cellulose, sawdust and wood chips (Herbert Jr. et al. 1998; Waybrant et. al. 1998; Benner et al. 1999). Zero valent iron (Fe0) can be used to stimulate sulfate-reduction, since Fe0 corrosion depletes oxygen and produces cathodic H2, a potential energy source for SRB (Gu et al. 2002). The selection of the reactive mixture significantly affects the permeability and reactivity of the barrier (Waybrant et. al. 1998; Amos and Younger 2003). Pea gravel can be mixed with organic substrate to increase the permeability of the barrier, and limestone may be added to provide additional alkalinity (Herbert Jr. et al. 1998; Waybrant et. al. 1998; Amos and Younger 2003). Reactive barriers typically have been employed in continuous barriers or in a funnel and gate configuration, as illustrated in Figures 12a and 12b, respectively (Richardson and Nicklow 2002). A prerequisite for successful in situ remediation is a thorough hydrogeological characterization of the site. Reactive barriers rely on natural groundwater flow to move contaminants through the treatment zone, which results in long treatment times. Depletion of organic substrates and clogging of the barrier due to metal precipitation may deteriorate the long-term effectiveness of the system (Richardson and Nicklow 2002).

Reactive gate abContinuous reactive barrier Funnel

Groundwater Contaminant Groundwater Groundwater Contaminant flow plume level flow plume

Figure 12. Contaminated groundwater plume capture by a) a continuous reactive barrier, b) funnel and gate system (adapted from Richardson and Nicklow 2002).

Anoxic ponds are water basins supplemented with organic substrate (Figure 13a). They can be used e.g. upstream of ALDs (Figure 13b) aiming at decreasing dissolved oxygen, reducing Fe3+ to Fe2+ and removing Al3+ (Figure 13f) (Gazea et al. 1996). Riekkola-Vanhanen and Mustikkamäki (1997) used a flooded open pit (Ruostesuo open pit near Pyhäsalmi mine, Finland), as a large-scale basin to treat AMD with SRB. Press-juice from silage and liquid manure were added as sources of electrons and SRB, respectively. Sulfate reduction activity was observed as a slow increase in water pH and a decrease in sulfate, zinc, iron and manganese concentrations and redox-potential (Riekkola-Vanhanen and Mustikkamäki 1997).

Wetlands have been recognized for several years as low-cost systems to improve the water quality of AMD (Hunstman et al. 1978; Wieder and Lang 1982). During the last two decades, constructed wetland systems have been developed from an experimental concept to full-scale

26 field application (Gazea et al. 1996) and used for removal of sulfate, metals and radionuclides from mine waters (Noller et al. 1994). The approach has become particularly popular in the USA, where hundreds of wetland sites have been operating for the treatment of wastewaters from coal mine areas in Appalachia alone (Gazea et al. 1996; Johnson 2000). Wetlands are highly complex ecosystems, where water quality is affected by a number of physical, chemical and biological processes including dilution, filtering of suspended particles, adsorption, complexation, ion exchange and uptake of metals, and precipitation by oxidative and reductive mechanisms (Eger et al. 1994; Evangelou and Zhang 1995; Gazea et al. 1996; Johnson 2000).

Constructed aerobic wetlands are shallow, vegetated and operate by causing the incoming water to flow over the wetland surface (Figure 13c) (Robb and Robinson 1995; Johnson and Hallberg 2002). The major objective is to enhance the oxidation and hydrolysis reactions of iron and other metals, and to retain the resulting metal precipitates by entrapment (Barton and Karathanasis 1999; Johnson and Hallberg 2002). The hydrolysis of metals produces acidity and, therefore, aerobic wetlands are applicable to the treatment of net alkaline waters (Robb and Robinson 1995; Gazea et al. 1996). Macrophytes are planted for aesthetic reasons, as well as to regulate water flow and stabilize the accumulating precipitates against erosion (Gazea et al. 1996; Johnson and Hallberg 2002). In long term, periodical solids removal may be needed, since the accumulation of metal precipitates will likely decrease the retention time in the wetland (Hedin 1996).

Wetlands supplemented with submerged organic substrate and limestone are called anaerobic wetlands or compost wetlands (Figure 13d). Materials that have been used as substrates include spent mushroom compost, manure, hay bales, peat, wood chips and saw dust (Eger 1994; Gazea et al. 1996; Johnson 2000). Anaerobic wetlands generate alkalinity through a combination of microbial activity and limestone dissolution (Gazea et al. 1996; Hedin 1996). The reduction of both iron and sulfate are considered important in anaerobic wetlands (Johnson 2000). Other biological activities may also contribute to the neutralization of AMD, including methanogenesis and ammonification (Johnson and Hallberg 2002). The biogenic hydrogen sulfide precipitates metals as sulfides (Robb and Robinson 1995; Barton and Karathanasis 1999). Vegetation growing on the submerged substrate can provide a continuous supply of carbon and energy for the underlying microbial community (Gazea et al. 1996; Mitsch and Wise 1998) and protect against wind erosion at periods when the water level drops below that of the substrate (Gazea et al. 1996). However, dense plant growth can also cause problems, such as preferential flow paths and decrease in retention time due to litter accumulation (Hedin 1996) or diffusion of oxygen from the roots into the surrounding substrate (Gazea et al. 1996). Dense vegetation may also attract muskrats, which can damage berms and create swimming channels that cause channelized flow through the wetland cells (Hedin 1996). It is likely that well-vegetated anaerobic wetlands will require periodic maintenance in order to maintain the designed retention times (Hedin 1996). Disposal of metal-laden precipitates from anaerobic wetlands may be more difficult than from aerobic wetlands (Hedin 1996).

27 f) a) Acid mine drainage Flow direction Organic material

Anoxic pond

2+ - 4 Fe(OH)3 + CH2O Æ 4 Fe + H2CO3 + 2 H2O + 8 OH 2- - SO4 + 2 CH2O Æ H2S + 2 HCO3 Gravel Liner HCO - + H+ Æ CO (g) + H O b) 3 2 2 H S + M2+ Æ MS(s) + 2 H+ (M2+ = metal) Flow direction Clay capping 2

Anoxic limestone drain

+ 2+ - CaCO3+H Æ Ca + HCO3 Limestone chippings Liner

c) Aerobic wetland Flow direction Vegetation CO2 + H2O + hv Æ O2 + CH2O 2+ + 3+ 4 Fe + 4 H + O2 Æ 4 Fe + 2 H2O 3+ + Fe + 3 H2O Æ Fe(OH)3 + 3 H

Organic topsoil Berm with Liner substrate spillway Anaerobic wetland 4 Fe(OH) + CH O Æ 4 Fe2+ + H CO + 2 H O + 8 OH- d) 3 2 2 3 2 2- - Organic material Vegetation SO4 + 2 CH2O Æ H2S + 2 HCO3 - + Flow HCO3 + H Æ CO2(g) + H2O direction 2+ + 2+ H2S + M Æ MS(s) + 2 H (M = metal)

Rockfilter CO + H O + hv Æ O + CH O Gravel Liner 2 2 2 2 Mn2+ + HCO - Æ MnCO + H+ e) 3 3

Flow direction Algae 2 MnCO3 + O2 Æ 2 MnO2 + 2 CO2 2+ - Mn + 2 OH Æ Mn(OH)2

Treated effluent Inert rockfill Liner Figure 13. Section through a) an anoxic pond, b) an anoxic limestone drain, c) an aerobic wetland, d) an anaerobic wetland and e) an aerobic rock filter. f) A flow chart of a passive acid mine drainage treatment system consisting of biological (green) and abiotic (black) units. Major abiotic (black) biologically mediated (green) removal mechanisms are shown (adapted from Cambridge 1995; Robb and Robinson 1995; Gazea et al. 1996).

Aerobic and anaerobic wetland units are often used consecutively and in combination with ALDs (Figure 13f) (Fyson et al. 1995; Sikora et al. 1995; Mitsch and Wise 1998; Johnson and Hallberg 2002). Wetland systems remediate AMD with low cost and minimal maintenance (Gazea et al. 1996). However, their treatment efficiency has been variable (Fyson et al. 1995; Kalin et al 1995; Sikora et al. 1995; Barton and Karathanasis 1999; Johnson and Hallberg 2002). The challenges of wetland treatment include the following: a) toxic effects of the AMD on wetland organisms (Roane et al. 1996; Brierley et al. 1989), b) disposal of nonviable

28 or excess biomass containing heavy metals (Roane et al. 1996; Brierley et al. 1989), c) seasonal variations (Brierley et al. 1989; Sikora et al. 1995; Johnson and Hallberg 2002), d) catastrophic system failures that may occur due to insufficient utilization of the treatment area, e) metal overloading and inadequate alkalinity production (Barton and Karathanasis 1999) and f) futile cycling of iron and sulfur (Johnson and Hallberg 2002). Wetland treatment may not be effective in arid and semi-arid climates (Noller et al. 1994; Johnson and Hallberg 2002), and exposure of the metal-sulfide sediments to oxygen in the periods of drought can lead to dissolution of the metals and acidification of the system (Tichý and Mejstřík 1996). Moreover, wetland treatment does not always result in effective removal of manganese (Johnson and Hallberg 2002).

Rock filters can be used to improve manganese removal (Figure 13e). They consist of aerobic cells that contain granite rocks which provide a large surface area for algal growth (Gazea et al. 1996). The algae create microenvironments with an elevated pH. This allows the precipitation of metals, such as manganese, which require high pH for removal (Cambridge 1995; Robb and Robinson 1995). Rock filters are often used downstream of compost wetlands as a final phase in passive treatment (Figure 13f) (Cambridge 1995; Gazea et al. 1996; Johnson and Hallberg 2002).

Passive biological treatment approaches offer low-cost and minimal maintenance solutions for treating AMD. However, the required treatment area may be large, metal recovery difficult, and control and predictability poor due to seasonal variations (Figure 14) (Gazea et al. 1996). The weaknesses of passive treatment have been overcome by enriching useful organisms from the environment and promoting their activity in controlled bioreactors (Johnson 2000).

passive Method active

low Costs high

small Need of labour high

large Treatment area small

difficult Metal recovery easy

poor Control good

poor Predictability good

Figure 14. Characteristics of passive and active biological treatment methods for acid mine drainage treatment (adapted from Gazea et al. 1996; Johnson 2000).

6.2.2 Bioreactor processes

Bioreactors rely on the retention of specific organisms that carry out desired conversions (Lens et al. 2002). Microbes utilized in AMD-treating bioreactors include e.g. sulfate- and sulfur-reducing bacteria (Hao 2000; Boonstra et al. 2002), Acidithiobacillus sp. and other iron oxidizers (Jianwei 1995; Umita 1996; Diz and Novak 1998), Bacillus sp. (Ibeanusi and Wilde 1998), Pseudomonas sp. (Ibeanusi and Wilde 1998; Scott and Karanjkar 1998), Clostridium sp. (Kauffman et al. 1986), Enterobacter sp., Serratia sp. (Scott and Karanjkar 1998),

29 Citrobacter sp. (Roig et al. 1997), Spirulina sp. and other algae (van Hille et al. 1999; Mallick 2002), Saccharomyces sp. yeast (Brady et al. 1994), and various mixed cultures (Ghosh and Bupp 1992). Bioreactors utilizing sulfate-reducing microorganisms provide the advantage of combined removal of acidity, metals and sulfate.

Numerous reactor designs for biological sulfate reduction have been reported (for reviews, see Hao 2000; Johnson 2000; Hulshoff Pol et al. 2001; Lens et al. 2002), such as batch reactors (Christensen et al. 1996), sequencing batch reactors (Herrera et al. 1991; Herrera et al. 1997; Sen and Johnson 1999; Luptakova 2003), continuously stirred tank reactors (Barnes et al. 1991b; White and Gadd 1996a), anaerobic contact processes (Haas and Polprasert 1993), anaerobic baffled reactors (Barber and Stuckey 2000), anaerobic filters (Maree and Strydom 1985), fluidized-bed reactors (Somlev and Tishkov 1992; Ma and Hua 1997; Somlev and Banov 1998), gas lift reactors (van Houten and Lettinga 1995), upflow anaerobic sludge blanket reactors (de Vegt and Buisman 1995; de Lima et al. 1996; de Vegt and Buisman 1996; Hammack and Dijkman 1999; Dijkman et al. 1999) and anaerobic hybrid reactors (Steed et al. 2000) (Figure 15). Benefits and drawbacks of various continuous flow reactors are listed in Table 10 (for a review, see Speece 1983).

The reactor configuration has implications for the ratio of sludge retention time/hydraulic retention time (SRT/HRT) in continuous flow reactors (Speece 1983). The loading rates of a process are largely dictated by the biomass retention in the reactor (Lettinga et al. 1980). Maximal sludge retention or biomass retention is desirable for process stability and minimal sludge production. Minimal HRT minimizes the reactor volume and thus reduces capital costs (Speece 1983). Continuously stirred tank reactors (CSTR) (Figure 15) are subjected to washout of active biomass (Speece 1983). Biomass retention has been enhanced by employing internal sedimentation systems and cationic flocculants (White and Gadd 1997). Anaerobic contact process (ACP) relies on biomass separation and recycling to increase the SRT/HRT (Figure 15) (Speece 1983). Several methods have been suggested for recovering biomass from the reactor effluent, including sedimentation, flocculation, centrifugation (Lettinga et al. 1980) and magnetic separation of sulfate-reducing bacteria (Watson et al. 1996; Bahaj et al. 1998).

Due to the slow growth rate and low biomass yield of SRB, various immobilized biomass reactors have gained increasing attention. In anaerobic filter reactors (AFR) biomass is retained as a biofilm on packing material as well as unattached in the packing interstices (Figure 15) (Speece 1983). AFRs have been operated in horizontal (Hammack and Edenborn 1992), upflow (Maree and Strydom 1985; Maree and Strydom 1987; Hammack et al. 1994a; Farmer et al. 1995; Elliott et al. 1998; Drury 1999; Estrada Rendon et al. 1999; Chang et al. 2000; La et al. 2003) or downflow modes (Wildeman et al. 1995; Groudeva et al. 1996; Zaluski et al. 1999). The downflow AFR allows the utilization of gravity and, thus, passive operation (Gusek and Wildeman 1999). Packing materials used in AFRs include cobbles (Zaluski et al. 1999), polypropylene pall rings (El Bayoumy et al. 1999b), glass beads (Kolmert et al. 1999; Kolmert and Johnson 2001) and alkaline minerals (Estrada Rendon et al. 1999; Zaluski et al. 1999). Biological sulfate reduction has been enhanced with solid organic materials (Béchard et al. 1993; Wildeman et al. 1995; Drury 1999; Estrada Rendon et al. 1999; Zaluski et al. 1999; Chang et al. 2000; Harris and Ragusa 2001) as well as liquid substrates (Maree and Strydom 1987; Elliott et al. 1998; El Bayoumy et al. 1999a; Kolmert et al. 1999; Glombitza 2001; Kolmert and Johnson 2001; Tuppurainen et al. 2002; Jong and Parry 2003). Solid substrates have a limited lifetime and have to be replaced or supplemented with liquid substrates once the original substrate is depleted (Tsukamoto and Miller 1999). AMD treatment using AFRs has been studied in laboratory (Maree and Strydom 1987), bench (Hammack et al. 1994a; Jong and Parry 2003), pilot (Dvorak et al. 1992; Farmer et al. 1995;

30 Wildeman et al. 1995) as well as full scale (Gusek and Wildeman 1999). The main shortcomings of AFRs are the channeling of the flow and clogging of the bed by precipitates (Somlev and Tishkov 1992).

G G I I E

E

Continuously stirred Anaerobic contact process (ACP) tank reactor (CSTR) G G I G E

E

I E Upflow anaerobic Downflow anaerobic I filter reactor (AFR) filter reactor (AFR) Fluidized-bed G E G E reactor (FBR)

G

E

I I I

Upflow anaerobic sludge Anaerobic hybrid blanket reactor (UASB) reactor (AHR) Anaerobic baffled reactor (ABR)

Figure 15. Various continuous flow reactors used in anaerobic wastewater treatment. I = wastewater influent, E = effluent, G = gas (Adapted from Speece 1983; Jung et al. 1997).

In the fluidized-bed reactor (FBR), channeling and clogging are avoided by fluidizing the inert biomass carrier (Figure 15) (Somlev and Tishkov 1992). Fluidization can be carried out with recycle water (Ma and Hua 1997) or using a gas stream in which case the reactor is called a gas lift reactor (van Houten and Lettinga 1995). Carrier materials used include iron chips (Somlev and Tishkov 1992), synthetic polymeric granules covered with iron dust (Somlev and Banov 1998), pumice particles (van Houten et al. 1994), porous glass beads (Nagpal et al. 2000b), and carbon dust (Ma and Hua 1997). The fluidized carrier provides a large surface area for biofilm formation (Speece 1983). Somlev and Tishkov (1992) compared the performance of a FBR and AFR to treat hydrometallurgical wastewater. Sulfate-reduction rates per reactor volume and carrier surface area were higher in the fluidized-bed reactor compared to the AFR (Somlev and Tishkov 1992). Combined removal of metals, acidity and sulfate using sulfate-reducing FBRs has not been previously extensively studied. The studies

31 by Somlev and Tishkov (1992), van Houten et al. (1994), and Nagpal et al. (2000b) were focused on the efficiency of sulfate reduction. Ma and Hua (1997) demonstrated Cd removal in lactate-fed laboratory scale FBR, but the pH change in the process was not reported. Somlev and Banov (1998) used FBR as one step in a three stage process for treating slightly acidic (pH 5.2) uranium mining wastewater. In addition to the FBR, the installation contained a secondary anaerobic stage supplied with crushed limestone, and an aerobic stage (Somlev and Banov 1998).

Table 10. Benefits and drawbacks of various reactor types. Reactor type Benefits (+) and drawbacks (-) Reference(s) Continuously + Consistent, reliable and rapid equilibrium Barnes et al. 1991a; Barnes stirred tank reactor conditions et al. 1991b (CSTR) - Poor retention of biomass Speece 1983 Anaerobic contact + Better retention of biomass as compared to CSTR Speece 1983 process (ACP) - Pumping of biomass breaks down flocks and sludge Speece 1983 Anaerobic filter + Low shear forces Barnes et al. 1991b reactor (AFR) + Longer sludge retention time than in CSTR Barnes et al. 1991b + Possibility to utilize gravitation in downflow mode Speece 1983 + Stripping of H2S effective in downflow mode Speece 1983 - Channeling of the flow possible Barnes et al. 1991b - Pressure gradients can be large Anderson et al. 1990 Fluidized-bed + Large surface area for biofilm formation due to Speece 1983; Anderson et reactor (FBR) fluidized carrier material al. 1990 + High retention of biomass on the carrier Marín et al. 1999 + Efficient mass transfer Marín et al. 1999 + Small pressure gradients Anderson et al. 1990 + No channeling of flow Anderson et al. 1990 + Influent concentrations diluted due to recycle flow Marín et al. 1999 + No clogging Speece 1983 + Selects for microbes with low Km values Melin et al. 1997; Melin et al. 1998 - Energy needed for carrier fluidization Speece 1983 - Shear forces can detach biomass Rittmann 1982 - Less volume available for biomass compared to the Yoda et al. 1989 UASB reactor due to the inert biomass carrier Upflow anaerobic + No channeling of flow Barnes et al. 1991b sludge blanket + No compacting of sludge Barnes et al. 1991b reactor (UASB) + No costs due to biomass carrier Speece 1983 + No clogging Speece 1983 + Possibility to obtain high treatment rates Speece 1983 - Biomass flush out common during process failures Speece 1983 - More susceptible to changes in influent quality Jhung and Choi 1995 compared to AFR Anaerobic hybrid + Less susceptible to clogging compared to AFR Steed et al. 2000 reactor (AHR) + Sludge removal easier than in AFR Steed et al. 2000 + Biomass retention better than in UASB Steed et al. 2000 Anaerobic baffled + Long sludge retention time Speece 1983 reactor (ABR) + No costs due to biomass carrier Speece 1983 + Good tolerance for shocks of hydraulic and organic Barber and Stuckey 2000 loading

In upflow anaerobic sludge blanket (UASB) reactors, biomass retention is based on good settling characteristics of granular sludge (Figure 15) (Lettinga et al. 1980; Speece 1983). The presence of methanogens in the biomass can enhance granulation (van Houten et al. 1994; Visser 1995; Omil et al. 1996). The produced biogas is trapped by a hood located below the water surface (Lettinga et al. 1980) and can be periodically burned in a flare (Scheeren et al. 1993). Due to the biomass granulation, no packing or carrier material is needed which reduces

32 the start-up costs of the UASB compared to AFR and FBR (Speece 1983). However, extensive biogas production may require extra instrumentation which increases capital costs. Moreover, methanogens compete for substrates (acetate and H2 + CO2) with sulfate reducers (Visser 1995; Oude Elferink 1998), resulting in decrease in the yields of H2S and alkalinity per amount of substrate added. Other problems encountered with UASB reactors are poor or slow granulation (Steed et al. 2000) and the rapid disintegration of the granular sludge under certain conditions (Speece 1983). Sulfidogenic UASB reactors have been used for metal recovery in bench (de Lima et al. 1996), pilot (Hammack and Dijkman 1999), demonstration (Scheeren et al. 1993; Dijkman et al. 1999) and full scale (de Vegt and Buisman 1995; de Vegt and Buisman 1996; de Vegt et al. 1997; de Vegt et al. 1998; Boonstra et al. 1999).

The anaerobic hybrid reactor (AHR) is a combination of UASB and AFR, where the granular sludge bed is in the lower section of the reactor and packing material in the upper section (Figure 15) (Colleran et al. 1998; Steed et al. 2000). Steed et al. (2000) compared the feasibility of the UASB, AFR and AHR for removing heavy metals from AMD. The performance of the AHR was superior based on effluent metal and sludge concentrations (Steed et al. 2000). Although the UASB reactor reduced soluble metal concentrations, it did not operate as an effective clarifier, and hence the concentration of total suspended solids in the effluent was high (Steed et al. 2000). The study did not address the removal of acidity by SRB, since the reactor pH was adjusted with a buffer solution (Steed et al. 2000). Another modification of the UASB reactor is an anaerobic baffled reactor (ABR) which is a staged reactor where biomass retention is enhanced by forcing the water through several compartments (Figure 15) (Barber and Stuckey 2000). ABRs and AHRs have been used for treating sulfate-laden and in some case metal-containing wastewaters (Colleran et al. 1998; Barber and Stuckey 2000; Steed et al. 2000), but their potential for concomitant removal of acidity, metals and sulfate has not been extensively studied.

Depending on the reactor type and process configuration, the metal sulfide sludge can be recovered from the bottom of the bioreactor, such as AHR (Steed et al. 2000), by back washing the AFR at regular intervals (Maree and Strydom 1987) or with a clarifier downstream of the precipitation unit (Hammack et al. 1994a). Several reactor-dependent, physico-chemical, microbiological and operational factors affect the efficiency of sulfate- reducing bioreactors (Table 11).

33 Table 11. Factors affecting the efficiency of sulfate reducing bioreactors. Factor Effect(s) Reference(s) Reactor design Available surface area Increases the potential populations of active SRB Bass et al. 1996; Lyew and for microorganisms and the amounts of sulfate reduced Sheppard 1997 Microbial composition Species Optimal growth conditions, substrates utilized and Hard et al. 1997; Widdel tolerance to toxic compounds vary among different 1988 species Attachment properties The ability of microbial species to colonise sludge Omil et al. 1997; Isa et al. of microorganisms granules or surfaces is variable 1986 Type and history of Competition between SRB and other Harada et al. 1994; Omil et inoculum microorganisms al. 1998; Lens et al. 2000 Influent composition Sulfate concentration Affects SRB growth and activity; SRB may be out- Overmeire et al. 1994; competed at low sulfate concentrations; Inhibition White and Gadd 1996a; at high concentration Crine et al. 1999 Metal concentrations Fe2+ required for SRB growth, high heavy metal de Smul and Verstraete concentrations toxic; Presence of Ca2+ and Mg2+ 1999; Hao 2000; Marchal et may enhance the dominance of SRB over al. 2001 methanogens pH Affects the growth and activity of SRB; pH Visser et al. 1992; van decrease increases the free H2S concentration; High Houten et al. 1995; Omil et pH (> 8) can favor SRB over methanogens; pH al. 1996; Omil et al. 1997; affects the configuration and diameter of biomass Crine et al. 1999; Marchal et aggregates al. 2001; García et al. 2001 Operational conditions Substrate Affects SRB growth and activity; sulfate reduction Harada et al. 1994; White concentration and enhanced with increased substrate concentrations; and Gadd 1996a; 2- loading rate High SO4 /substrate ratio favors SRB over other Kalyuzhnyi and Fedorovich microorganisms 1997; Omil et al. 1998; El Bayoumy et al. 1999b; Crine et al. 1999 Type of substrate Ethanol, propionate, butyrate and glucose can favor Polprasert and Haas 1995; growth of SRB whereas acetate and methanol may Omil et al. 1996; White and favor growth of methanogens; Lactate is good for Gadd 1996b; Greben et al. obtaining high biomass yields of SRB, Ethanol 2000 stimulates sulfide production; Glucose fermentation may decrease pH N and P addition Nutrient addition may be needed to achieve optimal El Bayoumy et al. 1999b; + C:N:P ratio; High NH4 concentrations toxic to Lens et al. 1999 bacteria Flocculant addition Improves settling of solids Scheeren et al. 1993 Temperature Affects the growth and activity of SRB and the Visser et al. 1992; Visser et solubility of gases; Temperature increase may help al. 1993; Crine et al. 1999; the SRB to outcompete methanogens Dean 1999 Exposure to oxygen SRB may better outcompete methanogens after a Omil et al. 1997 short term exposure to oxygen

H2S concentration High H2S concentration can inhibit SRB growth or Omil et al. 1996; van Houten favor SRB over other microorganisms et al. 1997 Stirring conditions Increased agitation power can decrease substrate Marchal et al. 2001 uptake Hydraulic retention Short HRT can lead to wash-out of biomass and Kalyuzhnyi and Fedorovich time (HRT) affect sulfate reduction rate 1997; Fedorovich et al. 2000

N2 flushing N2 flushing can decrease the concentration of toxic Marchal et al. 2001 H2S -1 Upward liquid velocity High υup (4-6 m h ) in granular sludge reactors can Omil et al. 1996 (υup) result in biomass washout

34 Biological sulfate reduction and metal precipitation using biogenic H2S can be applied in single or separated unit processes (Figure 16) (for a review, see Hao 2000). A single-stage approach for sulfate reduction and metal precipitation (Figure 16a) was used in the Palmerton pilot plant installed to treat metal-contaminated drainage from a smelting residues dump at the former New Jersey Zinc Company plant in Palmerton, Pennsylvania (Dvorak et al. 1992). The system consisted of two independent downflow AFRs filled with loosely packed spent mushroom compost (Dvorak et al. 1992). Another example of the single-state approach is a full-scale plant at the Budelco zinc refinery in Budel-Dorplein, Netherlands that has been used to remediate metal-containing groundwater (Scheeren et al. 1993). This technology, marketed under the trade name Thiopaq®, consists of two biological processes complemented with solid separations steps. The first biological step utilizes SRB in an ethanol-fed UASB reactor to generate alkalinity and produce H2S promoting the precipitation of metals as sulfides within the bioreactor. The second biological process involves an aerobic filter in which the excess sulfide is oxidized by aerobic colorless sulfur bacteria. The solids are removed in a tilted plate settler and a continuously cleaned sand bed filter is used as a solids polishing step before discharge. The elemental sulfur can be used for agricultural applications or to produce sulfuric acid (Scheeren et al. 1993; de Vegt and Buisman 1995; de Vegt and Buisman 1996; de Vegt et al. 1997; de Vegt et al. 1998; Boonstra et al. 1999). The Thiopaq technology has also been demonstrated for the treatment of heavily polluted AMD at the former Wheal Jane mine in Cornwall, UK (Boonstra et al. 1999). Boonstra et al. (1999) estimated that replacing the current active (lime) treatment operation employed at the Wheal Jane site with a biological plant would reduce the annual discharge from the site by > 600 kg iron, > 9900 kg zinc, > 120 kg copper, 3600 kg aluminium, 9600 kg manganese, 21 kg cadmium and 3895 tons of sulfate.

The single-stage treatment process (Figure 16a) is a low-cost solution for AMD treatment, but it may not be viable if the wastewater to be treated is very acidic or contains high concentrations of heavy metals (Hao 2000). Many single-stage treatment systems have utilized alkaline materials to generate additional alkalinity (Hammack and Edenborn 1992; Wildeman et al. 1995; Groudeva et al. 1996; Estrada Rendon et al. 1999; Gusek and Wildeman 1999; Zaluski et al. 1999; Harris and Ragusa 2001). In some cases several bioreactors have been used in series to enhance sulfate reduction and metal precipitation (Dvorak et al. 1992; Béchard et al. 1993). Another approach is to recycle part of the treated water to dilute the influent (Figure 16b) (Maree and Strydom 1987; Ma and Hua 1997; Steed et al. 2000). Recycling of the water requires additional pumps and thus increases the investment and operational costs (Hao 2000). In single-state processes the concentration of dissolved sulfide has to be maintained at a relatively high level to buffer the system against metal shock loads. On the other hand, dissolved sulfide is toxic to SRB, and therefore, sulfide product inhibition may be expected in single-stage processes (Kalyuzhnyi et al. 1997).

Metals can be pre-precipitated prior to the biological step by recycling either sulfide- containing water (Figure 16c) or H2S-containing gas (Figure 16d). Haas and Polprasert (1993) recycled sulfide-rich water for precipitating metals before a semi-continuous ACP process. Separation of the chemical sulfide precipitation and biological H2S production is also the basis of the patented BioSulphide process® (Rowley et al. 1997). The BioSulphide technology was demonstrated at the Britannia Copper Mine, in British Columbia, Canada (Figure 17) (Rowley et al. 1997). Copper and zinc were precipitated selectively in consecutive steps of the chemical circuit by using sulfide- and alkalinity-containing effluent from two bioreactors. Part of the treated water from the chemical circuit was fed to the bioreactors to provide sulfate for the SRB (Rowley et al. 1997). Based on the pilot results, the capital costs of a full-scale treatment plant at Britannia were estimated to be 2.2 million U.S. dollars (Rowley et al. 1997). For comparison, the capital costs of a lime treatment plant at Britannia were estimated to be 3.5 million U.S. dollars, and operating costs over 900 000 U.S. dollars per year without

35 consideration of sludge disposal costs. The BioSulphide plant was estimated to result in an annual net operating profit of 130 000 U.S. dollars, resulting from the sale of CuS and ZnS concentrates to smelters and mines (Rowley et al. 1997).

a) Metal- and sulfate- containing wastewater Sulfate reduction and metal precipitation

Substrate

- - Water recycle (HS , HCO3 ) b) Metal- and sulfate- containing wastewater Sulfate reduction and metal precipitation

Substrate

- - Water recycle (HS , HCO3 ) Metal- and sulfate- 2- c) containing wastewater Metal Water (SO4 ) Sulfate precipitation reduction

Substrate

Gas recycle (H2S) Metal- and sulfate- d) 2- + containing wastewater Metal Water (SO4 , H ) Sulfate precipitation reduction

Substrate

Metal-containing wastewater e) Sulfate-containing - - wastewater Sulfate- Water (HS , HCO3 ) Metal reduction precipitation

Substrate

Figure 16. Possible layouts of reactor processes utilizing sulfate-reducing bioreactors for metal precipitation (Adapted from Hao 2000).

36 H2+CO2 nutrients

SRB SRB Water 2- bioreactor bioreactor (SO4 ) Water - - (HS , HCO3 )

AMD Feed CuS ZnS Stripping Treated influent tank clarifier clarifier vessel effluent

CuS ZnS concentrate concentrate

Figure 17. A simplified flow diagram of the demonstration plant utilizing sulfate-reducing bacteria (SRB) for acid mine drainage (AMD) treatment at the Britannia Copper Mine, in British Columbia, Canada (adapted from Rowley et al. 1997).

Recycling of H2S-containing gas (Figure 16d) may assist the selective precipitation of valuable metals (such copper which precipitates as sulfide already at low pH), since no alkalinity is introduced to the precipitation step. However, the metal sulfide precipitation produces protons adding up to the acidity being fed to the bioreactor. Therefore, gas recycling has often been used in combination with chemical neutralization (Hammack et al. 1994a; Hammack et al. 1994b) or water recycling (de Vegt et al. 1997; Boonstra et al. 1999; Dijkman et al. 1999; Hammack and Dijkman 1999). Bhagat et al. (2004) compared the precipitation of metals using either biogenic H2S gas or liquid supernatant from a sulfate-reducing sequencing batch reactor. The precipitation using liquid supernatant resulted in higher metal removal efficiencies than the sparging with H2S gas (Bhagat et al. 2004). The precipitates produced comprised mainly metal sulfides, but in case of the liquid supernatant precipitation, they also contained unidentified complexes which may include nitrates, chlorides or carbonates (Bhagat et al. 2004). Hammack et al. (1994a) used biogenic H2S and NH4OH to recover copper and zinc selectively in a bench scale system treating acidic water from the Rio Tinto Mine, Nevada, USA The recycling of both H2S gas and bicarbonate-containing water has been applied in a pilot plant treating water from the Berkeley Pit, an abandoned open-pit copper mine in Butte Montana, USA (Hammack and Dijkman 1999). Hammack and Dijkman (1999) estimated that the costs of a full-scale treatment (19 000 m3 per day) of Berkeley Pit water utilizing ethanol as substrate in the UASB reactor would be 2.0 U.S. dollars per m3 water treated based on an annuity of 15 %. An alternative biological treatment using a gas lift reactor and steam-reformed natural gas as the substrate would cost 0.65 U.S. dollars per m3 water treated. Hammack and Dijkman (1999) estimated that plants using steam-reformed gas are more economical for treatment applications requiring sulfate removal in excess of 10 tons per day.

If the metals and sulfate to be removed are in different water fractions, a treatment approach depicted in Figure 16e can be used. This idea has been applied in the demonstration plant at Kennecott’s open pit copper mine in Bingham Canyon, Utah, USA (Figure 18) (de Vegt et al. 1997; de Vegt et al. 1998; Boonstra et al. 1999; Dijkman et al. 1999). Part of the H2S produced in the bioreactor is used as a gaseous form to selectively recover copper from a leach water stream. The bioreactor liquid effluent, which is rich in sulfide and alkalinity, is further used to precipitate metals from sulfate-containing water, and to produce elemental sulfur. The sulfate-containing effluent from the metal precipitator is fed to the bioreactor to

37 maintain biological sulfate reduction (de Vegt et al. 1997; de Vegt et al. 1998; Boonstra et al. 1999; Dijkman et al. 1999). De Vegt et al. (1997) estimated that the cost of a treatment system (15 % annuity) reducing 10 000 tons of sulfate per year would be approximately 330 U.S. dollars per ton of sulfate reduced when H2 and CO2 are used as energy and carbon source for the SRB.

- - Water (HS , HCO3 ) H2S Gas- Leach water Aerobic SRB Metal- liquid bioreactor bioreactor Sulfur sulfide contactor CuS clarifier Water Gas clarifier clarifier 2- (SO4 ) recycle

Air H2+CO2 Elemental Metal CuS sulfur sulfides concentrate Treated effluent AMD influent, nutrients

Figure 18. A simplified flow diagram of the demonstration plant utilizing sulfate-reducing bacteria (SRB) for treating acid mine drainage (AMD) and leach water at Kennecott’s open pit copper mine in Bingham Canyon, Utah, USA (adapted from de Vegt et al. 1997; de Vegt et al. 1998).

The separation of the biological sulfate reduction and metal sulfide precipitation with the biogenic sulfide alleviates toxicity on SRB, allows selective metal precipitation by the control of pH and H2S dosing, and reduces the amount of biomass and organic substrates in the metal sulfide sludge (Hammack et al. 1994a). Drawbacks are the increasing investment and operational costs due to increased and more complex instrumentation.

Biological treatment of AMD has several major advantages over traditional active chemical treatment (for a review, see Johnson 2000). Operating costs of the latter are high, sulfate and metal removal efficiencies tend to be relatively low, bulky sludge requiring dewatering and landfilling is produced, and potentially valuable metals which accumulate in the sludge are eventually lost. Biological treatment facilitates the recovery of metals (and sulfur when integrated with elemental sulfur production), the sludge produced is far denser (some 6-10 fold), and sulfate concentrations can be lowered to potable water levels. On the other hand, biological treatment requires sophisticated and expensive engineering systems, and the process needs an energy and carbon source for the bacteria (Johnson 2000). Compared to passive biological treatment, active bioreactors are more compact, offer consistent performance and control, and allow the recovery metals and sulfur (Johnson 2000). However, bioreactor plants require significant start-up capital and continuous monitoring (Johnson 2000). The overall costs of a biological treatment plant depend on the characteristics of the wastewater stream and the discharge criteria (de Vegt et al. 1997; Hammack and Dijkman 1999).

38 7 SULFATE-REDUCING PROKARYOTES

Sulfate-reducing prokaryotes (SRP) are a unique physiological group of microorganisms because they have the capability of using sulfate as the final electron acceptor in anaerobic respiration. SRP are broadly distributed on earth and display significant roles in nature by virtue of their potential for numerous interactions (Barton and Tomei 1995).

7.1 Taxonomy

Sulfate-reducing prokaryotes (SRP) constitute a diverse group of microorganisms (for reviews, see Thauer and Kunow 1995; Stackebrandt et al. 1995; Castro et al. 2000; Garrity et al. 2003). Various properties have been used in traditional classification of SRP, including cell shape, motility, guanine plus cytosine (GC) % content of DNA, type of sulfite reductases (desulfoviridin, P582, desulforubidin, desulfofuscidin), electron transfer proteins (cytochromes), respiratory menaquinones, fatty acids, optimal growth conditions and the ability to oxidize acetate (for reviews, see Akagi 1995; Chen et al. 1995; Stackebrandt et al. 1995; Castro et al. 2000). Electron donor utilization profiles are used for classification within a particular genus. The analysis of ribosomal RNA (rRNA) sequences has led to the reclassification of some SRP and allowed organization of various SRP species into four phyla of sulfate-reducing bacteria (SRB) (Proteobacteria, Firmicutes, Thermodesulfobacteria, Nitrospira) and one phylum of sulfate-reducing archaea (SRA) (Euryarchaeota) (Table 12) (Garrity et al. 2003; Mori et al. 2003; Audiffrin et al. 2003; Moussard et al. 2004).

The largest of the SRB groups is affiliated with the Proteobacteria, more specifically the delta (δ) class of Proteobacteria, which currently contains 32 gram-negative sulfate-reducing genera in 8 families. Most of the proteobacterial sulfate-reducers are mesophiles (Castro et al. 2000), but also some thermophilic and psychrophilic species have been described (Beeder et al. 1995; Isaksen and Teske 1996; Knoblauch and Amann 1999; Knoblauch et al. 1999). The phylum Firmicutes includes low DNA GC % content sulfate-reducers which are affiliated with the family of Peptococcaceae. Currently three genera, Desulfonispora, Desulfosporosinus and Desulfotomaculum, have been described within Peptococcaceae, all of which form heat-resistant endospores. The members of this group may stain gram positive or negative, but have a gram-positive cell-wall ultrastructure based on electron microscopic studies (Sleytr et al. 1969; Stackebrandt et al. 1995; Stackebrandt et al. 1997; Plugge et al. 2002). The family of Peptococcaceae contains both mesophilic and moderately thermophilic SRB (Castro et al. 2000). Thermophilic SRB have also been found in two other phyla, namely Thermodesulfobacteria and Nitrospira. These both contain, so far, only one SRB genus Thermodesulfobacterium and Thermodesulfovibrio, respectively. Additionally, the genus Thermodesulfobium in the family of Thermodesulfobiaceae awaits classification (Mori et al. 2003). The archaeal sulfate-reducers exhibit optimal growth temperatures above 80 ºC. Only one SRA genus, Archaeoglobus, has been described to date within phylum Euryarchaeota (Thauer and Kunow 1995; Castro et al. 2000; Garrity et al. 2003). The SRA are characterized by the lack of peptidoglycan cell wall, the presence of phytanyl ether lipids, and a relatively high salt optimum (1.5-1.8 % NaCl) (Thauer and Kunow 1995). The cell envelope of SRA consists of a single surface layer of glycoprotein subunits, which are closely associated with the outside of the cytoplasmic membrane (Thauer and Kunow 1995).

39 Table 12. Phylogeny of sulfate-reducing prokaryotes (Garrity et al. 2003; Mori et al. 2003; Audiffrin et al. 2003; Moussard et al. 2004) Phylum Class Order Family Genus Proteobacteria δ-Proteobacteria Desulfovibrionales Desulfovibrionaceae Desulfovibrio Desulfomicrobiaceae Desulfomicrobium Desulfohalobiaceae Desulfohalobium Desulfonatronovibrio Desulfonauticus Desulfothermus Desulfobacterales Desulfobacteraceae Desulfobacter Desulfobacterium Desulfobacula Desulfobotulus Desulfocella Desulfococcus Desulfofaba Desulfofrigus Desulfomusa Desulfonema Desulforegula Desulfosarcina Desulfospira Desulfotignum Desulfobulbaceae Desulfobulbus Desulfocapsa Desulfofustis Desulforhopalus Desulfotalea Desulfoarcales Desulfoarculaceae Desulfoarculus Syntophobacterales Syntrophobacteraceae Desulfacinum Desulforhabdus Desulfovirga Thermodesulfo- rhabdus Syntrophaceae Desulfobacca Desulfomonile Firmicutes Clostridia Clostridiales Peptococcaceae Desulfonispora Desulfosporosinus Desulfotomaculum Thermodesulfo- Thermodesulfo- Thermodesulfo- Thermodesulfo- Thermodesulfo- bacteria bacteria bacteriales bacteriaceae bacterium Thermodesulfatator Nitrospira Nitrospira Nitrospirales Nitrospiraceae Thermodesulfovibrio Unclassified Unclassified Unclassified Thermodesulfobiaceae Thermodesulfobium Euryarchaeota Archaeoglobi Achaeoglobales Archaeoglobaceae Archaeoglobus

7.2 Physiology

All SRP are characterized by their use of sulfate as a terminal electron acceptor during anaerobic respiration (for reviews, see Hansen 1994; Akagi 1995; Cypionka 1995; Hamilton 1998). Sulfate is unique among microbial electron acceptors in that it must be activated before it can be reduced (Hamilton 1998). Sulfate is activated by means of adenosine triphosphate (ATP). The enzyme ATP sulfurylase catalyzes the attachment of the sulfate to a phosphate of ATP, leading to the formation of adenosine 5'-phosphosulfate (APS) as shown in Figure 19 (Madigan et al. 2003). Although ATP is hydrolyzed, the reaction is energy requiring, and must be pulled to completion by the removal of the end products. Pyrophosphate (PPi) can be

40 hydrolyzed to phosphate by pyrophosphatase (Cypionka 1995). In dissimilative sulfate 2- reduction, the sulfate moiety of APS is reduced directly to sulfite (SO3 ) by the enzyme APS reductase with the release of adenosine monophosphate (AMP). In assimilative reduction, another phosphate group is added to APS to form phosphoadenosine 5'-phosphosulfate (PAPS), and only then is the sulfate moiety reduced to sulfite with the release of phosphoadenosine 5'-phosphate (PAP) (Madigan et al. 2003).

a) b) ATP (Adenosine triphosphate) ATP PPi ATP ADP 2- PAPS OH OH OH SO4 APS 1 6 Adenine CH -O-P-O-P-O-P-OH NADPH + 2 e- O 2 - 2 2 e HH O O O HHH AMP NADP+ PAP OH OH SO 2- SO 2- 3 3 APS (Adenosine 5‘-phosphosulfate) 2 e- 3 OH O 6 e- Adenine O CH -O-P-O-S-OH 3 S O 2- 2 6 e- 3 6 HH O O 2 e- 4 H H

2- S2O3 OH OH 5 PAPS (Phosphoadenosine 5‘- 2 e- phosphosulfate) OH O

H S H2S 2 Adenine O CH2-O-P-O-S-OH HH O O H H Organic sulfur OH Excretion compounds OH O-P-OH Dissimilatory Assimilatory O sulfate reduction sulfate reduction Figure 19. a) Schemes of dissimilative and assimilative sulfate reduction. The enzymes catalyzing the reactions include (1) ATP sulfurylase, (2) APS reductase, (3) sulfite reductase, (4) trithionate reductase, (5) thiosulfate reductase, and (6) APS kinase. b) Sulfate is activated by adenosine triphosphate (ATP) leading to the formation of adenosine 5'-phosphosulfate (APS) and phosphoadenosine 5'-phosphosulfate (PAPS) (Adapted from Akagi 1995; Madigan et al. 2003).

The sulfite is reduced to H2S through a pathway that has not been resolved to date (for reviews, see Akagi 1995; Cypionka 1995). One hypothesis is that a direct six-electron reduction of sulfite to sulfide occurs by the help of sulfite reductase without the formation of any isolable intermediate compound(s) (Figure 19) (Akagi 1995). Another hypothesis proposes a pathway, in which two intermediates are formed, trithionate and thiosulfate. This trithionate pathway involves a recycling mechanism of sulfite that is released during the reduction of trithionate to thiosulfate by the trithionate reductase, and the reduction of thiosulfate to sulfide by the thiosulfate reductase (Akagi 1995). In the dissimilative sulfate reduction, H2S is excreted into the environment whereas in the assimilative reduction, the H2S formed is immediately converted into organic sulfur compounds, such as amino acids (Madigan et al. 2003).

2- Sulfate and other sulfur-oxy aniones (SOx ) are taken up into the cells via co-transport with H+ or Na+ (marine strains) (Cypionka 1995) driven by the energy in the proton motive force (Figure 20) (Madigan et al. 2003). The energy investment for sulfate uptake will be regained when sulfide finally leaves the cell as H2S (Figure 20) (Cypionka 1995). Thus, before a cell

41 can derive any energy from dissimilative sulfate reduction, it must first expend considerable energy to transport sulfate and then to convert it to an activated form. The energy yield from sulfidogenic oxidation of electron donors is significantly lower than the energy available from aerobic respiration. As a consequence, large quantities of H2S are produced as a result of the extensive electron donor turnover necessary to supply sufficient energy for anabolic processes and growth (Hamilton 1998).

H+ ED e- - + HS H+ or Na+ + ETS + + _ + H+ 2- _ _ SO4 _ _ SP e- + ED’ _ + SO 2- H S + _ x 2 _ + _ ADP+ _ + Pi H+ or Na+ SP H+ ATPase H+ _ 2- _ SO3 + ATP + _ _ + _ ADP+ _ + + ATP Pi + SP _ _ AP _ ATPase _ Na+ + + H or Na + + 2- + H S2O3 + +

H+

Figure 20. A generalized scheme summarizing proton coupled activities in sulfate-reducing prokaryotes. SP = H+ + 2- + + or Na driven symporter that accounts for the transmembrane uptake of sulfur-oxy aniones (SOx ). AP = H /Na antiporter system. ATPase = ATP synthase that synthesizes ATP by proton motive force, or contributes to the charge across the membrane by pumping H+ outward. ETS = Electron transport system, that obtains electrons from the external or internal oxidation of electron donors (ED or ED’, respectively) and supplies electrons to 2- SOx reduction (adapted from Fauque et al. 1991; Cypionka 1995; Madigan et al. 2003).

In addition to sulfate, SRP can use a variety of alternative electron acceptors. Depending on the strain, other sulfur compounds (sulfite, thiosulfate, elemental sulfur, organic sulfur compounds) (Figure 21) (Badziong and Thauer 1978; Cord-Ruwisch and Garcia 1985; Lie et al. 1999), nitrate and nitrite (Widdel and Pfennig 1982; Barton et al. 1983; Dalsgaard and Bak 5+ 6+ 6+ 1994), CO2 (Klemps et al. 1985), As (Newman et al. 1997), U (Panak et al. 1998), Cr (Chardin et al. 2002; Michel et al. 2003), Mn+4 (Lovley and Phillips 1994), Fe3+ (Tebo and Obraztsova 1998), Tc7+ (Lloyd et al. 2001), Pd2+ (Lloyd et al. 1998), fumarate (Widdel and Hansen 1992) and even oxygen can be used (Krekeler and Cypionka 1995). In fact, SRP may use preferably some alternative electron acceptors, since their reduction yields more energy (Krekeler and Cypionka 1995). For example sulfite and thiosulfate do not require activation and therefore allow higher energy yields (Krekeler and Cypionka 1995). However, not all electron acceptors support growth of pure cultures (Lovley and Phillips 1994; Krekeler and Cypionka 1995). Incomplete reduction of sulfate and sulfite to thiosulfate by SRP has been observed under conditions of electron donor limitation (Figure 21) (Fitz and Cypionka 1990; Sass et al. 1992; Cypionka 1994). Some SRP are able to disproportionate sulfite, thiosulfate or elemental sulfur to produce both sulfate and H2S in an energy yielding reaction (Figure 21) (Bak and Pfennig 1987; Lovley and Phillips 1994). Some SRP can oxidize inorganic sulfur

42 compounds in the presence of O2, nitrate or nitrite (Figure 21) (Dilling and Cypionka 1990; Dannenberg et al. 1992; Cypionka 1994; Fuseler and Cypionka 1995; Krekeler and Cypionka 1995). In the absence of an electron acceptor, many SRP can ferment organic substrates (Widdel and Hansen 1992).

Increasing redox potential

Increasing supply of organic substrate

Complete Incomplete Disproportionation Oxidation with Sulfur Redox reduction reduction to to sulfate plus oxygen, nitrate or compound state to sulfide thiosulfate sulfide nitrite

2- SO4 +6

2- SO3 +4

2- S2O3 +2

S0 0

H2S -2

Figure 21. Transformations of inorganic sulfur compounds by sulfate-reducing bacteria under various environmental conditions (adapted from Cypionka 1994).

SRP are able to oxidize a variety of substrates, ranging from simple carboxylic acids to more complex aromatic hydrocarbons (Akagi 1995). Most of the substrates are typical fermentation products and intermediate breakdown products of larger molecules (Figure 22) (for a review, see Hansen 1994). Direct utilization of biopolymers by SRP is very rare (Hansen 1994). The oxidation of organic substrates in SRP may be complete, leading entirely to the production of CO2, or incomplete, with acetate usually being the end product. Incomplete oxidation is due to the absence of a mechanism for acetyl-Co-A oxidation (Widdel and Hansen 1992). Complete oxidizers may excrete acetate when growing e.g. on ethanol or butyrate. With limiting amounts of substrates, the excreted acetate may be oxidized further (Widdel and Hansen 1992). Some incomplete oxidizers are able to use acetate as a source of carbon if H2 is used as the electron donor (Pikuta et al. 2000). Furthermore, some SRP are true autotrophs being capable of growing with CO or CO2 as a sole carbon source (Klemps et al. 1985; Min and Zinder 1990).

SRP need nitrogen, phosphorus and trace nutrients for growth. Ammonium salts or urea and phosphate are good sources for nitrogen and phosphorus, respectively (Barnes et al. 1991b). Several strains can fix N2 and use nitrogen from amino acids as nitrogen source (Barton and Tomei 1995). Some strains can utilize nitrate, nitrite, and 2,4,6-trinitrotoluene (TNT) as a nitrogen source and electron acceptor (Barton and Tomei 1995). Metals are important cofactors for several enzymes of SRP. Nickel and selenium are required for hydrogenase activity. Iron is an essential element for cytochrome and hydrogenase production (Barton and Tomei 1995). High concentrations of heavy metals are toxic to SRP. For example zinc has been reported to be inhibitory in concentrations starting from 13-65 mg l-1 (Morton et al. 1991; Ueki et al. 1991; Hao et al. 1994; Poulson et al. 1997).

43

Proteins, carbohydrates Lipids

Hydrolysis

Amino acids, sugars Long chain fatty acids

Fermentation ß-oxidation

Carboxylic acids, alcohols

Acetogenesis

Acetate H2 , CO2 Methanogenesis

CH4, CO2

Sulfate reduction Sulfate reduction

CO2

SO 2- H S H S SO 2- 4 2 2 4 Figure 22. Anaerobic degradation of organic compounds in the presence of sulfate (adapted from Visser 1995).

The optimum pH for the growth of most known SRP is around 7 (Hao et al. 1996), but many species can tolerate pH values between 5 and 9.5 (Postgate 1984). Also some acidophilic or acidotolerant SRB have been found (Hard et al. 1997; Elliott et al. 1998; Johnson 2000). Sen and Johnson (1999) isolated acidophilic SRB at pH 3.6 on solid growth medium. Additionally they reported alkalinity production in a fermentor culture even at pH of 1.73 (Sen and Johnson 1999). SRB can also create microenvironments, where pH is higher than in the surrounding environment (Widdel 1988).

7.3 Ecological and economical importance

SRP are widely distributed on the earth: they have been found in soils, fresh, marine and brackish waters, hot springs and geothermal areas, oil and natural gas wells, sulfur deposits, estuarine muds, sewage, salts pans, corroding iron, the rumina of sheep and guts of insects (for a review, see Postgate 1984). In addition to their obvious importance to the sulfur cycle, SRP contribute to a variety of other processes in anaerobic environments (Table 13) (for reviews, see Postgate 1984; Widdel 1988; Barton and Tomei 1995; Hao et al. 1996; Hulshof Pol et al. 2001; Lens et al. 2002). For many years, sulfate reduction has been associated with the nuisance it causes during the anaerobic treatment of organic sulfate-rich waste and wastewater, toxicity and odor problems as well as the damage caused to materials (Table 13) (Barton and Tomei 1995; Hulshoff Pol et al. 2001; Lens et al. 2002). Since 1990s, interest has grown in applying SRP for biotechnical applications including remediation of wastewaters, gases and solid materials (Table 14) (for a review, see Lens et al. 2002).

44

Table 13. Effects of sulfate-reducing bacteria on natural and engineered environments. Effect Reference(s) Formation of metal sulfide deposits Shimazaki et al. 1985; Ehrlich 2002 Alteration of pH Postgate 1984 Degradation of organic matter Postgate 1984 Cycling of nutrients Postgate 1984; Barton and Tomei 1995 Mercury methylation Barton and Tomei 1995; King et al. 2000 Mobilization of radium from sulfate-bearing waste materials Landa 2003 Animal associations Postgate 1984; Barton and Tomei 1995; Deplancke et al. 2000

Fish kills due to toxic H2S Müezzinoğlu et al. 2000 Toxicity problems in anaerobic digestion O’Flaherty and Colleran 2000 Reduction in methane yield in anaerobic digestion O’Flaherty and Colleran 2000

Malodor of H2S Müezzinoğlu et al. 2000 Souring of oil and gas reservoirs Rosnes et al. 1991; Lizama and Sankey 1993 Biocorrosion of iron, metal alloys and concrete Postgate 1984; Schick 1990; Morton et al. 1991; Barton and Tomei 1995; Hao et al. 1996; Beech et al. 1998; Pogrebova et al. 2001; Fang et al. 2002 Deterioration of food Postgate 1984; Barton and Tomei 1995 Deterioration of cutting emulsions Barton and Tomei 1995 Product modification in paper industry Postgate 1984; Barton and Tomei 1995

Table 14. Biotechnical applications using sulfate-reducing bacteria. Application Reference(s) Treatment of acid mine drainage Tuttle et al. 1969 Removal of sulfate from artistic stone works Ranalli et al. 1997 Removal of sulfate from tannery wastewater Genschow et al. 1996 Removal of sulfate from landfill leachate Henry and Prasad 2000 Treatment of waste sulfuric acid Stucki et al. 1993 Treatment of organic sulfate-rich wastewaters Lens et al. 2000

Removal of SO2 from flue gases Selvaraj et al. 1997a; Selvaraj et al. 1997b; Weijma et al. 2000; Lens et al. 2003 Treatment of waste gypsum Maree and Hill 1989 Remediation of metal contaminated soil White et al. 1997; Groudev and Spasova 1997 Biotransformation of explosives Boopathy et al. 1998 Degradation of organic pollutants Rueter et al. 1994; de Best et al. 1998; Phelps and Young 1999; Ndon et al. 2000; Franzmann et al. 2002 Electricity production with a fuel cell Cooney et al. 1996 Stimulation of SRB with sulfate-containing amendments van der Gon et al. 2001 as a means to reduce methane emissions from rice fields

45 8 HYPOTHESES AND AIMS OF THE PRESENT WORK

In the present work, a one-stage sulfidogenic fluidized-bed treatment process for removing soluble metals and acidity from sulfate-containing wastewater was studied and developed. It was hypothesized, that continuous flow FBRs can be used to selectively enrich and maintain SRB. FBRs have been reported to efficiently retain biomass and allow high reaction rates (Speece 1983; Shieh and Keenan 1986; Marín et al. 1999). Moreover, the recycle flow in the FBR dilutes influent concentrations (Marín et al. 1999). Therefore, it was hypothesized that sulfidogenic FBRs are efficient in treating acidic metal-containing wastewater that would be toxic to suspended cells of SRB. Due to the efficient biomass retention it was expected that the FBR could be operated at low HRTs. As the fluidization of the carrier material affects the biofilm formation and the mass transfer in the FBR (Rittmann 1982; Speece 1983; Anderson et al. 1990; Marín et al. 1999), it was expected that the fluidization rate of the carrier material affects the process start-up and maintenance of the FBR. Various SRB species differ in their ability to oxidize different electron donors (see e.g. Widdel and Hansen 1992). Therefore, the choice of an electron donor was assumed to affect the process performance of the FBRs.

Previous research has shown that immobilized cells tolerate higher concentrations of toxic compounds than free cells (Keweloh et al. 1989). It has been reported that the outer layers can protect the inner layers of a biofilm from the inhibitory concentrations due to mass transfer resistance (Gantzer 1989). It was hypothesized that due to the protective nature of the biofilm, the sulfidogenic FBR process can be rapidly recovered after temporary shocks of acidity and high metal concentrations. As zinc and iron sulfides have lower solubility products of compared to the corresponding hydroxides (Dean 1999), it was expected that zinc and iron would precipitate as sulfides in the sulfidogenic FBR. It was also hypothesized that the metal precipitates do not clog the FBR but are flushed out due to the high upflow velocity in the FBR. Moreover, it was expected that the biomass yields obtained for the FBR enrichment cultures are comparable to those reported for pure cultures of SRB.

The high recycle ratio in the FBR ensures completely mixed conditions in the FBR (Rittman 1982). It was presumed that electron oxidation kinetics could be studied with batch FBR experiments and modelled using the Michaelis-Menten equation (Cornish-Bowden 1995). Sulfide product inhibition was expected in the single-state FBR processes where the concentration of dissolved sulfide has to be maintained at relatively high level to buffer the system against metal shock loads (Kalyuzhnyi et al. 1997). It was assumed that the sulfide inhibition of the FBR community can be described using the noncompetetive inhibition model (Maillacheruvu and Parkin 1996).

Previous studies report low species diversity in open ecosystems as a result of selection pressures, such as particular electron donors (Tiirola et al. 2002), electron acceptors (von Canstein et al. 2002), or inhibitors (Nyman 1999; Sandaa et al. 1999). Therefore it was hypothesized, that the bacterial diversity of the sulfidogenic FBRs, fed with single electron donors only, would be low. Additionally, it was expected that the choice of an electron donor affects the bacterial community structure of the FBRs. It has been recognized that the cultured and described species of bacteria represent only a minor fraction of the existing bacterial diversity (Amann et al. 1995). It was expected that the FBR communities enriched from mine sediments and methanogenic granular sludge contain novel bacteria that have not been previously described.

46 Based on the previous research and the hypotheses described above, the following aims were formulated for the present work:

• Evaluate the effects of carrier fluidization rate on the start-up and maintenance of the sulfidogenic FBR-process (Paper I)

• Evaluate the effects of the change of electron donor from lactate to ethanol on the FBR performance (Paper II).

• Determine the maximum metal loading and minimum influent pH for the effective precipitation of metals from acidic sulfate-containing water using FBR (Papers I and II).

• Evaluate the effects of HRT on the electron donor oxidation and the concomitant wastewater treatment performance in an ethanol-fed sulfate-reducing FBR (Paper III).

• Demonstrate the recovery of the FBR communities after process failures (Paper III).

• Estimate the biomass yield in the FBR (Paper I).

• Assess the composition and fate of the metal precipitates formed in the FBR (Paper I).

• Determine the maximum velocity (Vmax) and Michaelis-Menten constants (Km) for the oxidation of the ethanol and its major degradation intermediate, acetate in the FBR (Paper II).

• Determine the inhibition constants (Ki) of dissolved sulfide (DS) and H2S on the sulfidogenic oxidation of ethanol and acetate in the FBR (Paper III).

• Characterize the bacterial diversity of the ethanol- and lactate-utilizing FBRs using culture independent molecular (Paper IV) and phenotypic (Paper V) methods.

• Isolate and identify sulfate-reducing and other anaerobic bacteria from the ethanol- and lactate-utilizing FBRs (Paper V).

47 9 MATERIALS AND METHODS

9.1 Bioreactors

Fluidized-bed treatment of acidic wastewater was studied in laboratory scale FBRs (volumes 0.5-1.3 l) (Figure 23). Silicate mineral (Ø 0.5 to 1 mm, Filtralite, Norway) was used as the biomass carrier in the FBRs (Figure 24). The fluidization rates of the FBR carriers were 10-30 %. HRTs and loading rates were calculated based on the volume occupied by the fluidized carrier material. The FBR operation was compared to the performance of a laboratory scale UASB reactor (volume 2.35 l) (Figure 25). The temperature of the reactors was maintained at 35 °C with thermostatically controlled room or heating units. The FBRs and the UASB reactor were inoculated with sludge from a previous sulfate-reducing UASB reactor (Paper I) and with laboratory enrichment cultures originating from methanogenic granular sludge (Neson Oy, Jokioinen, Finland) and sediments from Outokumpu Pyhäsalmi Mine, Finland. A time diagram showing the timing of the various experiments conducted in the three FBRs and the UASB reactor is shown in Figure 26.

a Gas b bag Refrigerator

Sampling Effluent port

Water level adjustor Recycle flow Temperature probe and heating unit

Fluidized bed Glass beads

Refrigerator

Waste- water

Electron donor

Figure 23. Fluidized-bed reactor (FBR) a) configuration, b) photograph (Papers I-V).

48

Figure 24. Virgin silicate mineral (Ø 0.5 to 1 mm, Filtralite, Norway) used as the biomass carrier in the fluidized-bed reactors (Photo: Pasi Niskanen).

a Gas bag b

Sampling Refrigerator port Effluent

Recycle Water level flow adjustor Sludge bed

Glass beads

Refrigerator

Waste- water

Electron donor

Figure 25. Upflow anaerobic sludge blanket (UASB) reactor a) configuration, b) photograph (Paper I).

49 Bacterial isolations Continuous flow Continuous flow experiments Clone library PLFA experiments and DGGE DGGE DGGE [I] [II] [III] B1 B2 B3 B4 FBR1

FBR2

Electron FBR3 donors: Lactate UASB Ethanol

Time (d) 0100 200 300 400 500 600 700 800 900 1000 1100 1200 1300

Figure 26. A time diagram showing the timing of the various experiments conducted in the three fluidized-bed reactors (FBR) and the upflow anaerobic sludge blanket (UASB) reactor. Continuous flow experiments included experiments with increased feed Zn, sulfate and electron donor concentrations and decreased feed pH (Papers I and II), and experiments with decreased hydraulic retention time (Paper III). Two sets of batch experiments B1 and B2 addressed the sulfidogenic oxidation of ethanol and its major degradation product, acetate, respectively (Paper II) and other two sets B3 and B4 were conducted to study the sulfide inhibition of ethanol and acetate oxidation, respectively (Paper III). The feed to the FBRs between individual batch experiments was continuous. Microbiological studies included clone libraries and denaturing gradient gel electrophoresis (DGGE) of polymerace chain reaction (PCR) amplified 16S rRNA genes (Paper IV), Phospholipid fatty acid profiling (PLFA) (Paper V), and bacterial isolations (Paper V).

9.2 Continuous flow experiments

The reactors were fed with synthetic wastewater supplemented with lactate or ethanol as the electron donor for sulfate reduction. The FBR performance was evaluated with continuous flow experiments under increasing feed Zn, sulfate and electron donor concentrations and decreasing feed pH (Papers I and II), or decreasing hydraulic retention time (HRT) (Paper III). The effects of the change of electron donor from lactate to ethanol on the FBR performance were assessed (Paper II). The recovery of the bioprocesses after intentional overloading was demonstrated (Papers I and III). Additionally, the effects of carrier fluidization rate on the FBR, the composition and fate of the metal precipitates and biomass yields were evaluated as described in Paper I.

9.3 Batch kinetic experiments

Batch kinetic experiments were conducted in the ethanol-utilizing FBR to determine the maximum velocity (Vmax) and Michaelis-Menten constants (Km) for the oxidation of the ethanol and its major degradation intermediate, acetate (Reactions 13 and 14) (Paper II).

2- - - + 2 CH3CH2OH + SO4 Æ 2 CH3COO + HS + H + 2 H2O (13)

- 2- 2- - CH3COO + SO4 Æ 2 HCO3 + HS (14)

50 Additionally, the inhibition constants (Ki) of dissolved sulfide (DS) and H2S on the sulfidogenic oxidation of ethanol and acetate were determined (Paper III). Before and between individual batch experiments, the FBR was continually fed (HRT 24 h) with wastewater to wash out the excess hydrogen sulfide formed during the batch experiments, and to maintain a constant biomass content in the reactor. For each batch experiment, the feed flow was discontinued and the FBR was operated in recycle mode. A portion (50 ml) of the reactor liquid was replaced through the sampling valve placed in the recycling line with fresh solution (pH 7) containing Na2SO4, nutrients and ethanol or acetate, and in the inhibition experiments also NaS2. During the batch experiments, samples of the reactor liquid were collected through the sampling valve placed in the recycling line at regular intervals for ethanol and acetate analysis (Papers II and III).

9.4 Kinetic calculations

The ethanol and acetate oxidation rates obtained in the batch kinetic experiments were standardized to the total amount of biomass in the FBR determined as volatile solids (VS) and volatile suspended solids (VSS) (Papers II and III). Michaelis-Menten constant (Km) and maximum oxidation velocity (Vmax) were obtained using the Michaelis-Menten equation (Equation 15) and five different plotting approaches: Lineweaver-Burk plot (Equation 16), Hanes plot (Equation 17), Eadie-Hofstee plot (Equation 18), Direct linear plot (Equation 19) and Modified linear plot (Equation 20) (Cornish-Bowden 1995) (Paper II):

V ⋅ S v = max (15) Km + S 1 1 K 1 = + m ⋅ (16) v Vmax Vmax S S K 1 = m + ⋅ S (17) v Vmax Vmax v v = V − K ⋅ (18) max m S v V = v + ⋅ K (19) max S m 1 1 1 K = − ⋅ m (20) Vmax v S Vmax

In the equations 15-20, S = initial substrate concentration and v =oxidation velocity. Inhibition constants (Ki) for dissolved sulfide and H2S were calculated using the noncompetetive inhibition model (Equation 21) (Maillacheruvu and Parkin 1996) (Paper III):

V ⋅ S v = max , (21) ⎛ I ⎞ ⎜ ⎟ ()K m + S ⋅⎜1+ ⎟ ⎝ K i ⎠ where v = oxidation velocity, Vmax = maximum oxidation velocity, S = initial substrate concentration, Km = Michaelis-Menten constant, and I = inhibitor concentration.

51 9.5 Physico-chemical analyses

The physico-chemical analyses carried out in this work are summarized in Table 15.

Table 15. Summary of the physico-chemical analyses employed in this study. Analysis Instrument(s) Paper(s) Reference(s) Sulfate High performance liquid chromatograph I, II Papers I and II Sulfate Ion chromatograph II, III Paper II Dissolved organic Total organic carbon analyzer I, II Paper I carbon (DOC) Acetate, Ethanol Gas chromatograph with a flame II, III, V Papers II and V ionization detector Soluble metals Atom absorption spectrometer I, II, III SFS 1980a, SFS 1980b Dissolved sulfide Spectrophotometer I, II Trüper and Schlegel 1964 Dissolved sulfide Spectrophotometer II, III, IV, V Cord-Ruwish 1985 pH pH electrode I, II, III, IV, V Papers I, II and III Total alkalinity Potentiometric titration I, II, III SFS 1996 Total suspended Oven, balance I, II, III Paper I solids (TSS) Volatile suspended Furnace, balance I, II, III Paper I solids (VSS) Total solids (TS) Oven, balance I, II, III Paper I Volatile solids (VS) Furnace, balance I, II, III SFS 1990 Mineralogy of the X-ray diffractometer (XRD) I Paper I metal precipitates Elemental Scanning electron microscope with an I Paper I composition of the energy dispersive X-ray spectrometer metal precipitates (SEM-EDS)

9.6 Microbiological analyses

The bacterial diversity of the FBR communities was studied using clone libraries (Paper IV) and denaturing gradient gel electrophoresis (DGGE) (Paper IV) of polymerace chain reaction (PCR) -amplified 16S rRNA genes, and community phospholipid fatty acid profiling (Paper V) (Tables 16 and 17, Figure 27). Moreover, sulfate-reducing and other anaerobic bacteria were isolated using modified Postgate and Pfennig media and anaerobic plate and roll tube methods. The bacteria were identified by 16S rRNA gene sequencing (Papers IV and V).

Table 16. Summary of the microbial methods used in this study. Analysis Paper(s) Reference(s) DNA extraction and purification IV Plumb et al. 2001 Polymerase chain reaction (PCR) IV, V Papers IV and V Cloning IV Paper IV Denaturing gradient gel IV Muyzer et al. 1996; V electrophoresis (DGGE) Isolation of bacteria V Paper V Phospholipid fatty acid (PLFA) V Bligh and Dyer 1959; White et al. 1979; Guckert et al. profiling 1985; Skerratt et al. 1991; V DNA sequencing IV, V Papers IV and V

52 Biofilm on Extraction and analysis FBR carrier of PLFAs Community PLFA profiles

Enrichment of Community DNA bacteria extraction

Isolation of bacteria

PCR Rolltubes Partial 16S rRNA genes 16S rRNA Plates in with GC clamp genes anaerobic jars

DGGE Cloning to Escherichia 12 345 coli

Bacterial isolate in liquid culture

Plasmid Excising DNA purification fragments from the gel Direct lysis PCR PCR

16S rRNA Partial 16S rRNA gene gene without GC clamp Phylogenetic Sequencing tree

CTGAATCGTA Phylogenetic analysis

Figure 27. Analysis of the bacterial diversity of the fluidized-bed reactors (FBR) by denaturing gradient gel electrophoresis (DGGE) and cloning of polymerace chain reaction (PCR) -amplified 16S rRNA genes, bacterial isolations as well as phospholipids fatty acid (PLFA) profiling.

Table 17. Computer programs used in the microbiological analyses. Program Paper(s) Reference(s)/Source(s) BLAST IV, V Altschul et al. 1990, http://www.ncbi.nlm.nih.gov/BLAST/ Chromas IV http://www.technelysium.com.au/chromas.html Bioedit IV, V http://www.mbio.ncsu.edu/BioEdit/bioedit.html ARB IV, V http://www.mikro.biologie.tu-muenchen.de Chimera check IV Ribosomal Database Project, program http://rdp.cme.msu.edu/cgis/chimera.cgi?su=SSU Treeview IV, V Page 1996, http://taxonomy.zoology.gla.ac.uk/rod/treeview.html Excel 2000 IV Microsoft Office 2000, Redmond, WA, USA

53 10 RESULTS AND DISCUSSION

10.1 Start-up and process performance

The potential of sulfate-reducing FBR and UASB processes for concomitant removal of acidity, metals and sulfate from wastewater was studied with continuous flow experiments (Paper I). The three FBRs were operated with different fluidization rates to optimize the recycle flow of the reactors. In terms of sulfate reduction, sulfide production and effluent alkalinity, the start-up of the FBR with the 10 % fluidization rate was superior to the FBRs with 20-30 % fluidization rates. However, the FBRs with 20-30 % recycling rates performed better at the highest loading rates (Paper I). The results show that a low recycling rate, and, hence, smaller attrition supported faster biofilm formation on the carrier material during the enrichment period. On the other hand, at the higher loading rates, the FBRs with higher fluidization rates had higher sulfate reduction rates. At the higher recycle rates, dilution of metal concentrations and acidity were efficient. Another contributing factor is the improved mass transfer in the biofilm due to the higher upflow velocity (Shieh and Keenan, 1986). The results show that the high recycling rate of the FBRs allows treatment of wastewater with high concentrations of Zn and low pH (Paper I).

During the process start-up and the first loading experiments, lactate was used as an electron donor for sulfate reduction since it is known to be a good substrate for enriching SRB (Widdel 1988). The use of lactate may speed up the process start-up due to the relatively high growth yields of SRB growing on lactate (Nagpal et al. 2000a). However, in large scale operations the use of lactate would result in high operational costs (Nagpal et al. 2000a). A low-cost substrate, such as ethanol, is needed for large scale operations, once a reasonable biomass yield has been achieved (Nagpal et al. 2000a). In this study, the original lactate feed for enrichment and maintenance of one FBR culture was replaced stepwise with ethanol over 50 days at a HRT of 16 h. The change of the electron donor from lactate to ethanol did not significantly affect the FBR performance (Paper II). During the electron donor change, 51-72 % of the DOC was consumed by sulfate reduction with 80–97 % DOC removal. No significant acetate accumulation was observed during the electron donor change (Paper II).

10.1.1 Loading capacity

The robustness of the FBR was studied by increasing stepwise the Zn, sulfate and electron donor feed concentrations and decreasing the feed pH (Papers I and II), as well as decreasing the HRT (Paper III) until process failures occurred. A summary of the bioreactor performance during the continuous flow experiment along with performance of previously reported one- stage treatment systems using SRB is shown in Table 18. At a HRT of 16 h, both lactate- and ethanol-utilizing FBRs precipitated 350-360 mg Zn l-1 d-1 and 86 mg Fe l-1 d-1 from acidic wastewater containing 230-240 mg Zn l-1 and 57-58 mg Fe l-1, and increased the wastewater pH from 2.5 to 7.5-8.5 (Papers I and II). The corresponding percentual metal removal was over 99.8 % and effluent soluble Zn and Fe concentrations below 0.1 mg l-1. The residual DOC concentration during stable reactor performance with the highest loadings was 7-25 mg l-1 (Papers I and II). The FBRs allowed higher rates of metal precipitation and acidity neutralization than the UASB at a HRT of 16 h (Paper I) (Table 18).

Maximum metal precipitation rates were obtained with the ethanol-fed FBR at a HRT of 6.5 h: over 600 mg Zn l-1 d-1 and 300 mg Fe l-1 d-1 from wastewater containing 176 mg Zn l-1 and 87 mg Fe l-1 (Paper III). Under these conditions, ethanol oxidation was incomplete, percent

54 metal precipitation was over 99.9 %, and effluent metal concentrations remained below 0.08 mg l-1 (Paper III). The alkalinity produced in ethanol oxidation increased the initial pH of 3, resulting in effluent pH of 7.9-8.0 (Paper III).

The rates of metal precipitation and neutralization of acidity compare favourably with the rates obtained in previous studies with one-stage SRB-reactor systems (Table 18). In previous studies, high metal removal rates from acidic wastewater have been reached with systems where biomass retention is enhanced with flocculants (Table 18) (Barnes et al. 1991a; Scheeren et al. 1993), neutralization of the wastewater is enhanced chemically (Table 19) (Groudeva et al. 1996), or metal precipitation and sulfate reduction takes place in separate reactor units (Table 19) (de Vegt et al. 1998; Hammack and Dijkman 1999). Ma and Hua (1997) attained higher Cd precipitation rates with a FBR fed with lactate and yeast extract than the rates obtained in this study. However, the pH of the wastewater used in their study was not reported (Ma and Hua 1997). Metal precipitation and pH neutralization are interdependent, since the precipitation reaction (Reaction 12) produces acidity. In the present research, the feasibility of the sulfate-reducing FBRs to precipitate zinc and iron from acidic wastewater and produce alkalinity to neutralize the wastewater was demonstrated.

Zinc has been reported to completely inhibit sulfate-reduction in the concentration range of 25 to 60 mg l-1 (Morton et al. 1991; Ueki et al. 1991). In the present study, the sulfate-reducing FBR was able to treat wastewater with 240 mg Zn l-1 resulting in an effluent with less than 0.1 mg Zn l-1 (Paper II). High recycle ratio of the FBR dilutes feed metal concentrations and the acidity. In the completely mixed FBR, bacteria became exposed to Zn concentrations close to those of the effluent, and thus, the treatment of wastewater with inhibitory concentrations of heavy metals and low pH becomes amenable.

This study focused on the treatment of simulated zinc and iron containing wastewater. The sulfate-reducing FBR process can potentially be used to precipitate other toxic metals, which have high affinity to sulfide ions, such as Cu, Pb, Ni and Cd. The low solubility required for highly toxic metals such as mercury are often achievable only by speciation as sulfides, since most metal sulfides have solubility several orders of magnitude lower than the corresponding hydroxides throughout the pH range (Conner 1990).

55 Table 18, p1/2. Examples of treatment systems for acid mine drainage or metallurgical wastewaters using one-stage reactors with sulfate-reducing bacteria. pH Metals Sulfate Rector and process Operation Substrate(s) Feed Removal Feed Removal Reference(s) type Feed Effluent (mg l-1) (g m-3d-1) (%) (mg l-1) (g m-3d-1)(%) FBRa HRTb = 16 h Lactate 2.5 7.8 Zn 230 350 99.9 2290 2220 65 Paper I T = 35 ºC Fe 58 86 99.9 FBR HRT = 16 h Ethanol 2.5 7.7 Zn 240 360 99.9 1920 2320 81 Paper II T = 35 ºC Fe 57 86 99.9 FBR HRT = 6.5 h Ethanol 3 7.9 Zn 176 635 99.9 2080 4290 57 Paper III T = 35 ºC Fe 87 315 99.9 FBR HRT = 1 h Lactate, NRc NR Cd 100 2390 99.5 NR NR NR Ma and Hua 1997 T = 35 °C yeast extract FBR HRT = 8 h Lactate, NR NR Cd 1000 3000 99.9 NR NR NR Ma and Hua 1997 T = 35 °C yeast extract FBR HRT = 6.8 h Molasses 5.2 6.8 Ni 5.7 19 93.0 2100 6860 92 Somlev and Banov T = 30-31 °C U 1.2 2.8 66.7 1998 Cu 0.6 2.0 93.3 Cd 0.5 1.1 60.0 Zn 0.3 0.82 76.7 As 0.4 1.4 95.0 Pb 0.2 0.64 90.0 UASBd HRT = 16 h Lactate 3.0 7.7 Zn 170 250 99.99 1650 1860 75 Paper I T = 35 ºC Fe 54 81 99.97 UASB with a calm HRT = 4 h Ethanol, 3.2 7.0 Zn 112 672 99.9 1480 7010 79 Barnes et al. 1991a zone above sludge T = 30 ºC nutrients Cd 3.47 21 99.8 (+ flocculant) Co 2.05 12 99.1 Cu 2.98 18 99.7 UASB HRT = 4 h Ethanol, NR NR Zn 230 1380 99.87 1200 6240 87 Scheeren et al. 1993 (demonstration scale T = 20 °C nutrients Cd 1.20 7.1 99.2 Thiopaq process, (+ flocculant) Fe 54 312 96.3 Budelco zinc Pb 9 54 99.8 refinery, Budel- Cu 2.1 12 98.6 Dorplein, Co 0.2 1.1 90.0 Netherlands) Ni 0.3 1.6 90.0 Mn 33 170 84.9 As 0.10 0.54 90.0 UASB HRT = 19 h Sewage sludge NR NR Ni 16 19.3 96.2 84 66 62 de Lima et al.1996 aFBR=Fluidized-bed reactor, bHRT=hydraulic retention time, cNR=not reported, dUASB=Upflow anaerobic sludge blanket reactor

56 Table 18, p. 2/2. Continued. pH Metals Sulfate Rector and Operation Substrate(s) Removal Feed Removal Reference(s) process type Feed Effluent Feed (mg l-1) (g m-3d-1)(%) (mg l-1) (g m-3d-1)(%) AHRa with recycle HRTb = 310 h Acetate NRc NR Fe 840 65 99.9 5000 300 77 Steed et al. 2000 T = 30 ºC Zn 650 50 99.9 Mn 275 21 97.5 Cu 127 10 99.9 Four sequential HRT = 20 h Dextrose, urea, 3.5 6.5 Fe 200 240 99.8 654- NR NR Béchard et al. 1993 downflow AFRsd saccarose Al 25 30 99.6 868 Upflow AFR HRT = 115-192 h Cow manure, 2.5- 6.5 Fe < 300 NR 84 400- NR 80- Drury 1999 T = 14-24 ºC saw dust, 3.5 Mn < 22 40 1000 98 cheese whey Zn <100 99.7 Cu < 100 99 Cd < 2 99 As < 1 89 Upflow AFRs HRT = 50-100 h Cow manure, 7.0 7.0 Zn 49-57 13-26 96-99 390- 70-170 70 Farmer et al. 1995 (pilot plant T = 12-16 ºC paper products, Mn 1.8-2.4 0.5-1.0 71-91 440 Colorado, USA) alfalfa Cd 0.094-0.12 0.03-0.06 95-99 Upflow AFRs with HRT = 12 h Molasses 5.0 8.1 Fe 2.68 5.1 95.9 1331 2230 84 Maree and Strydom recycling T = 21-38 ºC Zn 0.32 0.59 92.2 1987 Mn 0.34 0.63 92.7 Al 0.934 0.18 9.6 Cr 0.11 0.15 68.2 Cd 0.01 0.01 50.0 Ni 0.135 0.17 63.0 Pb 0.135 0.17 63.0 Upflow AFR HRT = 20 h Acetate NR NR Cu < 200 240e 100e 2500 410f 14f El Bayoumy et al. T = 20 ºC Ni < 150 180 100 1999a Zn < 150 180 100 Cr < 75 89 100 Cd < 50 59 100 Pb < 40 47 100 Upflow AFR with HRT = 310 h Acetate NR NR Fe 840 65 99.8 5000 300 77 Steed et al. 2000 recycle T = 30 ºC Zn 650 50 99.9 Mn 275 21 97.1 Cu 127 10 99.9 aAHR = anaerobic hybrid reactor, bHRT = hydraulic retention time, cNR = not reported, dAFR = anaerobic filter reactor, eMetals were tested separately, fIn the absence of metals

57 Table 19, p.1/2. Examples of treatment systems for acid mine drainage or metallurgical wastewaters using reactors with sulfate-reducing bacteria in combination with alkaline chemicals/materials and/or pre-precipitation of metals. pH Metals Sulfate Rector and process Operation Substrate(s) Feed Removal Feed Removal Reference(s) type Feed Effluent (mg l-1) (g m-3d-1) (%) (mg l-1) (g m-3d-1) (%) Downflow AFRsa with HRTb = 216-408 h Mushroom 6.2 7.1 Zn 317 19-35 99.9 3000 36-68 20 Dvorak et al. limestone (Palmerton- T = 18-24 ºC compost Ni 0.9 0.05-0.10 97.8 1992 pilot, USA) Cd 0.3 0.02-0.03 99.0 Mn 26 1.5-2.8 98.1 Downflow AFR with HRT = 9.1-23.5 h Mushroom 2.1- 7.1-7.5 Fe 452-1081 1100-1190 99.8 1755- 3440- 88-90 Groudeva et limestone T = 20 ºC compost, 2.7 Cu 23-74 61-90 99.8-99.9 3740 4070 al. 1996 manure, As 10-19 16-26 99.8-99.9 straw, sawdust Upflow AFR with HRT = 1000-2000 h Ryegrass 2.3 6.5 Fe 2100 25-50 99.9 8100 60-120 20-40 Harris and alkaline soil materials T = 30-35 ºC Al 46 0.5-1 98.7 Ragusa 2001

Three sequential HRT = 120 h Mushroom 3.2 6.4 Al 7 1.4 100 1000 34 17 Dvorak et al. upflow AFRs with T = 10 ºC compost Fe 53 9.0 85 1992 limestone (Pittsburgh- pilot, USA) Two sequential upflow HRT = 57-113 h Straw 2-4 4.2-6.5 Fe 720-1480 NRc 15-90 2-3000 NR NR Estrada AFRs (one with As 95-230 90-100 Rendon et al. alkaline hydrated Pb 1-2 90-100 1999 calcium silicate) Zn 15-35 38-98 Mn 5-15 5-9 d Upflow AFR with pre- HRTtotal = 94 h Yeast 2.2 7.6-8.0 Fe 550 140 99.9 NR NR NR Hammack et d precipitation units (HRTbioreactor = 60 h) extract, beet Al 140 36 99.9 al. 1994a d using H2S, precipitator, molasses Cu 92 23 99.9 pH adjusted with Mn 76 19d 95.7 d NH4OH Zn 60 15 99.8 Co 4 1d 97.5 Ni 2 0.5d 95.0 aAFR=Anaerobic filter reactor, bHRT = hydraulic retention time, cNR = not reported, dRemoval rates calculated based on total HRT

58 Table 19, p.2/2. Continued. pH Metals Sulfate Rector and process Operation Substrate(s) Feed Removal Feed Removal Reference(s) type Feed Effluent (mg l-1) (g m-3d-1) (%) (mg l-1) (g m-3d-1)(%) a b Two mixed up-flow HRT bioreactor = 48 h H2 + CO2 , 3.1- NR Cu 12.1 129 99.9 1140- NR NR Rowley et al. bioreactors with two HRTchemical circuit = nutrients 4.5 Cd 0.085 0.80 89.2 1900 1997 pre-precipitation units 1.8-2.7 h Zn 16.2 171 99.0

using bioreactor Tbioreactor = 30 °C effluent and pH control (BioSulphide demonst- ration plant, Britannia Copper Mine, British Columbia, Canada) c d UASB with pre- HRTtotal = 206 h Lactate 2.3 6.5-8.1 Fe 620 21 76.9-99.9 NR NR NR Hammack et d precipitation using H2S (HRTbioreactor = 98 h) Cu 178 62 99.9 al. 1994b and limestone T = 24-30 ºC Zn 530 55-72d 99.9 neutralization Al 278 7-20d 99.9 Mn 191 32d 31.9-91.0

UASB with pre- Wastewater (=WW) H2+CO2, WW: WW: WW: NR WW: WW: NR WW: de Vegt et al precipitation units HRTWW: NR nutrients 2.5 8.5 Al 2200 99.9 30000 98 1998; (Thiopaq pilot, Leachwater (=LW) Cu 60 99.8 Boonstra et Kennecott Utah HRTLW: NR LW: LW: 2.2 Fe 675 99.9 al. 1999 Copper mine, USA) 2.6 Mg 4500 56.7 Mn 350 99.9 Zn 65 99.8 LW: LW: Cu 180 99.8 Fe 380 0.3 Zn 200 0.5 d d UASB with three pre- HRTtotal = 77 h Ethanol, 2.3 8.5 Fe 879 279 99.9 8150 1690 65 de Vegt et al d precipitation units (HRTbioreactor = 3.2 h) nutrients Al 24.4 7.7 99.6 1998; d using H2S gas and Tbioreactor = 26 °C Mn 150 47 98.8 Hammack bioreactor effluent Zn 523 166d 99.9 and Dijkman (pilot, Berkeley Pit, Cu 178 57d 99.9 1999 USA) Cd 1.96 0.62d 98.9 Co 1.51 0.47d 98.7 Ni 1.00 0.31d 98.0 aHRT = hydraulic retention time, bNR = not reported, cUASB = Upflow anaerobic sludge blanket reactor, dRemoval rates calculated based on total HRT.

59 10.1.2 The effect of hydraulic retention time on fluidized-bed reactor performance

The effect of HRT on the performance of fluidized-bed treatment of acidic Zn and Fe containing wastewater was studied by gradually decreasing the HRT from 20.7 to 6.1 h, while keeping the feed concentrations constant (Paper III). Decrease in HRT resulted in decreased percentual acetate utilization and sulfate reduction, along with reduced concentrations of dissolved sulfide and alkalinity in the FBR effluent (Figures 28 and 29). Ethanol utilization efficiency (%), as well as Zn and Fe precipitation remained high until process failures occurred at the lowest HRTs (Paper III). During the process failures caused by intentional overloading, sulfate reduction, substrate oxidation, metal precipitation, effluent sulfide, pH and alkalinity decreased (Paper III). After the performance failure the FBR process was recovered in a few days by adjusting the FBR liquid pH to 7 and interrupting the feed pumping for one day (Paper III). The FBR efficiently precipitated metals at HRT down to 6.5 h, but a further decrease in the HRT resulted in a process failure. At the HRT of 6.5 h and with an ethanol loading of 2.6 g l-1 d-1, Zn and Fe precipitation rates of over 600 mg Zn l-1 d-1 and 300 mg Fe l-1 d-1 were obtained. The corresponding metal removal efficiency was over 99.9 % and effluent Zn and Fe concentrations below 0.08 mg l-1. Under these conditions, ethanol oxidation was incomplete and acetate accumulated in the FBR. The operation of the FBR at shorter HRTs (at the corresponding loading rate) is limited by partial acetate oxidation, since excess alkalinity production by acetate oxidation is necessary for wastewater neutralization (Paper III).

a 5 b 100 Ethanol

) Acetate -1 4 80 ) d Sulfate -1 Zn 3 Fe 60

2 40 Ethanol Acetate Removal (% Sulfate Removal (g l (g Removal 1 20 Zn Fe 0 0 0 102030 01020 HRT (h) HRT (h)

Figure 28. The effects of hydraulic retention time (HRT) on a) quantitative and b) percent oxidation of ethanol and acetate, reduction of sulfate, and precipitation of Zn and Fe in the fluidized-bed reactor during the continuous flow experiment (Data obtained from paper III).

Unlike sulfidogenic ethanol oxidation (Reaction 13), lactate oxidation by SRB produces bicarbonate alkalinity in the first oxidation step, where lactate is converted to acetate (Reaction 22). In addition, complete lactate oxidation yields 3 moles of bicarbonate alkalinity per mole of substrate (Reactions 22 and 23), whereas ethanol oxidation yields only 2 moles (Reactions 13 and 14).

- 2- - - 2 CH3CHOHCOO + SO4 Æ 2 CH3COO + 2 HCO3 + H2S (22)

- 2- - - CH3COO + SO4 Æ 2 HCO3 + HS (23)

60 400 30 Sulfide 350 Alkalinity 25

) 300 pH -1 20 ) and pH

250 -1 200 15 150 10 Sulfide (mg l (mg Sulfide 100 5 50 Alkalinity (mmol l (mmol Alkalinity 0 0 0 5 10 15 20 25 HRT (h)

Figure 29. The effects of hydraulic retention time (HRT) on effluent dissolved sulfide, alkalinity and pH during the continuous flow fluidized-bed reactor (Data obtained from paper III).

Another substrate, the oxidation of which produces neutralizing capacity is hydrogen (Reaction 24) (van Houten et al. 1994).

2- - - 8 H2 + 2 SO4 Æ H2S + HS + 5 H2O + 3 OH (24)

In view of this it appears that lactate and hydrogen utilizing processes may be superior to ethanol-utilizing ones for treating acidic wastewaters because they are better at neutralizing the acidity in the wastewater influent. The drawback of lactate use is its high cost as compared to ethanol (Nagpal et al. 2000a). The production and use of synthesis gas (i.e. gas mixture of hydrogen, carbon monoxide and carbon dioxide) requires special infrastructure and specific reactor configurations and, therefore, this electron donor is considered economically feasible only in large-scale treatment (Boonstra et al. 1999).

In the present study, the maximum sulfate reduction rate (4.3 g l-1 d-1) obtained was at a HRT of 6.5 h. This is lower than the rates (6.33 g l-1 d-1) achieved by Nagpal et al. (2000b) using an ethanol-fed two-stage liquid-solid FBR inoculated with a defined culture of Desulfovibrio desulfuricans and Desulfobacter postgatei. Since the pH of the feed used by Nagpal et al. (2000b) was 6.3, the liquid-solid fluidized-bed operated well at HRTs down to 5.1 h, although percent sulfate reduction and acetate utilization decreased with decreasing HRT.

Kalyuzhnyi et al. (1997) studied the effect of HRT on sulfate reduction in a UASB reactor treating neutral-alkaline (pH 7-10) wastewater supplemented with excess of ethanol. In their study, percent sulfate reduction was not significantly affected by decreasing the HRT from 5 to 1 d, and decreased slightly when the HRT was further decreased to 8.4 h. Isa et al. (1986) studied the effect of HRT on an ethanol- and acetate-fed anaerobic digester with polyurethane sponges as the carrier matrix. They observed that an increase in the HRT from 0.5 to 10 d increased the percent electron flow towards the SRB and also the percentage of sulfate reduced. They concluded that the restricted competitive nature of the SRB was due to the wash-out of the SRB from the reactor and the preferential colonization of the carrier matrix by the methanogens (Isa et al. 1986). In our study using FBR treating acidic wastewater, percent sulfate reduction decreased when the HRT was decreased below 12 h (loading of 1.4 g ethanol l-1 d-1) (Figure 28b), whereas percentage of substrate utilized for sulfate reduction was not affected by the HRT (Paper III). Biogas was not produced in the experiments and, hence, methanogenesis remained insignificant. Moreover, the process failure due to the decreased HRT did not result in significant washout of biomass from the FBR. After the performance

61 failure, the process was recovered in a few days by adjusting the FBR liquid pH to 7 and interrupting the feed pumping for one day. In view of this, it appears that the decrease in percent sulfate reduction and acetate utilization was due to the low concentration of acetotrophic SRB and the preferential colonization of the carrier material by the SRB capable of only incomplete substrate oxidation.

10.1.3 Biomass yield

Previous studies have rarely reported biomass yields in continuous flow sulfate-reducing reactors. In this study, biomass yield was estimated in the reactor processes after enrichment with lactate (Paper I). Biomass yield in the FBR processes was 0.049-0.054 g dry biomass per g of lactate oxidized or 0.053-0.074 g dry biomass per g of sulfate reduced (Table 20). The corresponding values for the UASB reactor were 0.039 g dry biomass per g of lactate oxidized or 0.035 g dry biomass per g of sulfate reduced. The biomass yield per the amount of oxidized lactate in the continuous flow sulfate-reducing reactors was similar to those previously reported for pure cultures of SRB (Traore et al. 1982; Herrera et al. 1991) (Table 20), whereas the biomass yield per amount of reduced sulfate was an order of magnitude higher than that reported by Herrera et al. (1991). Part of the electron flow from lactate, especially during the early stages of continuous-flow enrichment, had other fates than sulfate reduction and, hence, the yields reported include also growth of other microbes in the system.

Table 20. Biomass yield in sulfate-reducing cultures. Yield (g dry Yield (g dry Culture biomass/ g lactate biomass/ g sulfate Reference oxidized) reduced) FBR1 0.054 0.074 Paper I FBR2 0.049 0.059 Paper I FBR3 0.052 0.053 Paper I UASB 0.039 0.035 Paper I a Desulfomicrobium norvegicum a 0.046 ± 0.007 NR Traore et al. 1982 Desulfovibrio gigas 0.042 ± 0.019 NR Traore et al. 1982 Desulfovibrio africanus 0.020 ± 0.001 NR Traore et al. 1982 Desulfovibrio desulfuricans strain 0.025 0.006 Herrera et al. 1991 NCIMB 9467 a Originally known as Desulfovibrio desulfuricans Norway 4 (Sharak et al. 1997), bNR= not reported

10.1.4 Composition and fate of metal precipitates

The metal precipitates of the water level adjustors of the FBRs and the UASB reactor were analyzed as XRD diagrams and SEM-EDS elemental analysis in order to verify their composition (Paper I). The metal precipitates were predominantly pyrite (FeS2), sphalerite (ZnS) and probably wurzite (ZnS) and iron pyrite (FeS), and contained insignificant amounts of oxide phases (Paper I). The precipitates from the water level adjustors also contained elemental sulfur, as a result of oxidation of the excess hydrogen sulfide in the water level adjustors. The Figure 30 shows a typical X-ray diffractogram of the metal precipitates and the Table 21 shows the corresponding main peak intensities of the precipitates and reference minerals.

62 1400

Pyrite (FeS2) Sphalerite (ZnS), cubic chrystal form 1200 Iron sulfide (FeS) Wurzite (ZnS), hexagonal chrystal form

1000 Elemental sulfur (α-S)

800 Counts 600

400

200

0 6 10 20 30 40 50 60 70 2θ (°) Figure 30. X-ray diffraction image of the metal precipitates obtained from the water level adjustor of a sulfate- reducing fluidized-bed reactor. A Cu-tube with the radiation wavelength of 1.54056 Å was used as the radiation source in the analysis.

Table 21. The main X-ray diffraction peak intensities of the metal precipitates obtained from the water level adjustor of a sulfate-reducing fluidized-bed reactor and the main peaks intensities of the reference minerals. The phase identification of the metal precipitates was performed using DIFFRACplus EVA software, version 7.0 (Bruker) with an attached ICDD Powder Diffraction File (PDF2) database. Peak intensities (%)a 2θ angle d value Metal Pyrite Sphalerite Iron sulfide Wurzite Sulfur (°) (Å) 0 precipitates (FeS2) (ZnS) (FeS) (ZnS) (S ) 11.7 7.53 11 - - - - 6 23.0 3.86 17 - - - - 100 25.8 3.45 11 - - - - 40 26.7 3.33 14 - - - 2 25 28.6 3.12 100 43 100 100 100 25 31.2 2.86 8 - - - - 18 32.9 2.72 30 100 10 6 4 - 37.0 2.43 16 58 - - - 14 40.7 2.22 12 43 - - - 2 47.5 1.91 58 42 51 55 50 8 56.3 1.63 41 74 30 35 35 6 59.1 1.56 8 11 2 2 2 2 61.5 1.51 6 11 - - - 2 64.2 1.45 6 13 - - - - 69.5 1.35 8 - 6 8 2 4 a The highest peak corresponds to 100 %

The fate of the metals was estimated during the continuous-flow loading experiment (Paper I). Most of the metal precipitates were retained in the reactors or water level adjustors, indicating that the metal precipitates can be easily separated from the effluent by e.g. gravitation (Paper I). The metal precipitates did not cause clogging of the reactors during the continuous operation. In a previous study, clogging has been reported in upflow anaerobic filter reactors (Steed et al. 2000).

In most studies concerning the treatment of metal containing wastewaters, the composition of metal precipitates has not been determined (e.g. Maree and Strydom 1985; Haas and

63 Polprasert 1993; Ma and Hua 1997; Farmer et al. 1995; Groudeva et al. 1996; de Lima et al. 1996; Estrada Rendon et al. 1999; Steed et al. 2000). In addition to sulfides, metals may precipitate e.g. as hydroxides or carbonates (Steed et al. 2000). Hydroxide and carbonate precipitation is of concern especially when the pH of the wastewater was 6.2-7.5 (Maree and Strydom 1985; Dvorak et al. 1992; Farmer et al. 1995), wastewater was treated with AFRs containing alkaline materials (Groudeva et al. 1996; Estrada Rendon et al. 1999) or when the pH of the treatment process was adjusted with e.g. NaOH, NH4OH or Na2CO3 containing buffers (Haas and Polprasert 1993; Steed et al. 2000). For example, Fe(OH)3 precipitation may be significant even at pH 3 (Hammack et al. 1994b). This study showed that the sulfidogenic FBR treatment allows the precipitation of metals as sufides, since pH adjustment with alkaline chemicals is not needed (Paper I). Sulfide precipitation is desirable, since most divalent metal sulfides are less soluble than the corresponding metal hydroxides (Table 6). In addition, the solubility of metal sulfides is not as sensitive to pH changes as that of metal hydroxides (Conner 1990). Metal sulfide sludges are relatively easy to dewater (Whang et al. 1982; Peters et al. 1985), and valuable metals can be recovered from the sludges (Boonstra et al. 1999).

10.1.5 Molar stoichiometry of the reactions in the fluidized-bed reactors

The molar stoichiometry of the reactions in the ethanol-oxidizing FBR was estimated from the continuous flow experiment with varying HRT (Paper III). The relationship of sulfate reduction and electron donor (ethanol plus its degradation intermediate, acetate) oxidation to sulfide production were as shown in Figure 31a and b, respectively. The sulfide production included dissolved sulfide and metallic sulfides (MS or MS2, where M is Zn or Fe). The extent of sulfide production calculated with the assumption that all metal sulfides were of the form MS2 followed the theoretical sulfide production expected from sulfate reduction rates (Figure 31a). When assuming that all metal sulfides were of the form MS, the sulfide production rate was close to theoretical at low sulfate reduction rates, but lower than theoretically assumed at high sulfate reduction rates. As seen in the XRD analysis of the metal precipitates (Paper I), metal sulfides of both MS and MS2 forms were produced in the FBRs. Therefore, the actual production of dissolved and metal sulfides was most likely between the two possibilities. Part of the H2S may have escaped into the headspace of the FBR during the experiments. However, no biogas was produced in the experiments. Additionally, the effluent pH was generally between 7.5 and 9 (Paper III), which is between the first and second pKa values of H2S (pKa1 = 7.04 and pKa2 = 11.96 at 25 °C) (Weast and Astle 1982). Therefore, - most of the dissolved sulfide was as HS rather than as H2S, and the volatilization of the H2S can be considered low. Another reason for the deviation from the theoretical sulfide production is that the production of dissolved sulfide was calculated from transient sulfide concentrations measured from the recycle flow whereas sulfate reduction rate was calculated from the the sulfate concentrations of 1 to 5 days integrated feed and effluent samples (Paper III).

The rate of sulfide production per oxidized electron donor was close to the theoretical rates with electron donor oxidation rates up to 50 mmol l-1 d-1 when assuming all metal sulfides to be of the form MS2 (Figure 31b). At higher electron donor oxidation rates or when assuming metal sulfides to be of the form MS, the sulfide production was lower than theoretically expected. The observed sulfide production deviated from the theoretical because part of the electron donors had other faiths than sulfate reduction, as shown in Figure 32.

64 a b

) 60 70 ) -1 -1 d

d 60

-1 50 -1 50 40 40 30 30 20 20 10 10 0 0 Sulfide production (mmol l 0204060Sulfide production (mmol l 050100 Sulfate reduction (mmol l-1 d-1) Electron donor oxidation (mmol l-1 d-1)

Figure 31. The relationship of (a) sulfate reduction and (b) electron donor oxidation (ethanol plus its degradation intermediate acetate) to the production of dissolved and metal sulfides in the ethanol-fed fluidized-bed reactor during continuous flow experiments (data obtained from Paper III). The data shown with the filled and unfilled circles were calculated assuming that all metal sulfides were of the form MS and MS2, respectively. The dashed line or the triangles show the theoretical sulfide production.

70 ) -1 60 d -1 50 40 30 20 10 0

Sulfate reduction (mmoll 0 50 100 Electron donor oxidation (mmol l-1 d-1)

Figure 32. The relationship of electron donor oxidation (ethanol plus its degradation intermediate acetate) to sulfate reduction in the ethanol-fed fluidized-bed reactor during continuous flow experiments (data obtained from Paper III). The triangles show the theoretical sulfate reduction.

As earlier discussed (Chapter 10.1.2), the alkalinity in sulfidogenic ethanol oxidation is produced by the oxidation of the degradation intermediate, acetate. The Figure 33 shows the relationship of the total and sulfidogenic acetate oxidation to the production of alkalinity. The sulfidogenic acetate oxidation was calculated using the percent reduction of sulfate and assuming that the % is the same for both ethanol and acetate oxidation. The production of alkalinity per total acetate oxidation is lower than the theoretical production of bicarbonate alkalinity (Reaction 23), which is due to the non-sulfidogenic part of the acetate oxidation (Figure 33a). The relationship of sulfidogenic acetate oxidation to production of alkalinity follows more closely the theoretical production of bicarbonate alkalinity (Figure 33b). The deviation from the theoretical can be explained by the other reactions producing or consuming protons, such as metal sulfide precipitation (Reaction 12) and oxidation of excess sulfide (Reaction 25) as well as the initial acidity of the wastewater (Wastewater pH was 3 and the end point of alkalinity titration at pH of 4.5).

65 - 0 - 2 HS + O2 Æ 2 S + 2 OH (25)

Part of the bicarbonate may have escaped as CO2 into the headspace according to the reaction 26.

+ - H + HCO3 ÅÆ H2CO3 Å Æ CO2 + H2O (26)

Some CO2 was observed in the gas phase of the FBRs during the enrichment with lactate, but no biogas accumulated into the gas bags during the experiments. Additionally, the effluent pH (7.5-9) was between the first and second pKa values of H2CO3 (pKa1 = 6.37 and pKa2 = 10.25 - at 25 °C) (Weast and Astle 1982). Therefore, HCO3 is expected to be the major form rather than H2CO3 or CO2, and the volatilization of the CO2 can be considered low. Based on the molar stoichiometry of the reactions, it seems that the extent of alkalinity production can to some extent be predicted on the basis of the sulfidogenic acetate oxidation.

a b 80 60 y

70 y 50 )

60 ) -1 -1 40 d

50 d -1 -1 40 30 30 20 (mmol l (mmol 20 l (mmol

Production of alkalinit 10 10 Production of alkalinit 0 0 0 10203040 0102030 -1 -1 Sulfidogenic acetate oxidation Acetate oxidation (mmol l d ) (mmol l-1 d-1)

Figure 33. The relationship of (a) total and (b) sulfidogenic acetate oxidation to the production of alkalinity (measured as the quantity of alkalinity in the treated effluent) in the ethanol-fed fluidized-bed reactor during continuous flow experiments (data obtained from Paper III). The dashed lines show the theoretical production of bicarbonate alkalinity from sulfidogenic oxidation of acetate.

10.2 Electron donor oxidation and sulfide inhibition kinetics

10.2.1 Ethanol and acetate oxidation kinetics

Ethanol and acetate oxidation kinetics were studied in batch FBR experiments. Acetate oxidation was the rate-limiting step in ethanol oxidation (Reactions 13 and 14), as demonstrated by the accumulation of acetate in the FBR during the batch experiments (Figure 34a). The effects of the initial substrate concentrations on the oxidation rates were as shown -1 -1 in Figure 34b. Michaelis-Menten constants (Km) were 4.3-7.1 mg l and 2.5-3.5 mg l for ethanol and acetate, respectively. The maximum oxidation velocities (Vmax) were 0.19-0.22 mg gVS-1 min-1 and 0.033-0.035 mg gVS-1 min-1, for ethanol and acetate, respectively. The results show that acetate was oxidized to CO2 considerably more slowly than ethanol was oxidized to acetate. The acetate oxidation step (Reaction 14) in ethanol oxidation produces the bicarbonate alkalinity, which neutralizes the acidic wastewater. In addition, acetate oxidation produces twice as much hydrogen sulfide per mole than does ethanol oxidation to acetate (Reaction 13).

66

a b

) 2.5 0.25 -1 ν (mg g VS-1 min-1) 2 0.2

1.5 0.15 Ethanol 1 0.1 Acetate

0.5 0.05

Ethanolor acetate (mmol l 0 0 0 10203040 050100 Time (min) S (mg l-1)

Figure 34. a) Ethanol concentration decrease (filled symbols) and acetate accumulation (unfilled symbols) in the fluidized-bed reactor during four batch experiments. b) The effect of initial substrate concentration (S) on ethanol and acetate oxidation rates (v) in the batch kinetic experiments (Paper II).

The Km and Vmax values for acetate oxidation were lower than previously reported for mesophilic sulfate-reducing cultures (Table 22). Most of the previous studies were performed as batch bottle assays and/or with pure cultures (Schönheit et al. 1982; Ingvorsen et al. 1984; Widdel 1988; Oude Elferink 1998). The batch bottle assays may not adequately represent the conditions in the FBR biofilm enrichment. The Km and Vmax values reported by Yoda et al. (1987) for an acetate-fed FBR were higher than the ones obtained in this study. The observed low Vmax in the present study indicates that only part of the total biomass in the FBR utilized acetate as an electron donor. Many of the known sulfate-reducers incompletely oxidize their substrates producing acetate as an end product (Castro et al. 2000).

Table 22. Kinetic parameters for acetate oxidation by mesophilic sulfate-reducing pure or enrichment cultures.

Strain/Culture T Km Vmax Reference (ºC) (mg l-1) (mg gVS-1 min-1) Desulfobacter postgatei 30 13.6 0.4-1.2a Schönheit et al. 1982 Desulfobacter postgatei 30 3.8-4.5 3.0-3.1 Ingvorsen et al. 1984 Desulfobacter postgatei 30 NRb 7.1-9.4 Widdel 1988 Desulforhabdus amnigenus 37 35 1.7a Oude Elferink 1998 Desulfobacca acetoxidans 37 35 2.5a Oude Elferink 1998 Enrichment culture 31 5.7 NR Middleton and Lawrence 1977 Mixed culture of SRB and 30 9.5 0.65 Yoda 1987 methanogens Enrichment culture 35 2.5-3.5 0.033-0.035 Paper II a Biomass determined as protein instead of VS, b NR = not reported.

Ethanol oxidation kinetics in sulfate-reducing systems has been previously reported by -1 Nagpal et al. (2000a). In their study, the Km for ethanol (0.2 g l ) was obtained using a batch- enriched sulfate reducing culture and batch bottle assays (Nagpal et al. 2000a) and was over one order of magnitude greater than that reported in the present study (4.3-7.1 mg l-1) (Paper II). The present study shows that the FBR system selectively enriches for sulfate reducers with high affinities for the electron donor. Based on the low Km values, the ethanol and acetate oxidation rates remain high even at low substrate concentrations. This implies low residual organic content in the effluent, and hence a potentially low environmental impact associated with the discharge of the effluent water. This was also shown as the low concentration of DOC in the effluent water of the FBR during the continuous flow operation at a HRT of 16 h.

67 In comparison, Nagpal et al. (2000b) did not report significant acetate consumption in the ethanol-fed sulfate-limited recirculating FBR containing Desulfobacter postgatei (complete oxidizer) and Desulfovibrio desulfuricans (incomplete oxidizer). Nagpal et al. (2000b) discussed that this could be due to the inhibition of acetate-consuming bacteria either by sulfide or acetate concentrations, or the better ability of D. desulfuricans to compete for the limiting amounts of sulfate (Laanbroek et al. 1984; Nagpal et al. 2000b).

10.2.2 Sulfide inhibition kinetics

In single-state SRB-processes the concentration of dissolved sulfide has to be maintained at relatively high levels to buffer the system against metal shock loads. However, dissolved sulfide is toxic to SRB and, therefore, sulfide product inhibition may be expected in single- stage processes (Kalyuzhnyi et al. 1997). Sulfide inhibition of ethanol and acetate oxidation was studied in the sulfate-reducing FBR with batch kinetic experiments (Paper III). The effects of initial DS and H2S concentrations on ethanol and acetate oxidation rates were as shown in Figure 35. The relationship between sulfide concentration and electron donor oxidation was non-linear and the oxidation was not completely inhibited at any concentration tested (Figure 35). The noncompetitive inhibition model described well the sulfide inhibition of the sulfate-reducing enrichment culture (Figure 35). DS inhibition constants (Ki) for ethanol and acetate oxidation were 248 mg S l-1 and 356 mg S l-1, respectively, and the -1 -1 corresponding Ki values for H2S were 84 mg S l and 124 mg S l , for ethanol and acetate, respectively (Paper III). During the continuous flow experiments with varying HRT, the DS -1 concentration in the FBR was 50-370 mg S l (Paper III). The H2S concentration remained below 50 mg S l-1 with the exception of process failure situations when the pH of the FBR dropped. At these time points, ethanol oxidation was most likely partially inhibited by the H2S (Paper III). Also during the continuous flow loading experiments with varying feed concentrations, over 400 mg l-1 DS concentrations were observed (Papers I and II). Excess dissolved sulfide causes chemical oxygen demand and, therefore, needs to be oxidized before discharging the effluent into the environment.

a 0.16 b 0.16 ) ) 0.14 0.14 -1 -1 0.12 0.12 min min

-1 0.1 -1 0.1 0.08 0.08 0.06 0.06

(mggVS 0.04 (mggVS ν 0.04 ν 0.02 0.02 0 0 0 500 1000 1500 0 200 400 600 -1 DS (mg S l ) -1 H2S (mg S l )

Figure 35. The effect of initial (a) dissolved sulfide (DS) and (b) hydrogen sulfide (H2S) concentration on ethanol (filled symbols) and acetate (unfilled symbols) oxidation rates (ν) in the batch kinetic experiments (paper III). The curves represent noncompetitive inhibition model fits.

An understanding of sulfide toxicity on populations of SRB is difficult to obtain from the literature, as the effect of pH has in many cases not been taken into account in experimental designs (O’Flaherty et al. 1999). Moreover, various studies have addressed the effect of sulfide either on SRB growth, electron donor oxidation or sulfate reduction, and these are not

68 always linearly correlated. For example, Visser et al. (1996) discovered that at high pH-levels, SRB growth is more strongly inhibited than the activity. The sensitivity of SRB to sulfide toxicity is also reported to depend on bacterial species (Hiligsmann et al. 1998; Maillacheruvu and Parkin 1996; O’Flaherty et al. 1998). Maillacheruvu and Parkin (1996) studied propionate oxidizing, acetate oxidizing and hydrogenotrophic SRB and reported that especially acetotrophic SRB were highly sensitive to H2S. Similarly, Yamaguchi et al. (1999) reported that acetate utilizers were more susceptible to sulfide inhibition than the hydrogen utilizers. In the present study, ethanol oxidation was more affected by sulfide toxicity than acetate oxidation.

Table 23 summarizes previous studies on the effects of sulfide on ethanol- and acetate- oxidizing SRB cultures along with the results obtained in this study. Most of the previous studies used serum bottle assays which do not represent the conditions in continuous flow reactors. Moreover, earlier studies with ethanol- and acetate-oxidizing SRB assessed the sulfide toxicity using granular or suspended sludge (Table 23). Visser et al. (1996) discovered that the nature of sulfide toxicity depends on the form of biomass and pH. They reported that at pH above 7 sulfide inhibition in granular sludge was caused by the total sulfide concentration, while in suspended sludge the undissociated H2S concentration determined the toxicity. The inhibition results obtained with granular or suspended sludge are probably not applicable to biofilm systems, such as FBRs. Okabe (1992) studied the kinetics and stoichiometry of sulfate reduction in suspended (chemostat) and biofilm (RotoTorque reactor) cultures, and reported that Desulfovibrio desulfuricans behave differently in biofilms than in suspension. Moreover, Teitzel and Parsek (2003) showed that biofilms of Pseudomonas aeruginosa were more resistant to copper and lead stress than planktonic cells. They suggested that the extracellular polymeric substances that encase a biofilm may be responsible for protecting cells from heavy metal stress by binding heavy metals and retarding their diffusion within the biofilm (Teitzel and Parsek 2003).

The overall effect of sulfide on the SRB growth and activity has been described qualitatively by many authors (Choi and Rim 1991; Hiligsmann et al. 1998; Kalyuzhnyi et al. 1997; Kolmert et al. 1997; McCartney and Oleszkiewicz 1991; McCartney and Oleszkiewicz 1993; O’Flaherty et al. 1998; Reis et al. 1992; Stucki et al. 1993; Visser et al. 1996; Yamaguchi et al. 1999). However, there are only few studies that quantitatively report the sulfide inhibition kinetics (Maillacheruvu and Parkin 1996; Okabe et al. 1995). The present study is the first to report quantitatively the toxic effect of sulfide on ethanol and acetate oxidation of a sulfate- reducing biofilm process. In this study, the sulfide inhibition constants for ethanol oxidation were lower than the sulfide concentrations that resulted in 50 % growth inhibition of Desulfococcus multivorans and of an anaerobic sludge (O’Flaherty et al. 1998) (Table 23). Sulfide inhibition constants for acetate utilization were significantly higher than those reported by Maillacheruvu and Parkin (1996) for a mixed SRB culture in a serum bottle assay but lower than sulfide concentrations that resulted in 50 % decrease in the acetate oxidation of UASB sludge in a serum bottle assay (Visser et al. 1996) (Table 23). The Ki values for acetate oxidation were also lower than the sulfide concentrations that resulted in 50 % growth inhibition of UASB sludge in a batch reactor assay (Visser et al. 1996) or of SRB pure cultures and anaerobic sludge in a serum bottle assay (O’Flaherty et al. 1998) or 50 % decrease in the sulfate reduction of UASB sludge in a serum bottle assay (Yamaguchi et al. 1999) (Table 23).

69 Table 23. The inhibitory effect of dissolved sulfide (DS) and undissociated H2S on the growth, sulfate reduction and electron donor oxidation of ethanol- and acetate-utilizing sulfate-reducing bacteria (SRB).

Substrate Biomass Vessel T (°C) pH DS (mg S/l) H2S (mg S/l) Inhibition Reference Ethanol Anaerobic sludge Serum 37 6.8 500-561 262-294 50%c O’Flaherty et al. 1998 bottle 7.2 788-880 241-269 7.6 878-990 131-147 8.0 1019-1130 66-74 8.5 1004-1164 22-25 Ethanol Desulfococcus multivorans Serum 37 6.8 498 261 50%c O’Flaherty et al. 1998 bottle 7.2 851 250 7.6 1383 206 8.0 1488 97 8.5 1570 34 Ethanol UASBa sludge UASB 35 7.1-8 1700-9951 570-610 100%d Kalyuzhnyi et al. 1997 b e Ethanol Mixed SRB culture in FBR FBR 35 6.9-7.3 248 84 Ki Paper III Acetate UASB sludge Batch 30 7 521 217 50%c Visser et al. 1996 reactor 7.5 569 107 8 921 64 9 943 7 Acetate Desulfonema magnum, Serum 30/37 6.8 443-583 233-306 50%c O’Flaherty et al. 1998 Desulfotomaculum acetoxidans bottle 7.2 671-926 205-283 or Desulfobacter postgatei 7.6 660-1360 98-203 8.0 659-1500 43-98 8.5 708-1450 15-31 Acetate Anaerobic sludge Serum 37 6.8 374 196 50%c O’Flaherty et al. 1998 bottle 7.2 550 168 7.6 867 129 8.0 990 65 8.5 1011 22 Acetate UASB sludge Serum 35 7 699 270 50%d Yamaguchi et al. 1999 bottle Acetate UASB sludge Serum - 7.2-7.4 615 161 50%e Visser et al. 1996 bottle 8.1-8.3 1125 54 e Acetate Mixed SRB culture Serum - - 35 8 Ki Maillacheruvu and Parkin. 1996 bottle e Acetate Mixed SRB culture in FBR FBR 35 6.9-7.3 356 124 Ki Paper III aUASB= upflow anaerobic sludge blanket reactor, bFBR= fluidized-bed reactor, cInhibition of growth, dInhibition of sulfate reduction, eInhibition of electron donor oxidation.

70 10.3 Microbiology of the fluidized-bed reactors

Sulfate-reducing bioreactor communities used for AMD treatment have not been previously extensively described. In this study, bacterial communities of the long-term (500-980 d) maintained FBRs were characterized using culture-independent molecular methods and phenotypic methods as well as bacterial isolations.

10.3.1 Clone libraries

Clone libraries constructed using the PCR-amplified 16S rRNA genes of the FBR communities resulted in 170 and 130 clones, for the lactate- and ethanol-utilizing FBRs respectively (Paper IV). One hundred clones from each of the clone libraries were selected randomly for analysis. Seventeen different operational taxonomic units (OTU) were found in the ethanol-FBR library and 13 OTUs in the lactate-FBR library (Table 24). Additionally, 11 and 8 sequences from ethanol- and lactate-FBRs respectively, were recognized as chimeric artefacts (Kopczynski et al. 1994). The distribution of the OTUs among the different phylogenetic groups was as shown in Table 24. The clone library results showed a significant difference in the OTU distribution between ethanol- and lactate-utilizing FBRs. Most notable was the large number of Proteobacterium sequences retrieved from the ethanol-fed FBR, whereas in the lactate-fed FBR, sequences clustering with Nitrospira phylum were abundant. The dominant lactate-FBR clone type that was affiliated with the Nitrospira phylum was identical to a previously described 16S rRNA gene clone TCE65 (AF349764), which was obtained from a TCE reducing enrichment culture. The dominant clone was also distantly related (86 % similarity) to Magnetobacterium bavaricum (MB16SRRN), a magnetotactic bacterium, which is yet to be grown in axenic culture. M. bavaricum has been observed in freshwater sediments and its proposed electron acceptors include oxygen, nitrate, sulfate or iron oxides (Spring et al. 1993). The closest related cultivated species is Thermodesulfovibrio sp. TGL-LS1 (AB021303), a thermophilic sulfate-reducing bacterium, sharing a sequence similarity of 85 % with the dominant OTU (Paper IV).

Within the phylum Proteobacteria, clone OTUs were found in beta (β), gamma (γ), delta (δ), and epsilon (ε) subdivisions (Paper IV). The most abundant clone type from the ethanol- utilizing FBR was affiliated with δ-Proteobacteria and was 96 % similar to a clone SJA-152 (UEAJ9496), which was previously obtained from an anaerobic trichlorobenzene- transforming microbial consortium (von Wintzingerode et al. 1999). The closest cultured relative to the dominant clone was Geobacter pelophilus (U96918) (94 % similarity). One of the defining physiological characteristics of Geobacter species is the oxidation of acetate coupled to the reduction of S0 and Fe3+ (Coates et al. 1996). A total of seven clones from both reactors were 99 % similar to Desulforhabdus amnigenus (DA16SRR) and one lactate-FBR clone was 99 % similar to Desulfobacca acetoxidans (AF002671). Desulfobacca and Desulforhabdus are gram-negative SRB, which oxidize their substrates completely to CO2 (Oude Elferink et al. 1995; Oude Elferink et al. 1999). Six clones from the ethanol-utilizing FBR affiliated with Desulfovibrio cluster, the closest cultured relatives of the three OTUs being Desulfovibrio sulfodismutans (DSU17764), Desulfovibrio alcoholovorans (AF053751), and Desulfovibrio sp. Zt10e (AF109469) with 97-99 % similarity (Paper IV). The members of the genus Desulfovibrio are gram-negative sulfate-reducers, most of which oxidize their substrates incompletely to acetate (Macfarlane and Gibson 1991).

71 Table 24. Distribution of the 16S rDNA clones and operational taxonomic units (OTU) obtained from ethanol- and lactate-utilizing fluidized-bed reactors (FBR) with clone library, denaturing gradient gel electrophoresis (DGGE) and isolation methods. Ethanol-fed FBR Lactate-fed FBR Clone library Clone library Phylogenetic group % of DGGE Isolate % of DGGE Isolate clones OTUs OTUs OTUs clones OTUs OTUs OTUs (n = 88) (n = 91) β-Proteobacteria 1.1 1 2 - 1.1 1 - γ-Proteobacteria 1.1 1 - - - - - δ-Proteobacteria 56.8 6 6 2 9.9 3 3 3 ε-Proteobacteria 2.3 1 1 - 4.4 2 - Chloroflexi 3.4 1 1 - 15.4 2 - Candidate division OP5 9.1 1 - - - - - Chlorobi - - 2 - 4.4 2 1 Cytophaga-Flexibacter- 20.5 3 2 1 4.4 1 2 1 Bacteroides Firmicutes - - - 11 - - - 2 Spirochaetales 5.7 3 1 - 1.1 1 1 Nitrospira - - - - 59.3 1 1 Undefined - - 1 - - - - Total 100 17 16 14 100 13 8 6

10.3.2 Denaturing gradient gel electrophoresis

DGGE analysis was applied to reveal successive changes in the microbial communities of the FBRs (Paper IV). Replication of the PCR and DGGE steps confirmed the consistency of the profiles. In the DGGE analysis, 16 and 8 different sequences were retrieved from the ethanol- and lactate-FBRs, respectively (Table 24). The DGGE genotypes clustered with β-, δ- and ε- Proteobacteria, Nitrospira, Cytophaga-Flexibacter-Bacteroides (CFB), Spirochaetales, Chlorobi and Chloroflexi groups. Many of the DGGE sequences were similar to those observed in the clone libraries. Additionally some sequences clustering with β-, δ- and ε- Proteobacteria, Chlorobi group and CFB phylum were found that were not detected in the clone libraries. The sequences representing the dominant clone types in the clone library were detected also with the DGGE method. Sequences related to Desulforhabdus amnigenus and sequences clustering with CFB group were found from both FBRs over the 393-day study period, whereas sequences similar to Desulfobacca acetoxidans, Desulfovibrio alcoholovorans, D. vulgaris, and D. sulfodismutans were detected only in some samples (Paper IV). Factors such as variation in DNA isolation and cloning efficiency (Oude Elferink et al. 1998a), PCR selection, drift (Wagner et al. 1994) and chimeric artefacts (Kopczynski et al. 1994; Wang and Wang 1996) can create biases in the apparent community composition and diversity determined by molecular studies.

10.3.3 Bacterial isolates

A total of 128 strains were screened microscopically following isolation using anaerobic plates and roll tubes (Paper V). A selection of 55 strains was characterized by partial sequencing of the 16S rRNA gene and 17 distinct phylogenetic types were identified and characterized further (Figure 36; Table 25). According to the 16S rRNA gene sequence data, six and fourteen different strains were isolated from the lactate- and ethanol-utilizing FBRs, respectively (Tables 24 and 25). Three of these strains from the lactate-FBR were identical to strains isolated from the ethanol-FBR (Table 25). Twelve of the 17 different strains were 72 mesophilic (growth at 35 ºC), and five were thermophilic (growth at 50 ºC). Three strains had vibrioid morphology, one had an oval cell shape, one had needle-like cells and the other twelve were rods (Figure 36; Table 25). Seven strains were dissimilatory sulfate reducers (SRB), whereas the other ten had a sulfite reducing, fermentative or acetogenic metabolism (Table 25). All of the SRB used ethanol as a source of carbon and energy, and all SRB except for strain RE35E1 utilized lactate (Table 25). Only two of the non-sulfate-reducing strains utilized lactate and ethanol, and one strain used lactate and ethanol only when sulfate was replaced with 10 mM sulfite in the Pfennig medium. The other non-sulfate reducing strains most likely grew on the yeast extract in the Pfennig medium, indicating that their role in the FBRs was probably to degrade complex organic constituents produced by the biofilm, presumably as extracellular exudates, polymers or as decay products of dead cells. None of the isolates utilized acetate as the sole source of carbon and energy although two of the strains were isolated using growth medium containing acetate as the sole carbon source (Paper V).

Analysis of the 16S rRNA gene (1017-1606 bp sequenced) of the isolates revealed that four mesophilic strains belonged to the δ-Proteobacteria, twelve to the Firmicutes and one to the Bacteroidetes (Tables 24 and 25) (Paper V). The sulfate-reducing strains were affiliated with Desulfovibrio vulgaris (99.6 % sequence similarity), Desulfovibrio alcoholovorans (99.7 %), Desulfovibrio desulfuricans (99.5 %), Desulfobulbus propionicus (99.7 %), Desulfotomaculum nigrificans (99.4 %), Desulfotomaculum thermocisternum (88.7 %) and, one with both Desulfotomaculum sapomandens and Desulfotomaculum thermosapovorans (93.7 %) (Table 25). The other strains were affiliated with members of the genera Desulfitobacterium, Clostridium, Caloramator, Oxobacter and Bacteroides (Table 25). Many of the isolated strains were only distantly related to previously described species and, thus, may represent novel species or genera (Paper V).

The traditional anaerobic microbiology techniques did not result in pure cultures of the dominant genotypes in the clone library and DGGE analysis, which were related to members of the genera Geobacter and Magnetobacterium (Papers IV, V). This relates to the difficulty of obtaining pure cultures of anaerobes or the inability of the organisms to grow on the enrichment or isolation medium used. It highlights the draw-backs of using culture work to describe bacterial communities, especially anaerobes. Because of these difficulties, the physiology and taxonomy of the dominant members of the FBR communities, as identified by the molecular analyses (Paper IV), remain undescribed. Some bacterial species (Desulfovibrio vulgaris, D. alcoholovorans, and one strain distantly related to genus Bacteroides) obtained as pure cultures (Paper V) correspond to the 16S rRNA gene sequences retrieved from the FBRs with molecular methods (Paper IV). Additionally, several species (including Desulfobulbus propionicus, Desulfovibrio desulfuricans, and many gram positives) were isolated (Paper V) that were not detected in the FBR using clone library or DGGE methods (Paper IV). Many of the strains isolated at 50 ºC or after the initial enrichment in liquid culture are probably less dominant organisms in the FBRs, since the optimal temperature for the thermophilic strains is higher than the operating temperature of the FBRs, and most of the fermentative or acetogenic strains did not utilize ethanol or lactate. Growth medium and temperature are known to select for fast-growing organisms that are best adapted to the cultivation conditions (Bull et al. 2000), whereas less common organisms are rarely detected by standard rRNA approaches due to methodological limitations (Amann et al. 1995). The combination of molecular and culturing methods allowed the detection of more species in the sulfate-reducing FBR communities than either approach alone (Papers IV and V). Therefore, comparative studies using both molecular and culturing studies are needed for obtaining a more comprehensive view of the diversity of complex microbial communities (Bull et al. 2000).

73

Figure 36. Phase contrast photomicrographs of strains isolated from the sulfate-reducing fluidized-bed reactors. Strains: (a) PL35L1; (b) PA35E4; (c) RPf35E1; (d) RPf35L17; (e) RE35E1 (f) RL50L1; (g) RA50E1; (h) RPf35Ei; (i) PPf35E1; (j) PPf35E6; (k) PPf35E8; (l) PPf35E10; (m) PPf35L1; (n) PPf50E1; (o) PPf50E2; (p) PPf35E4; (q) PPf50E4. All panels are at the same magnification.

74 Table 25. Characteristics of strains isolated from the sulfate-reducing fluidized-bed reactors (FBR). Strain Isolation Growth medium T Source- Closest relative(s) Cell form Spores Sulfate Oxidationc method (ºC) FBRa (% 16S rRNA sequence similarity)b reduction Ethanol Lactate PL35L1 Plated Minimal Postgate (L)e, f 35 L Desulfovibrio vulgaris (99.6) vibrioid - + + + PA35E4 Plated Minimal Postgate (A)g 35 E Desulfovibrio alcoholovorans (99.7) vibrioid - + + + h RPf35E1 Roll tube Pfennig without Na2S 35 E (L) Desulfovibrio desulfuricans (99.5) vibrioid - + + + RPf35L17 Roll tube Pfennig without Na2S 35 L Desulfobulbus propionicus (99.7) oval - + + + RE35E1 Roll tube Rich Postgate (E)i 35 E Desulfotomaculum sapomandens and rod + + + - Desulfotomaculum thermosapovorans (93.7) RL50L1 Roll tubej Rich Postgate (L)k 50 L (E)l Desulfotomaculum nigrificans (99.4) rod + + + + RA50E1 Roll tubej Rich Postgate (A) 50 E Desulfotomaculum thermocisternum (88.7) rod + + + + m m RPf35Ei Roll tube Pfennig without Na2S 35 E Desulfitobacterium dichloroeliminans (95.0) rod + - + + PPf35E1 Platej Pfennig 35 E Clostridium sartagoformum (99.9) rod + - - - PPf35E4 Platej Pfennig 35 E Clostridium subterminale (97.3) rod + - - - PPf35E6 Platej Pfennig 35 E Clostridium aminobutyricum (96.6) rod + - - - PPf35E8 Platej Pfennig 35 E Clostridium tetanomorphum, Clostridium rod + - - - pascui and Clostridium peptidovorans (94.3) PPf35E10 Platej Pfennig 35 E Clostridium pascui (94.0) rod + - - - PPf35L1 Platej Pfennig 35 L Clostridium glycolicum (99.5) rod + - + + PPf50E1 Platej Pfennig 50 E Caloramator coolhasii (99.7) rod + - - - PPf50E2 Platej Pfennig 50 E (L) Bacteroides distasonis (83.0) needle - - - - PPf50E4 Platej Pfennig 50 E Oxobacter pfennigii (91.8) rod + - - - a E = ethanol-FBR, L = lactate-FBR. b At least 1074 bp were sequenced for each strain. c None of the strains oxidized acetate. d A similar strain was also isolated using roll tubes. e Electron donor in the Postgate medium given in the parenthesis (A=acetate, E=ethanol, L=lactate). f Similar strains were also isolated using Minimal Postgate g medium containing acetate, and Pfennig medium. Similar strains were also isolated using H2 + CO2 atmostphere and Minimal Postgate medium with and without acetate, Rich Postgate medium containing ethanol, and Pfennig medium. h A similar strain was also isolated from the lactate-FBR. i A similar strain was also isolated using Rich Postgate medium containing acetate. j Roll tubes or plates were prepared after enrichment in liquid culture. k A similar strain was also isolated using Rich Postgate medium containing ethanol. l A similar strain was also isolated from the ethanol-FBR. m In the presence of 10 mM sulfite.

75 10.3.4 Phospholipid fatty acid profiling

Previous studies have shown that the cells of SRB contain some distinct fatty acids that have value as biomarkers in environmental studies (Ueki and Suto 1979; Taylor and Parkes 1983; Dowling et al. 1988; Vestal and White 1989; Řezanka et al. 1990; Vainshtein et al. 1992; Kohring et al. 1994; Green and Scow 2000). In the present study, the FBR communities were characterized by PLFA profiling (Paper V). Total of 34 and 27 individual PLFAs in the range of C14-C18 were identified from the ethanol- and lactate-FBRs, respectively (Figure 37). The FBR community members contained large amounts (80-87 %) of saturated fatty acids. The proportions of branched PLFAs were 67 % and 58 % in ethanol- and lactate-FBR, respectively. The most abundant (12-19 %) PLFAs in both FBRs were i15:0 and 16:0 (Paper V). Significant amounts of i15:0 have previously been found in Desulfovibrio desulfuricans, D. vulgaris, D. baculatus, D. simplex, D. termitidis, Desulfomonas pigra (Vainhstein et al. 1992) and Desulfotomaculum nigrificans (Ueki and Suto 1979), but also some non-sulfate- reducers, such as Geobacter metallireducens (Lovley et al. 1993) and Cytophaga sp. (Haack et al. 1994). Palmitic acid (16:0) is present in many different species of bacteria (Rajendran et al. 1997). Considerable amounts (5-10 %) of a15:0, predominant in Desulfovibrio alcoholovorans, D. carbinolicus, D. fructosovorans, D. giganteus, D. gigas, D. sulfodismutans (Vainhstein et al. 1992) and Desulfotomaculum nigrificans (Ueki and Suto 1979) were detected in both FBRs (Paper V). Lactate-FBR community members contained higher amounts of 15:0 compared to the ethanol-FBR. Significant amounts of 15:0 were previously found in Desulfobulbus sp. (Taylor and Parkes 1983). A proposed biomarker for Desulfobacter sp., 10me16:0 (Dowling et al. 1986) was also detected in the FBR samples (Paper V). Fatty acid 16:1ω7c, which accounted for 3.5-3.6 % of the total PLFAs in the FBRs (Paper V), was reported to be a major fatty acid in G. metallireducens (Lovley et al. 1993), but is also present in some SRB including Desulfomonile tiedjei (Ringelberg et al. 1994). No cy15:1, a proposed biomarker PLFA for Clostridia (Parkes 1987), was detected in the FBR samples (Paper V).

The PLFA analysis did show signatures representing taxa that were members of the communities in both FBRs (Papers IV, V). The exactness of the interpretation was hindered by the similarity of dominant fatty acids in the profiles of different taxa, the lack of unique signature fatty acids in significant proportions in the profiles of many taxa, and taxonomic variation in the fatty acid yields (Haack et al. 1994). New additional biomarkers are needed to distinguish the major groups observed in the molecular analysis (Paper IV). Although the application of PLFA profiling to environmental samples is free of culturing biases, the information on the occurrence of individual markers in defined groups of microorganisms can only be derived from the analysis of isolated pure cultures (Spring et al. 2000). The PLFA fingerprints from both FBRs were remarkably similar (Figure 37) (Paper V) despite the fact that the molecular methods showed quite different communities (Paper IV). Obviously, the resolution of the PLFA approach is not high enough to distinguish most of the individual bacterial species in complex communities (Spring et al. 2000). The outcome highlights some drawbacks of interpretation of phospholipid data and its limited use. However, the PLFA profiling is free of PCR biases and provides rapidly information on the active microbial community.

76 25 %-ratio Ethanol-FBR 20 Lactate-FBR 15

10

5

0 14:0 15:0 16:0 18:0 17:0 i14:0 i15:0 i16:1 i16:0 i18:1 i18:0 i17:0 a15:0 a17:0 br15:0 br17:0 br17:0 a18:0* cy19:0 cy17:0 18:1w7 18:1w6 18:1w5 i17:1w5 16:1w7c 17:1w5c i17:1w10 16:0 3OH 16:0 2OH 18:0 10me15:0 10me16:0 i17:0 2OH i18:0 2OH a17:0 2OH 16:1w7t/6c Phospholipid fatty acids

Figure 37. The percentages of phospholipid fatty acids (PLFA) in order of increasing retention time in the carrier material of ethanol- and lactate-utilizing fluidized-bed reactors (FBR) (n = 2 samples). *Includes small amounts of i16:0 3OH, i16:0 2OH and 18:1. The absolute concentrations of PLFAs corresponding to 100 % were 17.2 ± 0.4 and 6.4 ± 3.3 µg per g of FBR carrier material in the ethanol and lactate-FBRs, respectively.

10.3.4.1 Community diversity

Based on this study the FBR communities, originally enriched from methanogenic granular sludge and mine sediments, contained SRB related to the members of the genera Desulfovibrio, Desulfobulbus, Desulfotomaculum, Desulfobacca and Desulforhabdus (Papers IV and V). The SRB communities used for AMD-reactor applications had not been extensively described prior to this study. Desulfotomaculum, Desulfovibrio and Desulfobulbus species have been previously detected in mine environments (Chang et al. 2001; Nakagawa et al. 2002) and various bioreactors treating organic waste materials (Table 26) (Okabe et al. 1999; Poulsen et al. 1993; Raskin et al. 1995; Raskin et al. 1996; Oude Elferink et al. 1998b; Sekiguchi et al. 1998; Santegoeds et al. 1999; Plugge et al. 2002). The type species of the genera Desulforhabdus and Desulfobacca were isolated from UASB reactors (Oude Elferink et al. 1995; Oude Elferink et al. 1999).

Previous studies report low species diversity in open ecosystems as a result of selection pressures, such as particular electron donors (Tiirola et al. 2002), electron acceptors (von Canstein et al. 2002) or inhibitors (Nyman 1999; Sandaa et al. 1999). The present study demonstrated that sulfate-reducing FBR-communities enriched and maintained on single substrates (ethanol and lactate) and treating acidic metal-containing wastewater comprised a diverse mixture of bacteria including many previously undescribed species. Species diversity has been previously reported to improve the mercury retention efficiency of a packed-bed bioreactor under rapidly increasing mercury concentrations (von Canstein et al. 2002). In addition, Whiteley et al. (2001) showed that resistant organisms contribute to the overall biofilm disinfectant resistance in a glucose-limited chemostat. Thus, microbial diversity provides a reservoir of strains with complementary ecological niches that improves the stability of bioreactor performance under changing environmental conditions (von Canstein et al. 2002). Fernandez et al. (1999; 2000) studied the community structure of methanogenic bioreactor communities fed with glucose, and reported that stability is linked to community flexibility reflected in the ability to shift the electron and carbon flow through various alternative guilds in an efficient manner (Fernandez et al. 2000). Moreover, Fernandez et al. (1999) concluded that an extremely dynamic community can be developed in a simple, functionally stable ecosystem. In sulfate-reducing reactors for mining applications, diversity and flexibility of microbial community may enhance biofilm robustness under varying

77 operational conditions, such as metal and acidity shocks, and the maintenance of anaerobic conditions. In this study, the FBR-communities contained also many thermophilic strains with optimal growth temperatures higher than the operating temperature of the FBRs. Moreover, the biofilm harbored many non-sulfate reducing bacteria which probably degraded complex organic constituents produced by the biofilm, such as extracellular exudates, polymers and the decay products of dead cells. The degradation products (e.g. carboxylic acids, ethanol, H2 and CO2) of these complex organic materials may again be utilized as energy and carbon sources for SRB (Widdel and Bak 1992; Widdel and Hansen 1992). This study showed that sulfate- reducing FBRs fed with simple electron donors only, can contain bacterial communities of large diversity.

Table 26. Examples of sulfate-reducing bacteria detected in various wastewater-treatment systems and mine environments. Wastewater treatment systems Environmenta Mine sites F F R B B B G G G G G G A Desulfobacca + + Desulfobacter + + Desulfobacterium + + Desulfobulbus + + + + + + + + Desulfococcus/ + + + + Desulfosarcina/ Desulfobotulus Desulfotalea-Desulfofustis + Desulforhabdus + + Desulfotomaculum + + + + Desulfovibrio + + + + + + + + + + Thermodesulfobacterium + Thermodesulforhabdus + Thermodesulfovibrio + + Referenceb 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 aEnvironments: F = fluidized-bed reactor, R = rotating disk reactor, B = anaerobic filter reactor, G = upflow anaerobic sludge blanket reactor, A = activated sludge. bReferences: 1 = Paper V, 2 = Paper IV, 3 = Okabe et al. 1999, 4 = Poulsen et al. 1993, 5 = Raskin et al. 1995, 6 = Raskin et al. 1996, 7 = Oude Elferink et al. 1998b, 8 = Sekiguchi et al. 1998, 9 = Santegoeds et al. 1999, 10 = Oude Elferink et al. 1995, 11 = Oude Elferink et al. 1999, 12 = Plugge et al. 2002, 13 = Santegoeds et al. 1998, 14 = Chang et al. 2001, 15 = Nakagawa et al. 2002, 16 = Karnachuk et al. submitted.

78 11 CONCLUSIONS

This study demonstrated the feasibility of sulfidogenic FBR for concomitant removal of acidity, metals and sulfate from wastewaters. Process design parameters for the FBR process were determined. The study also contributed to the understanding of the wide microbial diversity in the sulfate-reducing FBRs fed with simple electron donors only. Further, several bacterial strains possibly representing novel species or genera were isolated and characterized. Based on the study, the following conclusions and can be drawn:

• Low fluidization rate (10 %) in the FBR enables faster biofilm formation during start-up, whereas higher fluidization (20-30 %), i.e. recycling rate is advantageous at increased loading rates (Paper I). • The gradual change of the electron donor from lactate to ethanol does not significantly affect the FBR performance at a HRT of 16 h (Paper II). • At a HRT of 16 h, the FBR is able to precipitate over 350 mg Zn l-1 d-1 and 80 mg Fe l-1 d-1 from acidic water containing 230-240 mg Zn l-1 and 57-58 mg Fe l-1. The corresponding percentual metal removal is over 99.8 % and effluent soluble Zn and Fe concentrations below 0.1 mg l-1. The alkalinity produced in lactate and ethanol oxidation increases the wastewater pH from 2.5 to 7.5-8.5 (Papers I and II). • A decrease in HRT results in decreased acetate oxidation, which is accompanied with reduced concentrations of dissolved sulfide and alkalinity in the FBR effluent. The FBR efficiently precipitates metals at HRTs down to 6.5 h, but a further decrease in HRT results in a process failure. At the HRT of 6.5 h, Zn and Fe precipitation rates of over 600 mg Zn l-1 d-1 and 300 mg Fe l-1 d-1 can be obtained from wastewater containing 176 mg Zn l-1 and 87 mg Fe l-1. The corresponding percentual metal removal is over 99.9 % and effluent soluble Zn and Fe concentrations below 0.08 mg l-1. The alkalinity produced in the ethanol oxidation increases the wastewater pH from 3 to 7.9-8.0 (Paper III). • After performance failure due to intentional overloading the FBR process can be recovered in a few days by adjusting the FBR liquid pH to 7 and interrupting the feed pumping for one day (Paper III). • With the FBR enrichment and process operation approach presented in this study, biomass yield is about 4.9-5.4 % (g/g lactate oxidized) or 5.3-7.4 % (g/g sulfate reduced) (Paper I). • Zinc and iron precipitate in the sulfidogenic FBR as sulfides. Most of the metal precipitates are retained in the FBRs or water level adjustors, indicating that the metal precipitates can be easily separated from the effluent by e.g. gravitation (Paper I). • Michaelis-Menten constants (Km), determined in batch FBR experiments, are 4.3- 7.1 mg l-1 and 2.5-3.5 mg l-1 for ethanol and acetate oxidation, respectively. The -1 -1 maximum oxidation velocities (Vmax) are 0.19-0.22 mg gVS min and 0.033- 0.035 mg gVS-1 min-1, for ethanol and acetate, respectively. Acetate oxidation is the rate limiting step in sulfidogenic ethanol oxidation (Paper II). • Dissolved sulfide inhibition constants (Ki) for ethanol and acetate oxidation are 248 -1 -1 mg S l and 356 mg S l , respectively, and the corresponding Ki values for H2S are 84 mg S l-1 and 124 mg S l-1. Therefore, ethanol oxidation is more inhibited by sulfide toxicity than the acetate oxidation (Paper III). • Lactate and ethanol support diverse sulfate-reducing communities in FBRs treating acidic metal- and sulfate containing wastewater (Paper IV).

79 • The FBR communities, originally enriched from methanogenic granular sludge and mine sediments, contain SRB related to the members of the genera Desulfovibrio, Desulfobulbus, Desulfotomaculum, Desulfobacca and Desulforhabdus, and also many species that do not reduce sulfate (Papers IV and V). • The FBR communities contain many previously undescribed bacteria (Papers IV and V). • The clone library method showed significant difference in the communities of the lactate- and ethanol utilizing FBRs, and DGGE profiling some changes in the communities over time (Paper IV). • The PLFA profiling showed some signatures consistent with sulfate-reducing communities (Paper V). • The PLFA analysis was unable to show any substantial difference in the microbial communities in the reactors (Paper V). • Many of the strains isolated from the FBRs are only distantly related to previously described species and, thus, may represent novel species or genera (Paper V). • In this work, the combination of molecular and culturing methods allowed the detection of more species in the sulfate-reducing FBR communities than either approach alone (Papers IV and V). Therefore, comparative studies using both molecular and culturing studies are needed for obtaining a more comprehensive view of the diversity of complex microbial communities.

80 12 RECOMMENDATIONS FOR FUTURE RESEARCH

In the present work, the focus of the FBR process performance studies was on the optimization of metal sulfide precipitation and neutralization of the wastewater acidity, whereas effluent sulfate, sulfide and organic carbon concentrations were not minimized. Excess sulfide and organic carbon induce oxygen demand in the receiving water bodies. Moreover, released sulfate induces sulfidogenic activity in the anaerobic sediments of the receiving water systems. The production of H2S by the indigenous SRB in the sediments consumes the oxygen from the body of water and may even result in fish kills (Müezzinoğlu et al. 2000). To minimize the oxygen demand of the effluent and to avoid the potential problems with malodor or health concerns caused by the H2S, a unit process for the oxidation of excess sulfide should be developed and incorporated downstream of the FBR. Additionally, future research is needed to find ways to further minimize the residual sulfate and organic carbon concentrations in the treated effluent.

The type of the biomass carrier affects the biomass concentration and achievable loading rates of the FBR (Yee et al. 1992; Leenen et al. 1996). Silicate mineral was used as the biomass carrier in the present study. In the future work, the suitability of other carrier materials should be tested to maximize the biomass retainment and loading rates. The electron donors used in this study included lactate and ethanol. Alternative electron and carbon sources may decrease the operating costs of the FBR process. The synthetic wastewater used in this study contained zinc and iron. Before the scale-up and cost-analysis of the FBR process, the FBR treatment performance needs to be tested with real waste- or process waters.

In the present study, the FBRs were operated at 35 °C, which is a suitable growth temperature for the mesophilic SRB. More research is necessary to find the optimal temperature for the enrichment culture. For a treatment process operating at the low temperatures of the boreal regions, psychrophilic SRB need to be enriched. Alternatively, a high-temperature process containing thermophilic SRP could be developed for the treatment of hot process waters. The elevated temperature has advantages due to faster microbial metabolism (van Lier 1995) and lower solubility of toxic H2S (Dean 1999).

The wastewater acidity and metal concentrations were efficiently diluted due to the high recycle ratio of the FBR. Therefore, the SRB did not have to tolerate low pH and high metal concentration in the present study. Low pH conditions are needed if various metals are to be precipitated selectively. This can be achieved by dividing the biological sulfate reduction and chemical metal precipitation with biogenic H2S into separate unit processes or using acidophilic and metal tolerant SRB. Future research is needed to develop optimal ways to separate the biogenic H2S from the FBR effluent. Moreover, acidophilic and metal tolerant SRP could be bioprospected for developing a FBR process operating at low pH and high metal concentrations.

The thorough knowledge of the structure of the microbial communities of the FBRs is a good starting point for process optimization. Molecular methods allow tracking of critical groups of microorganisms and specific metabolic reactions that are the key to the satisfactory performance of the FBR system. The DGGE approach revealed some changes in the FBR communities during this study, but the resolution was not good enough to detect all SRB present in the FBRs. DGGE with nested PCR or primers specific for SRB are necessary to improve the specificity of the DGGE to SRB and to monitor the changes in the SRB community structure (Daly et al. 2000; Dar et al. 2004). Primers targeting the functional genes, such as sulfite reductase genes, could be used in addition to the primers targeting the 16S rRNA genes (Voordouw et al. 1990; Nakagawa et al. 2004). Fluorescent in situ

81 hybridization (FISH) with taxon specific probes would allow the quantification of the various SRP in the FBR communities (Amann and Ludwig 2000). FISH could be also combined with microautoradiography to study the physiological traits of the community members (Lee et al. 1999). The development of rapid analysis methods, such as microarrays, enables the fast detection of the community members and the expression of various genes (Loy et al. 2002; El Fantroussi et al. 2003).

The present study resulted in the isolation of several bacterial strains that may represent new species or genera. Some of these strains are being characterized further and formally described. The specificity of the existing 16S rRNA probes for the new strains should be tested and new taxon specific probes could be designed. Additionally, the PLFA composition of the new strains should be determined to find potential biomarkers for the new strains. The traditional anaerobic microbiology techniques did not result in pure cultures of the dominant genotypes of the clone library and DGGE analysis. Micromanipulation systems, such as separation of single cells using microcapillary tubes or optical tweezers, could facilitate the isolation of yet uncultured bacteria (Fröhlich and König 2000).

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114 I

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Reprinted with kind permission of Elsevier Science Ltd.

II

Kaksonen, A.H., Franzmann, P.D. and Puhakka, J.A. 2003. Performance and ethanol oxidation kinetics of a sulfate-reducing fluidized-bed reactor treating acidic metal- containing wastewater. Biodegradation 14: 207-217.

Reprinted with kind permission of Kluwer Academic Publishers.

Electronic version: With kind permission of Springer Science and Business Media.

Biodegradation 14: 207–217, 2003. 207 © 2003 Kluwer Academic Publishers. Printed in the Netherlands.

Performance and ethanol oxidation kinetics of a sulfate-reducing fluidized-bed reactor treating acidic metal-containing wastewater

∗ Anna H. Kaksonen1,2 , Peter D. Franzmann2 & Jaakko A. Puhakka1,2 1Institute of Environmental Engineering and Biotechnology, Tampere University of Technology, P.O. Box 541, FIN- 33101 Tampere, Finland; 2CSIRO Land and Water, Underwood Avenue, Floreat Park, Private Bag No. 5, Wembley WA 6913, Australia (∗author for correspondence: e-mail: anna.kaksonen@tut.fi)

Key words: acetate, ethanol, fluidized-bed reactor, oxidation kinetics, sulfate reduction, wastewater

Abstract The treatment of simulated acidic wastewater (pH 2.5–5) containing sulfate (1.0–2.2 g l−1), zinc (15–340 mg l−1) andiron(57mgl−1) was studied in a sulfate-reducing fluidized-bed reactor (FBR) at 35 ◦C. The original lactate feed for enrichment and maintenance of the FBR culture was replaced stepwise with ethanol over 50 days. The robustness of the process was studied by increasing stepwise the Zn, sulfate and ethanol feed concentrations and decreasing the feed pH. The following precipitation rates were obtained: 360 mg l−1 d−1 for Zn and 86 mg l−1 d−1 for Fe, with over 99.8% Zn and Fe removal, with a hydraulic retention time of 16 h. Under these conditions, 77– 95% of the electrons were accepted by sulfate reduction. The alkalinity produced from ethanol oxidation increased the wastewater pH from 2.5 to 7.5–8.5. Michaelis–Menten constants (Km) determined in batch FBR experiments, were 4.3–7.1 mg l−1 and 2.7–3.5 mg l−1 for ethanol and acetate oxidation, respectively. The maximum oxidation −1 −1 −1 −1 velocities (Vmax) were 0.19–0.22 mg gVS min and 0.033–0.035 mg gVS min , for ethanol and acetate, respectively. In summary, the FBR process produced a good quality effluent as indicated by its low organic content and Zn and Fe concentrations below 0.1 mg l−1.

Abbreviations: COD – chemical oxygen demand; DOC – dissolved organic carbon; FBR – fluidized-bed reactor; HPLC – high performance liquid chromatograph; HRT – hydraulic retention time; Km – Michaelis–Menten constant; RFBR – recirculating fluidized-bed reactor; S – initial substrate concentration; SRB – sulfate-reducing bacteria; T – temperature; TSS – total suspended solids; v – oxidation velocity; Vmax – maximum oxidation velocity; VS – volatile solids; VVS – volatile suspended solids

Introduction Recently, the interest in the application of sulfate re- duction as the dominant step of wastewater treatment The exploitation of sulfide minerals results in the ox- has been growing (for a review, see Hulshoff Pol et al. idation of iron and sulfur, and thus in the production 2001). The process is based on biological hydrogen of acidic metal-containing wastewaters (Foucher et sulfide production (Equation (1)) by sulfate-reducing al. 2001). The techniques traditionally used for the bacteria (SRB), metal sulfide precipitation (Equation treatment of these wastewaters have been based on (2)) and neutralization of the water with the alkalinity chemical neutralization and precipitation. The dis- produced by the microbial metabolism (Equation (3)) advantages of chemical treatment include high cost (Dvorak et al. 1992; Christensen et al. 1996): of the chemical reagents and production of a bulky sludge, which must be disposed of (García et al. + 2− → + − 2CH2O SO4 H2S 2HCO3 ,(1) 2001). In view of this, efforts have been made to de- velop biological alternatives for wastewater treatment. where CH2O = electron donor 208

methanogens support the formation of granules (Vis- 2+ + H2S+M → MS(s) + 2H ,(2)ser 1995; Omil et al. 1996). On the other hand, 2+ the methanogenic activity reduces the sulfide yield. where M = metal, such as Zn A large-scale process utilizing an ethanol-fed UASB − reactor is being used in the Netherlands for the treat- + + → + . HCO3 H CO2(g) H2O (3) ment of metal-contaminated groundwater (Barnes et The sulfidogenic process has several advantages al. 1991; de Vegt & Buisman 1995). In fluidized- over conventional chemical processes, such as lime bed reactors (FBR), biomass is retained on an inert addition. Most metal sulfides form a denser sludge carrier material, which is fluidized with recycled wa- and are less soluble (i.e., more stable) than the hy- ter or a gas stream (gas lift reactors). The advantage droxides produced by chemical treatment. In addition, of a fluidized-bed reactor (compared to packed-bed metal sulfides can be recovered and recycled (Jalali & or UASB reactors) is enhanced mass-transfer of both Baldwin 2000). substrates and toxic products, such as H2S (Nagpal et Acidic metal-containing wastewaters usually con- al. 2000b). For acidic, metal containing wastewaters, tain relatively low concentrations of organic sub- the high recycle ratio dilutes the metal concentrations stances. Therefore, addition of a suitable carbon and acidity in the reactor influent, thus allowing the source and electron donor for sulfate reduction is of- biological sulfate reduction and metal precipitation to ten necessary to promote biogenic H2S production occur in a single reactor. In addition, due to the in- (Kolmert & Johnson 2001). SRB utilize several low tensive mixing, the FBRs are less subject to clogging molecular weight substrates, such as lactate, formate, compared to fixed-bed reactors. acetate, ethanol, and hydrogen. Some SRB oxidize In a previous study, Ma & Hua (1997) demon- organic substrates completely to CO2, while others strated the suitability of lactate-fed FBR for the pre- oxidize them incompletely to acetate (Widdel 1988). cipitation of Cd from synthetic wastewater (pH not Failures of anaerobic wastewater treatment processes reported). Furthermore, lactate-supplemented sulfate- as a result of shock loadings or decreasing pH are reducing FBRs were used to precipitate Zn and Fe often characterized by an increasing concentration of from acidic (pH 2.5) wastewater (Kaksonen et al., in acetic acid in the effluent (Kus & Wiesmann 1995). press). The alkalinity produced in lactate oxidation in- Lactate is a good growth substrate for most SRB creased the initial pH of 2.5 resulting in effluent pH of (Postgate 1979), but the use of lactate in wastewater 7.5–8.5 (Kaksonen et al., in press). In the completely treatment processes would result in high operational mixed FBR, the reactor pH was close to that of the ef- costs (Barnes et al. 1991; Nagpal et al. 2000a). The fluent. Nagpal et al. (2000b) attained sulfate reduction − − use of lactate may speed-up the process start-up due ratesupto6.33gsulfatel 1 d 1 in an ethanol-fed sys- to the relatively high growth yields of sulfate-reducing tem consisting of a FBR and a separate recirculation populations growing on lactate. A low-cost substrate, vessel (recirculating fluidized-bed reactor RFBR) at a such as ethanol, is needed for large-scale operations, hydraulic retention time (HRT) of 5.1 h. In their ex- once a reasonable biomass yield has been achieved periments, the efficiency of sulfate reduction increased (Nagpal et al. 2000a). Another factor that needs con- considerably as the HRT increased until the bacteria sideration is the organic content in the effluent from became strongly substrate-limited at 55 h HRT. the reactor that results from the unused electron donor. In the present work, the effects of the change of For example waste products, such as sewage sludge, electron donor from lactate to ethanol on the FBR produce a secondary chemical oxygen demand (COD) performance were studied. The loading limits of the load, which restricts its disposal to the environment. ethanol-fed process were determined in continuous- TheslowgrowthrateofSRBhasresultedinthe flow experiments. In addition, in batch FBR kin- development of several immobilized biomass react- etic experiments, maximum velocities (Vmax) and ors for treatment of large volumes of wastewater. Michaelis–Menten constants (Km) were determined Anaerobic fixed-bed or packed-bed bioreactors are for the oxidation of the ethanol and its major de- effective in wastewater remediation (Foucher et al. gradation intermediate, acetate (Equations (4) and 2001; Kolmert & Johnson 2001). However, they are (5)). subject to clogging and channelling which decreases + 2− → − + − their long term efficiencies (Kolmert & Johnson 2001). 2CH3CH2OH SO4 2CH3COO HS + In upflow anaerobic sludge blanket reactors (UASB), + H + 2H2O, (4) 209

− Table 1. Composition (mg l 1 except for pH) of the re- actor feeds. The changes in the feed during the loading experiment were as presented in Figure 2.

Loading Batch kinetic experiments experimentsa

Sodium lactate 780 → 0 Ethanol 0 → 710 160 KH2PO4 56 28 NH4Cl 110 55 Ascorbic acid 11 5.5 Thioglycolic acid 11 5.5 Na2SO4·10H2O 1430 → 5690 Na2SO4 1820 MgSO ·7H O 1130 580 Figure 1. Schematic diagram of the process configuration of the 4 2 · fluidized-bed reactor (FBR). FeSO4 7H2O 280 140 ZnCl2 30 → 700 15 pH 5.0 → 2.5 4

a Feed to the reactor between individual batch experi- − + 2− → − + − ments was continuous. CH3COO SO4 2HCO3 HS .(5)

DOC and soluble metal concentrations, respectively. Samples for the determination of total dissolved sulf- Materials and methods ide were taken directly from the sampling port placed in the effluent line (see Figure 1). Bioreactors

A laboratory-scale FBR with a 20% fluidization rate Batch kinetic experiments was used to enrich and maintain SRB at 35 ◦C(Fig- ure 1). Silicate mineral (particle size 0.5–1 mm) was Batch kinetic experiments were conducted in the FBR used as the carrier material. The reactor was ori- to obtain the kinetic parameters (Vmax and Km) for ginally inoculated with sediment from a mining area the oxidation of electron donor. Before and between and methanogenic granular sludge, respectively (Kak- individual batch experiments, the reactor was con- sonen et al. in press). A modified Postgate B medium tinually fed (HRT 24 h) with wastewater (Table 1) (Postgate 1979) (Table 1) containing sulfate (1.0– to wash out the excess hydrogen sulfide formed dur- 2.2gl−1), zinc (15–340 mg l−1) and iron (57 mg l−1) ing the batch experiments, and to maintain a constant was fed to the reactor at a HRT of 16 h. biomass content in the reactor. The pH of the feed was 4 and the resulting pH in the reactor and effluent Replacement of the electron donor and loading was 6.5–7.5. Sulfate was added in excess (three times experiments as much as stoichiometrically required) for complete oxidation of the electron donor. For each batch exper- For enrichment and maintenance of the FBR culture, iment, the feed flow was discontinued and the FBR lactate was fed to the reactor for 210 days. Maxim- was operated in recycle mode. A portion (50 ml) of ally, 60–75% of the electrons were used for sulfate the reactor liquid was replaced through the sampling reduction. The lactate feed was replaced stepwise with valve placed in the recycling line (Figure 1) with fresh ethanol over 50 days. The sulfate, ethanol and Zn in- solution (pH 7) containing Na2SO4, nutrients and eth- fluent concentrations were increased stepwise and feed anol or acetate. This solution was purged with nitrogen pH was decreased gradually until process failures oc- before adding it to the FBR. During the batch exper- curred. The experimental design is shown in Figure 2. iments, samples of the reactor liquid were collected Rates of sulfate reduction, dissolved organic carbon through the sampling valve placed in the recycling (DOC) removal and metal precipitation were determ- line at regular intervals and analyzed for ethanol and ined as the difference of feed and effluent sulfate, acetate. In addition, sulfate and dissolved sulfide were 210

photometer (Philips PU9200X, Great Britain) accord- ing to the Finnish standards SFS 3044 (SFS 1980a) and SFS 3047 (SFS 1980b). Liquid pH was determ- ined in unfiltered samples using a pH electrode (WTW SenTix41, Germany). Total alkalinity was analysed by titrating unfiltered samples with 0.02 M HCl to pH 4.5 according to the standard SFS-EN ISO 9963-1 (SFS 1996). Total suspended solids (TSS) and volatile sus- pended solids (VSS) were analysed in effluent samples by filtering a measured volume of sample through a glass fiber filter (Whatman GF/A), drying the filter for 1 hour at 105 ◦C, to determine TSS, with subsequent ignition at 550 ◦C for 1 hour, for the determination of VSS. Ethanol and acetate were determined using a gas chromatograph (Hewlett Packard 5890A, USA) equipped with a 25 m × 0.32 mm (ID) capillary column coated with a 0.25 µm film of polyethyl- ene glycol treated with TPA (BP21, SGE Australia) and a flame ionisation detector. The amount of bio- mass in the reactor was estimated as volatile solids (VS) according to the Finnish standard SFS 3008 (SFS 1990). The ethanol and acetate oxidation rates were standardized to the total amount of biomass in Figure 2. Experimental design of continuous-flow fluidized-bed the FBR. Michaelis–Menten equation (Equation (6)) reactor experiments showing changes in the feed composition. and five different plotting approaches: Lineweaver– Burk plot (Equation (7)), Hanes plot (Equation (8)), Eadie–Hofstee plot (Equation (9)), Direct linear plot determined in the first and last sample of each batch (Equation (10)) and Modified linear plot (Equation experiment. (11)) were used to obtain the kinetic parameters (Vmax and Km) (Cornish-Bowden 1995). Analytical methods V · S ν = max (6) For sulfate, DOC, ethanol, acetate, soluble metal and Km + S sulfide analyses, the samples were filtered through 1 1 Km 1 0.45 µm polyethersulfone membrane syringe filters. = + · (7) For the loading experiments, sulfate was determined ν Vmax Vmax S by ion chromatography (Dionex DX-120, USA) and S K 1 total dissolved sulfide was analyzed spectrometrically = m + · S (8) (Shimadzu UV-1601, Japan) by the methylene blue ν Vmax Vmax method (Trüper & Schlegel 1964). For batch kinetic ν experiments, sulfate was determined using a high per- ν = Vmax − Km · (9) formance liquid chromatograph (HPLC) system con- S sisting of a Millipore Waters IC-PakTM Anion column ν (4.6 mm × 5 cm), GBC LC 1610 Autosampler, Vmax = ν + · Km (10)
S GBC LC 1150 HPLC pump and a Millipore Waters Model 430 Conductivity Detector, and total dissolved 1 = 1 − 1 · Km sulfide was analyzed spectrometrically (Helios Ep- . (11) Vmax ν S Vmax silon, 9423 UVE 100E, USA) by the colorimetric method described by Cord–Ruwisch (1985). DOC was measured with a TOC-5000 Analyzer (Shimadzu, Japan), and metals with an atomic absorption spectro- 211

Figure 3. Quantitative sulfate load and reduction rate (a), DOC load and degradation rate (b), Zn (c) and Fe (d) load and precipitation rate, and percent sulfate reduction (e), DOC degradation (f), Zn (g) and Fe (h) precipitation for the fluidized-bed reactor (FBR) during the electron donor change (period indicated by the arrow and dashed lines) and subsequent loading experiments.

Results and discussion substrate for a sulfidogenic UASB during an elec- tron donor change from lactate to ethanol. Methanol Replacement of the electron donor and loading accelerated the growth of methanogens and hence experiments complete degradation of the acetate occurred. In the present study, no significant acetate accumulation was During the electron donor change (between days 210 observed during the electron donor change. and 260), 51–72% of the DOC was consumed by The capacity of the FBR-process was studied sulfate reduction (Figure 3a, b, e and f) with 80– by increasing stepwise the Zn, sulfate and ethanol 97% DOC removal. In a previous study by Barnes feed concentrations and decreasing the feed pH (Fig- et al. (1991) methanol was used as an intermediate 212 ure 2). The following precipitation rates were ob- tained: 360 mg l−1 d−1 for Zn (Figure 3c) and 86 mg l−1 d−1 for Fe (Figure 3d), with over 99.8% Zn and Fe removal (Figure 3g and h), at a HRT of 16 h. Ma & Hua (1997) attained greater Cd precipita- tion rates with a lactate-fed FBR; the pH of the feed was not reported. Metal precipitation and pH neut- ralization are interdependent, since the precipitation reaction (Equation (2)) produces acidity. In the present study, the pH (2.5) of the influent water was neut- ralized by the alkalinity produced (Figure 4b and c) and the effluent pH remained at 7.5–8.5 during stable reactor performance. Zinc has been reported to completely inhibit sulfate-reduction in the concentration range of 25 to 60 mg l−1 (Morton et al. 1991; Ueki et al. 1991). In the present study, the sulfate-reducing FBR was able to treat wastewater with 240 mg Zn l−1 resulting in an effluent with less than 0.1 mg Zn l−1.Inthecom- pletely mixed FBR, bacteria became exposed to Zn concentrations close to those of the effluent, and thus, toxicity problems were avoided. The effluent DOC during stable FBR performance with the highest loadings was 7–20 mg l−1. Concen- trations of TSS (Figure 4d) and VSS (Figure 4e) in the effluent increased as loading rates were increased. The efficiency of the ethanol-fed FBR to precipitate metals and neutralize acidity was similar to that repor- ted with lactate-fed FBRs (Kaksonen et al., in press). However, at the highest loadings, a greater percentage of the added electron donor was utilized for sulfate reduction in the ethanol-fed FBR (77–95%) as com- pared to the lactate-fed FBR (60–75%). The rest of the electron flow was most likely coupled with ferment- ative reactions. Biogas was not produced and hence, methanogenesis was insignificant. This study focused on treatment of simulated zinc and iron containing wastewater. The sulfate-reducing FBR process can potentially be used to precipitate other toxic metals, which have high affinity to S2−- ions, such as Cu, Pb and Cd. The low solubilities required for highly toxic metals such as mercury are often achievable only by speciation as sulfides, since most metal sulfides (As and Cr are exceptions) have solubilities several orders of magnitude lower than the hydroxides throughout the pH range (Conner 1990). Figure 4. Effluent sulfide (a), feed and effluent pH (b), feed and effluent alkalinity (c), effluent total suspended solids (TSS) (d) and Batch kinetic experiments volatile suspended solids (VSS) (e) for the fluidized-bed reactor (FBR) during the electron donor change (period indicated by the Ethanol and acetate oxidation kinetics were stud- arrow and dashed lines) and subsequent loading experiments. ied in batch FBR experiments. Oxidation of ethanol 213

Figure 5. Oxidation of ethanol (a and b) and acetate (c and d) in the batch kinetic experiments with different initial substrate concentrations. and acetate at different initial concentrations was as than ethanol was oxidized to acetate. Sulfate and sulf- presented in Figure 5. Acetate oxidation was the rate- ide analysis confirmed that excess sulfate was present limiting step in ethanol oxidation (Equations (4) and in the FBR during the experiments and initial sulfide (5)), as demonstrated by the accumulation of acetate concentrations remained below 100 mg l−1. in the FBR during the batch experiments (Figure 6). The Vmax for ethanol oxidation was greater than The total quantity of biomass in the FBR was 4.0 g and the observed oxidation velocities during the continu- 3.8 g during the ethanol and acetate oxidation exper- ous loading experiments, whereas Vmax for acetate was iments, respectively. Approximately 85–86% of the lower (Figure 3b), if the biomass content of the FBR biomass was carrier bound, 10–11% on the metal pre- is assumed to be the same as during the batch experi- cipitates accumulated above the carrier material and ments. The Km and Vmax values for acetate oxidation 3–4% occurred in the bulk water. The effects of the were lower than previously reported for mesophilic initial substrate concentrations on oxidation rates are sulfate-reducing cultures (Table 3). The observed low shown in Figure 7a and the Lineweaver–Burk plot, Vmax indicates that only part of the total biomass in Hanes plot, Eadie–Hofstee plot, Direct linear plot the FBR utilized acetate as an electron donor. Many of and Modified linear plot in Figures 7b to 7f, respect- the known sulfate-reducers incompletely oxidize their ively. Michaelis–Menten constants (Km) were 4.3– substrates producing acetate as an end product (Castro 7.1mgl−1 and 2.7–3.5 mg l−1 for ethanol and acetate, et al. 2000). As shown in the loading experiments, up respectively (Table 2). The maximum oxidation velo- to 77–95% of the added electron flow was used for −1 −1 cities (Vmax) were 0.19–0.22 mg gVS min and sulfate reduction, the rest was coupled most likely to 0.033–0.035 mg gVS−1 min−1, for ethanol and acet- fermentative reactions. ate, respectively (Table 2). The results show that To the best of our knowledge, ethanol oxidation acetate was oxidized to CO2 considerably more slowly kinetics in sulfate-reducing systems have been pre- 214

Table 2. Michaelis–Menten constants (Km) and maximum oxidation velocities (Vmax) for ethanol and acetate oxidation obtained with different linearization approaches.

−1 −1 −1 Linearization Km (mg l ) Vmax (mg gVS min ) approach Ethanol Acetate Ethanol Acetate

Lineweaver–Burk 5.0 2.9 0.20 0.034 Hanes plot 7.1 3.5 0.22 0.035 Eadie–Hofstee plot 4.9 2.7 0.20 0.034 Direct linear plot 5.1 2.9 0.20 0.034 Modified linear plot 4.3 2.5 0.19 0.033

Table 3. Kinetic parameters for acetate oxidation by mesophilic sulfate-reducing pure and enrichment cultures.

TKm Vmax ◦ − − − Strain/culture ( C) (mg l 1) (mg gVS 1 min 1) Reference

Desulfobacter postgatei 30 13.6 0.4–1.2a Schönheit et al. (1982) Desulfobacter postgatei 30 3.8–4.5 3.0–3.1 Ingvorsen et al. (1984) Desulfobacter postgatei 30 7.1–9.4 Widdel (1988 Desulforhabdus amnigenus 37 35 1.7a Oude Elferink (1998) Desulfobacca acetoxidans 37 35 2.5a Oude Elferink (1998) Enrichment culture 31 5.9 Middleton & Lawrence (1977) Mixed culture of SRB and methanogens 30 9.5 Yoda (1987) Enrichment culture 35 2.7–3.5 0.033–0.035 This study

a Biomass determined as protein instead of VS.

(0.2 g l−1) was obtained using a batch-enriched sulfate reducing culture and batch bottle assays (Nagpal et al. 2000a) and was over one order of magnitude greater than that reported in our study (5 mg l−1). This shows that the FBR system selectively enriches for sulfate reducers with high affinities for the electron donor. Based on the low Km values, the ethanol and acet- ate oxidation rates remain high even at low substrate concentrations. This implies low residual organic con- tent in the effluent, and hence a potentially low en- vironmental impact associated with the discharge of the effluent water. This was also shown as the low concentration of DOC in the effluent water of the FBR during the continuous flow operation. In compar- ison, Nagpal et al. (2000b) did not report significant acetate consumption in the ethanol-fed sulfate-limited RFBR containing Desulfobacter postgatei (complete Figure 6. Ethanol concentration decrease (filled symbols) and acet- oxidizer) and Desulfovibrio desulfuricans (incomplete ate accumulation (unfilled symbols) in the sulfate-reducing FBR oxidizer). Nagpal et al. (2000b) discussed that this during four batch experiments. could be due to the inhibition of acetate-consuming bacteria either by sulfide or acetate concentrations, or the better ability of Desulfovibrio to compete for the viously reported only by Nagpal et al. (2000a). In their study, the half saturation constant for ethanol 215

Figure 7. (a) Effect of initial concentrations (S) on ethanol (filled symbols) and acetate (unfilled symbols) oxidation rates (v) in the batch kinetic experiments. Linearization of the results with (b) Lineweaver–Burk plot, (c) Hanes plot, (d) Eadie–Hofstee plot, (e) direct linear plot and (f) modified linear plot. The arrows point to the groups of ethanol and acetate plots, respectively. 216 limiting amounts of sulfate (Laanbroek et al. 1984; Cord-Ruwisch R (1985) A quick method for the determination of Nagpal et al. 2000b). dissolved and precipitated sulfides in cultures of sulfate-reducing The acetate oxidation step (Equation (5)) in eth- bacteria. JMM 4: 33–36 Cornish-Bowden A (1995) Fundamentals of enzyme kinetics. Re- anol oxidation produces the bicarbonate alkalinity, vised edition. Portland Press Ltd, London which neutralizes the acidic wastewater. In addition, Dvorak DH, Hedin RS, Edenborn HM & McIntire PE (1992) the acetate oxidation produces twice as much hydro- Treatment of metal-contaminated water using bacterial sulfate- reduction: Results from pilot-scale reactors. Biotech. Bioeng. 40: gen sulfide per mole than does ethanol oxidation to 609–616 acetate (Equation (4)). In summary, acetate oxida- Foucher S, Battaglia-Brunet F, Ignatiadis I & Morin D (2001) Treat- tion is the rate limiting step, the kinetics of which ment by sulfate-reducing bacteria of Chessy acid-mine drainage define the design criteria for the ethanol-supplemented and metals recovery. Chem. Eng. Sci. 56: 1639–1645 García C, Moreno DA, Ballester A, Blázquez ML & González F sulfate-reducing bioprocess for metal precipitation. (2001) Bioremediation of an industrial acid mine water by metal- tolerant sulphate-reducing bacteria. Min. Eng. 14(9): 997–1008 Hulshoff Pol LW, Lens PNL, Weijma J & Stams AJM (2001) Conclusions New developments in reactor and process technology for sulfate reduction. Water Sci & Technol 44(8): 67–76 Ingvorsen K, Zehnder AJB & Jørgensen BB (1984) Kinetics of The sulfate-reducing FBR culture, that was originally sulfate and acetate uptake by Desulfobacter postgatei. Appl. enriched with sodium lactate, utilized ethanol with a Environ. Microbiol. 47: 403–408 higher efficiency (77–95% of the electrons used for Jalali K & Baldwin SA (2000) The role of sulphate reducing bacteria in copper removal from aqueous sulphate solutions. Water Res. sulfate reduction) when compared to the efficiency 34: 797–806 of lactate utilization (60–75%). The FBR-treatment Kaksonen AH, Riekkola-Vanhanen M-L & Puhakka JA (in press) precipitates metals as sulfides and neutralizes the Optimization of metal sulfide precipitation in fluidized-bed treat- wastewater from an influent pH of 2.5, at a HRT of ment of acidic wastewater. Water Res Kolmert Å & Johnson DB (2001) Remediation of acidic waste wa- 16 h. Acetate oxidation is the rate-limiting step in ters using immobilised, acidophilic sulfate-reducing bacteria. J. sulfidogenic ethanol oxidation. The sulfidogenic FBR Chem. Technol. Biotechnol. 76: 836–843 system enriches selectively bacteria with low Km val- Ku¸s F & Wiesmann U (1995) Degradation kinetics of acetate and ues for ethanol and acetate oxidation. In conclusion, propionate by immobilized anaerobic mixed cultures. 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Kaksonen, A.H., Franzmann, P.D. and Puhakka, J.A. 2004. Effects of hydraulic retention time and sulfide toxicity on ethanol and acetate oxidation in sulfate-reducing metal- precipitating fluidized-bed reactor. Biotechnology and Bioengineering 86: 332-343.

Reprinted with kind permission of Wiley Periodicals, Inc.

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Kaksonen, A.H., Plumb, J.J., Franzmann, P.D. and Puhakka, J.A. 2004. Simple organic electron donors support diverse sulfate-reducing communities in fluidized-bed reactors treating acidic metal- and sulfate-containing wastewater. FEMS Microbiology Ecology 47: 279-289.

Reprinted with kind permission of the Elsevier B.V.

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Kaksonen, A.H., Plumb J.J., Robertson, W.J., Franzmann, P.D., Gibson, J.A.E. and Puhakka, J.A. 2004. Culturable diversity and community fatty acid profiling of sulfate- reducing fluidized-bed reactors treating acidic metal-containing wastewater. Geomicrobiology journal (in press).

Reprinted with kind permission of Taylor & Francis, Inc.

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