Microbiology Fundamentals - NTUAold-2017.metal.ntua.gr/uploads/4433/4_microbiology.pdf · Microbiology Fundamentals 990 metabolism and homeostasis appear to be a major concern for

CHAPTER 4

Microbiology Fundamentals

15th International Biohydrometallurgy Symposium (IBS 2003) September 14-19, Athens, Hellas

"Biohydrometallurgy: a sustainable technology in evolution"

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A model for iron uptake in Acidithiobacillus ferrooxidans based upon genome analysis

R. Quatrini1, F. Veloso1,3, E. Jedlicki2 and D.S. Holmes1,3* 1 Laboratory of Bioinformatics & Genome Biology, University of Santiago, Santiago,

Chile 2 Program of Cellular and Molecular Biology, ICBM, University of Chile, Santiago, Chile

3 Millenium Institute of Fundamental and Applied Biology, Santiago, Chile

Abstract During growth at acid pH, Acidithiobacillus ferrooxidans is confronted with an

abundant supply of soluble iron. Therefore, it may have developed iron uptake mechanisms quite distinct from those of neutrophilic organisms who are faced with the problem of the uptake of insoluble forms of iron or of soluble iron at very low concentrations. An analysis of the nearly complete genome sequence of A. ferrooxidans reveals the presence of a number of potential genes and regulatory pathways involved in iron uptake typical of those found in other organisms. However, in addition, bioinformatic analysis suggests that A. ferrooxidans exhibits several unusual genetic features that may reflect special requirements for iron uptake and homeostasis required for life at acidic pH. Finally, an analysis of iron uptake mechanisms in A. ferrooxidans supports the growing contention that the organism can live at a variety pHs between pH1 and, at least, pH 5.5.

Keywords: Acidithiobacillus ferrooxidans, iron uptake, iron regulation, FeoB, TonB-dependent receptors, genome analysis

1. INTRODUCTION One of the critical problems confronting all organisms is the acquisition and uptake of

iron to fulfill metabolic requirements. Because iron is largely insoluble at neutral pH under aerobic conditions, neutrophilic organisms have evolved a variety of sophisticated iron uptake mechanisms. In addition, they have developed homeostatic controls to tightly balance iron intracellular levels, guarding cell integrity against the deleterious effects of excess iron (1,2).

The regulation of iron metabolism and its coupling with the regulation of defences against oxidative stress must be strictly adjusted in response to environmental iron bioavailability. Judging by the plethora of genes and response mechanisms involved, iron

* Corresponding author: David Holmes: [email protected]. Work supported by Fondecyt No. 1010623 and the Millenium Institute of Fundamental and Applied Biology, Santiago, Chile. We thank the Institute of Genome Research (TIGR) and Integrated Genomics, Inc. (IG) for the use of their partial sequence of the Acidithiobacillus ferrooxidans genome.


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metabolism and homeostasis appear to be a major concern for life in the presence of oxygen.

We are interested in understanding how the bacterium Acidithiobacillus ferrooxidans copes with iron uptake and homeostasis considering that it can grow aerobically at acidic pHs in environmental situations where soluble iron is abundant. In addition, a special feature of A. ferrooxidans, not found in the majority of organisms, is its use of iron as energy and electron sources. This utilization must be balanced with the need for iron in metabolism suggesting that A. ferrooxidans might exhibit novel iron regulatory mechanisms not observed in neutrophilic heterotrophic organisms.

Contrary to what happens in most aerobic environments, where FeII rapidly autooxidizes to FeIII which precipitates as ferric hydroxides, in acidic conditions FeII is relatively stable and FeIII is much more soluble (>0.1M), exceeding by far the typical requirements of bacteria of 10-8M (3,4). Given that A. ferrooxidans is usually confronted with an abundant supply of both FeII and FeIII at acidic pHs, one can ask if it has evolved novel mechanisms for Fe uptake that are not present or are substantially different from those exhibited by neutrophilic organisms. One can also ask if this abundant supply of Fe has necessitated the development of special or different mechanisms in A. ferrooxidans for iron storage and for the avoidance of oxidative stress due to excess iron.

Studies concerning the strategies used by A. ferrooxidans for iron assimilation, iron homeostasis and avoidance of the oxidative stress threat imposed by a unique iron-rich and O2-plentiful environment are lacking and prompted the present investigation. Knowledge of the physiology and metabolism of A. ferrooxidans has been impeded by the paucity of classical genetic and molecular biological tools that have been so successfully used to unravel the metabolism of other organisms. Due to its applied interest for biomining and to circumvent this limitation, A. ferrooxidans was the first biomining microorganism to have its genome almost completely sequenced (5) making it a candidate for bioinformatic gene predictions and subsequent functional and metabolic model building (6). In this paper, we present a partial bioinformatic survey of potential genes of A. ferrooxidans involved in iron uptake. Candidate genes and pathways were first identified by comparison to known iron uptake genes and pathways in other organisms. Additional candidate genes were then identified by linkage analysis to the previously known genes. Contributing to the elucidation of the physiological responses of A. ferrooxidans is of fundamental interest and might, eventually, have a biotechnological impact on the mining industry.

2. METHODS Known metabolic pathways of iron acquisition were obtained from BIOCYC

(http://biocyc.org:1555/META/server.html), KEGG (http://genome.ad.jp/kegg/kegg4. html) and ERGO (http://wit.integratedgenomics.com/WIT2/). Amino acid sequences derived from genes identified as being involved in iron acquisition were used as query sequences to search the partial genome sequence of A. ferrooxidans ATCC 23270 in the TIGR (http://www.tigr.org/) and ERGO data bases using TBlastN and BlastP respectively. When a prospective candidate gene was identified in TIGR or ERGO its predicted amino acid sequence was then used to formulate a BlastP (http://www.ncbi.nlm.nih.gov/BlastP/) search of the non-redundant database at NCBI. Only bidirectional best hits were accepted as evidence for putative homologs. Candidate genes and their translated proteins were further characterized employing the following bioinformatic tools available in the web: primary structure similarities (http://www.ebi.ac.uk/ClustalW/), secondary structure


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predictions (HMM-based Protein Sequence Analysis http://www.cse.ucsc.edu/research/ compbio/HMM-apps/T99-query.html; JPred http://www.compbio.dundee.ac.uk/Software /JPred/jpred.html), transmembrane predictions (http://www.ch.embnet.org/software/ TMPRED_form.html), motif predictions (http://www.blocks.fhcrc.org/, http://www.ebi. ac.uk/interpro domain predictions (http://www.tolouse.inra.fr/prodom.htlm) and prediction of protein localization sites (http://psort.nibb.ac.jp/).

3. RESULTS AND DISCUSSION Bioinformatic similarity searches for orthologous genes (evolutionarily homologous

gene with the same function in different species) in A. ferrooxidans involved in iron uptake has revealed a number of candidate genes that can be assigned with a reasonable level of confidence, due to their similarity with experimentally proven iron uptake genes in other organisms (Table 1). Given the presence of such genes, metabolic models of iron uptake in A. ferrooxidans can be inferred based upon known models in other organisms. A summary of these results is shown in Figure 1. Additional work is in progress to identify genes and pathways involved in iron sensing, homeostasis, storage and detoxification.

A. ferrooxidans encodes a putative FeII inner-membrane transport system termed FeoAB (Figure 1A). FeoAB is present in most of the sequenced genomes but has been observed to function only under anaerobic conditions where FeII is more stable and soluble (7-9). As in other microorganisms, the feoB gene in A. ferrooxidans is adjacent to a small open reading frame (ORF) termed feoA, of unknown function. Interestingly, in A. ferrooxidans feoA and feoB appear to constitute an operon together with an ORF encoding a putative porin (herein designated porA), which could facilitate FeII entrance in the periplasm. Upstream of this porin is a putative Fur-box, suggesting typical iron-dependent repression of FeII uptake mediated by a Fur-like repressor as has been demonstrated in Escherichia coli (10). Bioinformatic analysis of the putative FeoB reveals the presence of a highly conserved N-terminal GTPase motif and eight hydrophobic-rich transmembrane segments supporting the contention that it is a membrane transport protein. In addition, the second and third periplasmic loops have motifs similar to cytochrome CB5 and B6 which could represent iron recognition or iron interacting portions of the protein.

Sequence comparison of the putative A. ferrooxidans FeoB ortholog and the other three FeoB proteins so far characterized (GenBank accession #s:P33650, P74884, D71909) reveals substantial amino acid sequence similarity except in the so-called connector loop and in the C-terminal half of the protein. FeoB is an inner membrane protein with several exposed loops facing the periplasm and it is these exposed loops that harbour the major differences between the FeoB of A. ferrooxidans and the other three orthologs of FeoB.

We speculate that these differences evolved as a response to the need for function and structural integrity of the A. ferrooxidans FeoB in a low pH environment. In addition, FeoB proteins of A. ferrooxidans and Helicobacter pylori have predicted periplasmic loops with higher isoelectric points than E. coli. This could be a reflection of the functional adaptation of these proteins to acidophilic or acid tolerant conditions. It has recently been demonstrated that the FeoB of H. pylori plays an important role in iron acquisition in the low-pH and low-oxygen environment of the stomach (7). We propose that the FeoB of A. ferrooxidans may serve a similar function in iron uptake at low pH.

Inspection of the A. ferrooxidans genome also reveals the presence of several putative outer membrane iron receptors (OMRs) (Figure 1B) with predicted affinities for either hydroxamate-type siderophores or for FeIII-dicitrate. The degree of primary structure


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similarity and the general accordance of secondary structure predictions between the FhuA (FeIII-hydroxamate siderophores uptake) and FecA (FeIII-dicitrate uptake) of A. ferrooxidans and E. coli highlights the conserved organization of the proteins. These similarities include similar signal peptides, 2 to 4 N-terminal α-helices, 22 β-strands in the putative β-barrel domain and a TonB interaction C-terminal motif (11). Several of the genes encoding OMRs are next to genes encoding TonB energy-transducing systems (exbBDtonB) and some of them cluster together with all the necessary elements (OMR, TonB system, ABC transporter components) to transport FeIII iron into the cytoplasm (Figure 2) as observed in other organisms (12). The conservation of the siderophore receptor proteins and organization of their genes into putative operons together with relevant genes encoding other necessary components for iron uptake, allows us to propose that mechanisms of siderophore uptake exist in A. ferrooxidans that are similar to those found in other organisms.

Figure 1. Proposed model for iron uptake and storage A. ferrooxidans based upon a bioinformatics analysis of its genome. (A) FeII uptake via an outer membrane porin (black) and inner membrane FeoAB transporter complex (white). (B) FeIII uptake via an outer membrane TonB dependent receptor OMR (black) and a periplasmic siderophore binding and inner membrane uptake mechanism involving ABC transporter Fec-like proteins (white) and a TonB energy transducing system (dark gray). (C) Cytoplasmic FeII storage via a bacterioferritin-like protein Bfr.

Interestingly, the several OMRs of A. ferrooxidans exhibit a wide range of isoelectric points (Table 1) that, by comparison to the OMRs of other organisms, would potentially allow them to import iron at different pHs ranging from pH 1 to at least pH 5.5. This supports the growing contention that A. ferrooxidans can live in various environments with marked differences in pH (13, 14). It is not known how A. ferrooxidans might regulate and coordinate the expression of these different OMRs to correspond to particular environmental requirements for iron uptake.

An interesting aspect of the organization of the proposed FeIII-siderophore transport operon in A. ferrooxidans is the presence of a small ORF potentially encoding a hemoglobin-like protein (Figure 2). A bioinformatic analysis suggests that it belongs to a


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family of genes known as truncated hemoglobins, identified in prokaryotes, protozoa, algae and plants. The role of this hemoglobin-like protein is still unclear, although it has been shown in plants to have low oxygen affinity and that the expression of its mRNA decreases in microaerophilic conditions (15). The hemoglobin-like gene is associated with other transporters annotated in genome database of ERGO in Burkholderia cepacia (downstream of a putative molybdenum transport), Ralstonia eutropha (downstream of an uncharacterized permease) and Sphingomonas aromaticivorans (JGI) (downstream of an OMR). In addition to transporting and storing O2 and facilitating its diffusion, several novel functions have emerged regarding hemoglobins, including control of nitric oxide levels in microorganisms, binding and transport of sulfide in endosymbiont-harboring species and protection against sulfide, scavenging of O2 in symbiotic leguminous plants, O2 sensing in bacteria and archaebacteria, and dehaloperoxidase activity useful in detoxification of chlorinated materials (16). In the case of A. ferrooxidans, its association with a siderophore operon might represent a strategy to diminish the risks of O2 toxicity caused by the entering iron until the metal has been targeted to its final destination.

Table 1. Identification and properties of candidate genes in A. ferrooxidans involved in iron uptake. (A) name of genes involved in FeII uptake, FeIII-hydroxamate uptake and FeIII-dicitrate uptake. (B) The % similarity of the predicted proteins from the candidate genes to known proteins deposited in data banks. (C) lists the GenBank accession number of the known proteins and (D) the organisms from which they are derived. (E) lists the predicted location (loc) of the proteins where C: Cytoplasm, OM: Outer Membrane and IM: Inner Membrane. (F) : predicted isoelectric point (pI) of the mature proteins.

(A) (B) (C) (D) (E) (F) Gene % S Accesion # Organism Lo pI

porA 42 ZP00053838 Magnetospirillum magnetotacticum OM 5.85 feoA 55 CAB49974 Pyrococcus abyssi C 11.14FeII

uptake feoB 53 BAC09292 Thermosynechococcus elongatus IM 7.28 omr1 50 AAM72768 Chlorobium tepidum OM 6.98 omr4 57 AAM43451 Xanthomonas campestris OM 8.18 omr6 37 ZP00125990 Pseudomonas syringae OM 9.02 omr7 38 ZP00003769 Nitrosomonas europaea OM 9.15 omr9 37 AAM35906 Xanthomonas axonopodis OM 6.53

FeIII hydroxamate

uptake

omr11 37 AAM42549 Xanthomonas campestris OM 8.63 omr2 37 ZP00125403 Pseudomonas syringae OM 7.26 omr3 38 ZP00125403 Pseudomonas syringae OM 9.02

FeIII dicitrate uptake omr10 38 AAM35740 Xanthomonas axonopodis OM 8.22

In accordance with what has been described for several siderophore transporting genes in other bacteria this putative operon could be iron regulated, since a well conserved Fur-box like sequence has been identified within the predicted promoter region of the globin-like gene (unpublished data). The existence in A. ferrooxidans of an gene with 79% similarity to the E. coli Fur ortholog (unpublished data) strengthens this interpretation and promulgates further intriguing questions about iron regulation in this organism.


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Figure 2. Gene organization of the proposed FeIII-siderophore uptake region of the A. ferrooxidans genome. A: Globin-like gene; B: Outer Membrane TonB-dependent Receptor I; C: TonB energy transducing system (exbB-exbD-tonB); D: ABC transporter I Fhu-like proteins (perplasmic binding protein, permease, ATP-binding protein); E: Outer Membrane TonB-dependent Receptor II and F: ABC transporter II Fec-like proteins (perplasmic binding protein, permease, ATP-binding protein). Also shown is the putative 19 nt Fur–box. In the upper right corner is a scale bar in base pairs (bp)

4. CONCLUSIONS A. ferrooxidans seems to have a complexity of iron uptake systems that rivals that

found in neutrophilic organisms. This argues against a simplistic view that A. ferrooxidans has an easy time encountering and taking up readily available soluble iron in an acid environment. It is possible that it requires this complexity despite the readily available iron. However, a more reasonable argument is that it sometimes lives at higher pHs than those typically associated with its growth in acid ferrous sulfate and sulfur, in which it must scavenge less soluble forms of iron perhaps in strong competition with other organisms. Notable is the plethora of siderophore uptake systems of A. ferrooxidans characteristic of most organisms that have to scavenge and compete for poorly soluble FeIII at neutral pHs. Demonstrations of the capacity of A. ferrooxidans to live at higher pHs have appeared in the literature (13,14) and our findings, regarding iron acquisition systems, support this point of view.

On the other hand, A. ferrooxidans iron uptake systems include components that seem to be adapted for function at low pH, most obvious are the higher isoelectric points of the periplasmic loops of membrane-associated iron uptake pumps. This characteristic is expected for the maintenance of structural integrity and possible function of proteins exposed to low pH.

The presence of a hemoglobin-like protein in A. ferrooxidans possibly associated with a FeIII siderophore uptake operon is unusual, although descriptions of a similar protein are now appearing in the literature. The best guess for its function is that it protects cells from excess oxygen. Alternatively, and not mutually exclusive with this hypothesis, is the possibility that it may measure oxygen levels and regulate genes in response to the changes in oxygen concentration.

In this paper, we have focused on iron uptake mechanisms. Future studies will address mechanisms to sense environmental levels of iron and the regulation of expression of iron uptake and storage genes and how these might be coupled to genes involved in the important capacity of A. ferrooxidans to oxidize iron for energy and electron gain. Preliminary evidence points to the presence of a complex network of Fur regulated operons and experimental work is underway to validate some of the bioinformatic models we have developed.


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REFERENCES 1. J.M. De Freitas and R. Meneghini. Mutat. Res., 475 (2001) 153. 2. D. Touati, M. Jacques, B. Tardat, L. Bouchard and S. Despied. J. Bacteriol., 177

(1995) 2305. 3. K. Hantke and V. Braun. In: Bacterial Stress Responses. (Chapter 19) G. Storz & R.

Hengge-Aronis. (Eds). ASM Press, Washington, D.C., 2000. 4. V. Braun and H. Killmann. Trends. Biochem. Sci., 24 (1999) 104. 5. E. Selkov, R. Overbeek, Y. Kogan, L. Chu, V. Vonstein, D.S. Holmes, S. Silver, R.

Haelkorn & M. Fonstein. Proc. Natl. Acad. Sci. USA, 97 (2000) 3509. 6. D.S. Holmes, J. Valdéz, C. Dominguez, M. Barreto, C. Arriagada, S. Silver, S. Bueno

and E. Jedlicki. International Biohydrometallurgy Symposium (2001) 237. 7. J. Velayudhan, N.J. Hughes, A.A. McColm, J. Bagshaw, C.L. Clayton, C.S. Andrews

and D.J. Kelly. Mol. Microbiol., 37 (2000) 274. 8. R.M. Tsolis, A.J. Baumler, F. Heffron and I. Stojiljkovic. Infect. Immun., 64 (1996)

4549. 9. I. Stojiljkovic, M. Cobeljic and K. Hantke. FEMS Microbiol. Lett., 108 (1993) 111. 10. M. Kammler, C. Schön and K. Hantke. J. Bacteriol., 175 (1993) 6212. 11. R. Koebnik, K.P. Locher and P. van Gelder. Mol. Microbiol., 37 (2000) 239. 12. J.H. Crosa. Microbiol. Molec. Biol. Rev., 61(1997) 319. 13. N. Ohmura, K. Sasaki, N. Masumoto and H. Saiki. J. Bacteriol., 184 (2002) 2081. 14. M. Barreto, M. Rivas, E. Jedlicki and D.S. Holmes. International Biohydrometallurgy

Symposium (2003) (submitted, the Proceedings of this Congress). 15. R.A. Watts, P.W. Hunt, A.N. Hvitved, M.S. Hargrove, W.L. Peacock and E.S. Dennis

Proc. Natl. Acad. Sci. USA, 98 (2001) 10119. 16. R.E. Weber and S.N. Vinogradov. Physiol. Rev., 81(2001) 569.



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Activity and occurrence of leaching bacteria in mine waste at Cartagena, Spain, in the years 1991 until 2000

W. Sand, D. El Korchi-Buchert, T. Rohwerder*

Department of Microbiology, Institute for General Botany, University of Hamburg, Ohnhorststr. 18, D-22609 Hamburg, Germany,

Tel/Fax: +49 40 428 16 423, e-mail: [email protected]

Abstract During the years 1991 until 2000 a total of 31 sampling campaigns have been

performed in the abandoned mining area between Portman and La Union near the city of Cartagena in the province Murcia, Spain. Sampling was done at mine waste heaps near and in the open pit Brunita, in the open pit Gloria, and at a rock formation with an outcropping pyrite vein. More than 340 samples have been taken and analyzed for numbers of leaching bacteria, microbial activity, and occurrence of sulfur compounds. Besides, physicochemical factors such as pH, temperature, redox potential etc. were measured. The statistical evaluation of all these data yielded the following results: (i) The area is a typical habitat for acidophilic leaching bacteria. Consequently, a significant positive correlation between the proton concentration and the numbers of leaching bacteria, the concentrations of leaching products (iron ions, sulfate, and elemental sulfur), and the microbial activity was obtained. (ii) The temporal development of leaching activity indicates that in this climate biological leaching needs about 5 years to almost fully oxidize all metal sulfides exposed at the surface (0-30 cm). (iii) Furthermore, bioleaching is restricted to the humid season from November until April.

Keywords: AMD, open pit mining, mining waste, monitoring, microcalorimetry

1. INTRODUCTION Huge abandoned mining areas with waste heaps, open pits, and acidic lakes are the

unpleasant but often tolerated side effect of the metal winning activities of modern times. Because these sites are often exposed to air and water and in most cases still hold significant amounts of metal sulfides they tend to pollute rivers and groundwater with acidic heavy-metal-containing solutions. This contamination, generally attributed to as acid mine or rock drainage (AMD, ARD), is caused by the activity of acidophilc iron(II)- and sulfur-oxidizing bacteria [1]. Therefore, risk assessment studies and rehabilitation projects are only possible if information is provided on the parameters which determine the bioleaching process. In particular, data on leaching kinetics and the overall temporal development of the microbial flora in a special mining area have to be collected before

* Present address: UFZ Center for Environmental Research Leipzig-Halle, Permoserstr. 15, D-04318 Leipzig, Germany


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starting rehabilitation activities. To meet these prerequisites we have studied for nearly 10 years the changes and interdependencies of leaching-relevant parameters in the abandoned mining district between Portman and La Union in the southeast of Spain (province Murcia). Sampling campaigns were performed at waste heaps near and in the open pit mines Brunita and Gloria as well as at natural leaching sites (Fig. 1). The ore-rich region near the city of Cartagena has suffered from human mining activities for thousands of years. Thanks to the excellent natural harbor known as Portman (a contraction of the Latin Portus Magnus) at the Mediterranean coast the richest ore deposits of the hills surrounding this bay were already mined by the Romans for silver and lead. Intensive exploitation by surface extraction technologies started in the late 1950s by S. M. M. Penarroya. During the next 30 years waste materials from several open pit mines, ore processing, and smelters were deposited as heaps or dumped in the bay of Portman [2]. The disposal in the Mediterranean Sea was stopped in 1990, leaving the bay completely filled with 54 million tons of heavy-metal-containing mining waste [2,3]. Finely, metal-winning activities ceased with the closure of the open pit mine Los Blancos III in 1991 and the closure of the Pb smelter near Cartagena in March 1992 [4].

Figure 1. Mining area between La Union and Portman in the Cartagena mountain range (Autonomous Region of Murcia, southeast Spain). The dotted area marked the abandoned mining district with the arrowhead pointing at the position of the open pit mines Brunita and Gloria

2. MATERIALS AND METHODS

2.1 Sample sites and sample analyses The investigated area between La Union and Portman (Fig. 1) belongs to a geological

formation that contains Fe-, Pb-, Ag-, Cu-, Zn-, Mn-, and Sn-rich ore deposits [5]. Sampling campaigns were done at 4 sites: (a) A waste heap near La Esperanza to the north of the open pit mine Brunita, which was piled up with material out of the mine as well as


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with flotation residues; (b) the abandoned open pit mine Brunita; (c) a natural leaching site at a rock formation with an outcropping pyrite vein, about 300 m to the south of Brunita; (d) the abandoned open pit mine Gloria. Analyses of leaching parameters were performed as described previously [6]. Briefly, sample humidity, temperature, pH, redox potential, calorimetric activity, and conductivity were recorded. Furthermore, concentrations of polythionates, sulfate, elemental sulfur, and iron ions were determined. For the assessment of the aerobic microbial flora acidophilic iron(II) oxidizers, acidophilic sulfur oxidizers, acidophilic chemoorganotrophs, neutrophilic sulfur oxidizers, and neutrophilic chemoorgano-trophs were counted by MPN enumeration. In order to correlate the data of sample analyses with the climatic situation in the province Murcia long-time mean values of temperature and precipitation for San Javier were applied [7]. On the basis of this data we were able to distinguish two seasons: “winter” covers the humid period from November to April and “summer” covers the dry period from May to October.

2.2 Statistical methods In a first statistical treatment the samples were compared with respect to all leaching

relevant parameters after dividing them in groups that had been classified with respect to sample sites (see 2.1) and sample types (solid surface samples or from a depth of 20 to 30 cm, solid samples from the pyrite vein, lake sediment samples, and water samples). As the collected data showed no normal distribution only non-parametric (i. e. distribution-free) statistical tests could be applied. Multiple and bivariate group comparisons were performed with the Kruskal-Wallis and Mann-Whitney U-test, respectively. In order to elucidate the relationship between microbial activity and in situ conditions non-parametric, bivariate correlation analyses were performed (Kendall S-test). In this way obtained results were tested with partial correlation analyses. Changes in the leaching processes during the total campaign period were further tested by multiple linear regressions. In this case, climatic influences were also considered by using the season at which the samples were collected as a covariable. Generally, the significance level was set for 5% and was adjusted to the number of statistical tests performed (Bonferroni´s adjustment). The software SPSS 10.0.7 was used for the statistical analyses.

3. RESULTS AND DISCUSSION

3.1 Microbial activity From previous studies it is already known that the microcalorimetrically measured

activity is one of the most reliable parameter in order to characterize a certain leaching biotope [6,8]. A comparison of the microcalorimetric activity with the other parameters recorded for the samples from the La Union mining district confirms this finding, too. The mean activity of all samples was highest within the first 3 years of the campaigns (1991 to 1993) and then decreased continuously (Fig. 2). This temporal development of the leaching activity can be explained with the closure of most open pit mines in the late 1980s and early 1990s. Obviously, in this mining area it takes about 5 years to almost fully oxidize the metal sulfides exposed to the surface (0-30 cm). Samples which showed a correlation between calorimetric heat production and cell numbers of leaching bacteria had activity values strongly depending on the season at which the samples were collected. As expected, the calorimetric activity was high in the humid winter season and low in the dry summer time (Fig. 3A). In other words, significant biooxidation of metal sulfides at the surface of the leaching sites (0-30 cm) only occurred between November and April. In typical leaching biotopes the microbial processes are dominated by acidophilic bacteria.


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Consequently, calorimetric values correlated negatively with the pH of the samples (Fig. 3B).

Figure 2. Temporal development of the microcalorimetrically measured activity of samples collected from 4 different leaching sites at the La Union mining district (sampling depth 0-30 cm)

Figure 3. Dependency of microbial activity of samples from the La Union mining district (sampling depth 0-30 cm) on season (A) and pH (B)

3.2 Group comparisons and correlation analyses of other parameters The interdependencies of the various leaching parameters were evaluated by group

comparisons and bivariate correlation analyses (see 2.2). Generally, it could be confirmed that all sampling sites are typical acidophilic leaching biotopes (data not shown). Although the total number of samples was comparatively low and the sampling sites consisted of very heterogeneous materials a positive correlation was found between the proton concentration and numbers of acidophilic leaching bacteria (iron(II)- and FeS2-oxidizing bacteria). In contrast, but in conformity with the previous finding, the proton concentration correlated negatively with cell numbers of neutrophilic bacteria (thiosulfate-oxidizing and chemoorgano-heterotrophic microorganisms). As the leaching activity was highest in


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acidic samples (Fig. 3B) the amounts of leaching products (iron ions, sulfate, and elemental sulfur) showed a highly significant correlation with the proton concentration.

4. CONCLUSIONS From the results of this study it can be concluded that rehabilitation projects should

start concomitantly with or immediately after mining, if climatic conditions are comparable to the La Union district. In this region leaching bacteria oxidizes surface-exposed metal sulfides (0-30 cm) almost completely within 3 to 5 years. Consequently, AMD/ARD can only be stopped if mitigation measures accompany the piling of waste heaps and the other depositing techniques. On the other hand, it is clearly shown that the simple closure of open pit mines is not an adequate solution for the problem. The consequences of this inactivity are acidic heavy-metal-contaminated soil and surface water. This contamination is build up in a few years but will leave an abandoned mining area unsuitable for many years with respect to farming and urban development as well as for tourism.

ACKNOWLEDGMENTS The help of J.I. Manteca (Universidad Politécnica de Cartagena) in providing detailed

information about the mining district La Union and the help of R. Buchert (University Hospital Eppendorf, Hamburg) with the statistical analyses are greatly acknowledged.

REFERENCES 1. H. L. Ehrich, Geomicrobiology, Marcel Dekker, New York, 2002. 2. J. M. M. Orozco et al., Landscape Urban Plann., 23 (1993) 195. 3. C. Auernheimer and S. Chinchon, Environ. Geol., 29 (1997) 78. 4. S. Moreno-Grau et al., Atmos. Environ., 34 (2000) 5161. 5. I. S. Oen et al., Econom. Geol., 70 (1990) 1259. 6. A. Schippers et al., Appl. Environ. Microbiol., 61 (1995) 2930. 7. http://www.inm.es/wwc/html/dclimat/SAN_JA.html 8. W. Sand et al., Appl. Microbiol. Biotechnol., 40 (1993) 421.



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An AFM-study on the adhesion of Acidithiobacillus ferrooxidans and Leptospirillum ferrooxidans to surfaces of pyrite

K. Kinzlera, W. Sanda*, J. Telegdib and E. Kalmanb a Institut für Allgemeine Botanik, Mikrobiologie, Universität Hamburg,

Ohnhorststrasse 18, D-22609 Hamburg, Germany b Hungarian Academy of Sciences, Chemical Research Center,

Pusztazeri ut 59-67, H 1025 Budapest, Hungary

Abstract The attachment of Acidithiobacillus ferrooxidans and Leptospirillum ferrooxidans to

pyrite surfaces has been analyzed using atomic force microscopy (AFM). Cells of both species attach preferentially to sites with visible surface defects. Although a part of the cells from the inoculum was still planktonical, a complete coverage of the pyrite surface never became detectable within 24 hours of attachment experiments. Only after the onset of growth, the surface became fully EPS-covered and, thus, turned into an "organic" surface. The extracellular polymeric substances (EPS) fill the space between the bacterial cell (wall) and the substratum surface. Furthermore, the EPS extend beyond the cell body. Especially the strain of L. ferrooxidans exhibited strong EPS-production and, consequently, an enormous surface coverage. Due to this coverage, the cells have an enlarged metabolic radius-of-action. Whereever EPS are in contact with the surface, the metal sulfide dissolution will be enhanced, due to the EPS-complexed iron(III)ions.

1. INTRODUCTION Bioleaching is the bacterially mediated dissolution of metal sulfides to sulfuric acid

(sulfate) and metal cations. This process is known for decades and has been shown to be of industrial importance as well as of environmental significance (acid rock drainage = ARD) [1,2,3,4]. Since its discovery there has been a debate about the mechanism, by which the microorganisms dissolve metal sulfides. Generally, a direct and an indirect mechanism are proposed. The argument pro and contra continue still nowadays [5,6,7].

Recent work of our group indicated that bioleaching is a process, which combines characteristics of both mechanisms [8,9,10]. This means that the dissolution of a metal sulfide is a chemical process caused by oxidative attack of iron(III)ions and/or protons. The iron(III)ions may be involved in any dissolution of metal sulfides, whereas protons are restricted to those metal sulfides, which are amenable to a hydrolytic dissolution. The group of metal sulfides, which obligately need iron(III)ions for a dissolution, comprise pyrite (FeS2), molybdenite (MoS2), and tungstenite (WS2). Their characteristic is the fact that the outer electron orbitals result from the metallic part of the compound, and not from the metal-sulfur bond. Thus, the oxidative attack by iron(III)ions is a requirement. The other metal sulfides have outer orbitals from the sulfur-metal bond and, thus, are attackable by protons and by iron(III)ions. In the presence of the latter the rate of


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dissolution may become considerably enhanced, which often causes erroneously the statement of a direct mechanism [11].

Furthermore, bioleaching is caused mainly by attached cells [8,10]. The planktonic population plays a negligible role under most conditions. Only in case of a high concentration of dissolved iron(III)ions planktonic cells could contribute. The iron(III)ions are provided by the leaching bacteria complexed in their extracellular, polymeric substances (EPS). The complex most likely consists of two moles of glucuronic acid and one mole of iron(III)ions. Both, At. ferrooxidans and Leptospirillum ferrooxidans, have these glucuronic acid - iron(III)ion complexes in their EPS. By an electrostatic interaction between the negatively charged pyrite surface (pyrite assumes a negative charge at pH below 3) and the net positive charge of the iron(III)ions-containing EPS, surface attachment of the cells becomes possible as the primary event in adhesion. Whether in later stages additional factors like pili etc. function, remains an open question. Some indications exist that such interactions may contribute also to the formation of robust surface attachment. Another open question is the site for attachment, whether it is a randomly or a deliberately chosen place. Some considerations seem to suggest that in case of bioleaching bacteria random attachment does not play the major role. Especially chemotaxis seems in this context to be a very important surface site recognizing system [12,13].

The present work was executed to visualize the attachment behavior of cells of two of the main bioleaching bacteria, At. ferrooxidans and L. ferrooxidans. The idea was to find criteria to judge the importance of EPS for attachment, to clarify the attachment mechanism, and to obtain information about the characteristics of the attachment site. For this purpose pure cultures of these species, which previously had been selected for good attachment properties, and coupons of selected pyrite cubes were incubated jointly. The coupons were analyzed for changes of the surface morphology in time-dependent adhesion and biofilm formation experiments using atomic force microscopy (AFM). AFM, a variety of tunneling scanning microscopy, allows to record the height profiles of plain surfaces by a surface scanning mechanism, and to put this together to produce a topographical map – AFM-images – of the investigated sample. Since this scanning action can be performed in air or water, living cells may be investigated without any interference by vacuum etc. Thus, the big advantages of this technique are the simple sample preparation and the measuring conditions at atmospheric pressure and room temperature. Consequently, for an investigation under natural conditions, AFM is the method of choice. The application of AFM for monitoring of microorganisms has been described in several papers [14,15,16].


2.1 Strains For these investigations the strains At. ferrooxidans strain R7 and L. ferrooxidans

strain R3 were selected. They are kept in our culture collection in nutrient solution at pH 1.9 [17].

2.2 Pyrite sample preparation Pyrite was obtained from mineral shops as cubes of at least 1 cm³ volume. To obtain

suitable samples for AFM investigations, the cubes were cut by a diamond saw into plates of about 2 mm thickness. Afterwards, the plates were washed with ethanol and hydrochloric acid to remove any organic contaminants and/or oxidation products like iron(II/III)ions, elemental sulfur, or polythionates. This procedure was followed by


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sterilization in 70% ethanol for 12h. These plates were introduced into the flasks with the bacterial cultures.

2.3 AFM experiments For experimental purposes, 500 mL cultures were grown, harvested by centrifugation

after two-thirds of the iron(II)ions had been oxidized, and resuspended in 100 mL nutrient-free solution ([17], without iron(II)ions). After a period of 24h for substrate depletion (to allow the cells to oxidize any remaining substrate), pyrite samples were introduced into the cultures and incubated for periods of up to 96h with stirring. At different times pyrite samples were removed and analyzed using the AFM. The atomic force microscope (NanoScope III, Digital Instruments, Veeco, USA) was operated in air in contact mode. A silicium nitride tip (SiN4) was used for surface scanning. All figures contain images, which have been obtained without any processing except autoflattening.

3. RESULTS AND DISCUSSION A sterile pyrite surface prior to any attachment experiment is shown in Figure 1. The

various layers besides minor surface defects are clearly visible on the (typical) pyrite surface.

Figure. 1. AFM-image of a sterile pyrite surface with a typical layered surface structure and two inorganic crystals

2.1 Attachment of At. ferrooxidans strain R7 to pyrite Attachment experiments were performed with At. ferrooxidans and immersed pyrite

plates. The following figures show the patterns of surface colonization by cells of At. ferrooxidans R7. For evaluating the question for preferential attachment, especially images in the time range of 2 to 8 h are useful. Such an image is given in Figure 2. Images obtained lateron yielded only equivocal information due to growth effects. It becomes obvious from Figure 2 and similar images (not shown) that the cells attached preferentially to sites with visible surface defects. More than 80% of all cells were attached to distortions of the pyrite surface (grooves, channels, edges of layers etc.) Images obtained at a later stage of the surface colonization process prove the growth of the attached bacteria (Fig. 3). Figure 4 demonstrates the existence of pits underneath bacteria. After removal of the cells, pits with the shape and size of cells become visible. Besides the bacteria footprints of cell-free EPS may remain on the surface (Fig. 5). Typical patterns of time-dependent surface colonization are presented in Figure 6A-D. Within 6 h several


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cells have attached to the pyrite surface. The amount of attached cells increases up to 24 h. Afterwards, until 48 h the amount of attached cells remains almost similar. One week later, the surface morphology exhibits considerable alteration. Only a few single cells remain visible. The majority is embedded in copious amounts of an amorphous mass, the EPS (measurements in the friction mode of the AFM proved that the amorphous mass is soft, whereas the free pyrite surface remained rough). The EPS cover a considerable part of the total surface. In Figure 7 a scheme of a surface-attached, bioleaching cell of At. ferrooxidans is shown.. It combines all our evidence collected from previous work [8] and from AFM images. Most important is the finding that the EPS extends over the limits of the cellbody and in this way considerably enlarge the metabolic radius-of-action.

Figure 2. AFM-image of a pyrite surface after 8 h of incubation with Acidithiobacillus ferrooxidans strain R7. Cells attached preferentially to sites with visible surface defects

Figure. 3. AFM-image of a pyrite surface after 24 h of incubation with Acidithiobacillus ferrooxidans strain R7. Pit formation is visible in the centre of the image

Figure 4. AFM-image of a cleaned pyrite surface after 24 h of incubation with cells of Acidithiobacillus ferrooxidans strain R7

Figure 5. AFM-image of a single cell of Acidithiobacillus ferrooxidans strain R7 attached to pyrite after 24 h of incubation. Around the cell cell-free exopolymers (footprints) are visible


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Figure 6A-D. Typical patterns of time dependent surface colonization by Acidithiobacillus ferrooxidans strain R7 over the course of 8 d

3+ Fe2+S0

S O2 32-

attack

CellA cell of Acidithiobacillus

A ferrooxidans

Figure 7. Schematical graph of a cell of Acidithiobacillus ferrooxidans attached to pyrite. Included is a model for the mechanism of the indirect or contact-mode leaching attack as catalyzed by A. ferrooxidans


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2.2 Attachment of Leptospirillum ferrooxidans strain R3 to pyrite The same experiments, as shown for At. ferrooxidans, were performed with cells of L.

ferrooxidans strain R3. In Fig. 8 an image with a few cells of L. ferrooxidans attached to a pyrite surface is shown. In the middle of the image a dividing cell or two recently divided cells are visible, which are jointly embedded in their EPS. The cells are surrounded by copious amounts of EPS. The amount of EPS is considerably higher than the amount of the EPS of At. ferrooxidans strain R7 (Fig. 5). Furthermore, the EPS cover a considerably larger part of the surface than the naked cells would be able to. The surface coverage is in principle similar to the one shown in Figure 7 for At. ferrooxidans. Time-dependent adhesion is shown in Fig. 9A-C. In the course of the colonization a serious change of the surface morphology becomes visible. After 6 h of incubation (Fig. 9A) several cells have attached to the pyrite surface and excreted some EPS around the cells. After 24 h (Fig.9B) the morphology already has changed considerably. Single cells with their surrounding EPS are still identifiable, although the EPS starts to merge together. After 48 h (Fig. 9C) the surface of the pyrite is covered by an almost homogenous film of EPS (sea of EPS) with some embedded cells. Consequently, the structure of the pyrite is not visible anymore. The presence of the EPS again was proven by experiments using the friction-mode of the AFM (data not shown).

Figure 8. AFM-image of cells of Leptospirillum ferrooxidans attached to pyrite after 24 h of incubation

Figure 9A-C. AFM-image of cells of Leptospirillum ferrooxidans attached to pyrite after 6 (A), 24 (B) or 48 (C) h of incubation


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3. CONCLUSIONS Summarizing, it is obvious that an attachment of these bioleaching bacteria takes

place at selected sites. These sites have a surface defect, which most likely causes a high surface charge by electrical imbalance. This agrees well with the hypothesis of an electrostatic interaction mechanism between EPS and pyrite surface. It also becomes clear from these images that the EPS are always the compounds mediating the contact between a bacterial cell and pyrite.

Once the cells are attached, they start to grow and, within a few days, may cover the surface with large amounts of EPS. Consequently, the previously inorganic surface turns to an organic one. It remains to be elucidated, what this means for further attachment. Preliminary data indicate that this organic surface film hinders other bacteria to attach, even those of the same strain. In any case, the EPS with their complexed iron(III)ions are able to oxidatively attack the pyrite. Thus, the cells radius-of-action becomes considerably larger than the dimensions of an EPS-free cell would indicate. Especially for L. ferrooxidans, whose only substrate are the low-energy iron(II)ions, this enlarged radius-of-action might confer an improved environmental competitiveness.

REFERENCES 1. A. Schippers and K. Bosecker, Z. Angew. Geol., 48 (2002) 38. 2. D. Fortin, B. Davis and T.J. Beveridge, FEMS Microbiol. Ecol., 21 (1996) 11. 3. K.J. Edwards, T.M. Gihring and J.F. Banfield, Appl. Environ. Microbiol., 65 (1999)

3627. 4. N. Kuyucak, Min. Environ. Managem., 11 (2001) 12. 5. W.Sand, T. Gehrke, R. Hallmann and A. Schippers, Appl. Microbiol. Biotechnol., 43

(1995) 961. 6. M.P. Silverman, J. Bacteriol., 94 (1967) 1046. 7. F.K. Crundwell, In: V.S.T. Ciminelli and O. Jr. Garcia (eds.), Biohydrometallurgy:

Fundamentals, Technology and Sustainable Development, Elsevier, New York, 2001, Part A, p. 149.

8. W.Sand, T.Gehrke, P.-G.Jozsa and A. Schippers, Hydromet., 59 (2001) 159. 9. T. Rohwerder, P.-G. Jozsa, T. Gehrke and W. Sand, In: G. Bitton (ed.) Encyclopedia

of Environmental Microbiology, Vol 2, Wiley, New York, 2002, p. 632. 10. D.E. Rawlings, Annu. Rev. Microbiol., 56 (2002) 65. 11. O.H. Tuovinen, In: H.L. Ehrlich and C.L. Brierley (eds.) Microbial Mineral Recovery,

McGraw-Hill, New York, 1990 p. 55. 12. J. Acuna, J. Rojas, A.M. Amaro, H. Toledo and C.A. Jerez, FEMS Microbiol. Lett., 96

(1986) 37. 13. G.Meyer, T. Schneider-Merck, S. Böhme and W. Sand, Acta Biotechnol., 22 (2002)

391. 14. Y.F. Dufrene, Ch.J.P. Boonaert, H.C. van der Mei, H.J. Busscher and P.G. Rouxhet,

Ultramicroscopy, 86 (2001) 113. 15. Y.F. Dufrene, Micron, 32 (2001) 153. 16. J. Telegdi, Zs. Keresztes, G. Pálinkás, E. Kalman and W. Sand, Applied Physics A, 66

(1998) S639. 17. M.E. Mackintosh, J. Gen. Microbiol., 105 (1978) 215.



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An X-ray photoelectron spectroscopy study of the mechanism of microbially assisted dissolution of chalcopyrite

A. Parkera, C. Klauberb*, M. Stottb, H.R. Watlingb and W. van Bronswijka a A.J. Parker Cooperative Research Centre for Hydrometallurgy, School of Applied

Chemistry, Curtin University, P.O. Box U1987, Perth, WA 6845, Australia b A.J. Parker Cooperative Research Centre for Hydrometallurgy, CSIRO Minerals, P.O.

Box 90, Bentley, WA 6982, Australia

Abstract The acidic oxidative dissolution of samples of p-type chalcopyrite (Mt Isa Mines

flotation concentrate and massive Mt Lyell) via the iron- and sulphur-oxidising microbes Sulfobacillus thermosulfidooxidans (50°C) and Sulfolobus metallicus (65°C) has been examined using X-ray photoelectron spectroscopy. This enabled surface speciation, encompassing sulphides, disulphides, elemental sulphur, sulphate and various jarosites, during the main dissolution stages. When compared to the case for the abiotic dissolution of chalcopyrite conducted under similar chemical conditions, both the similarities and differences can be highlighted. Principal differences between abiotic and microbially assisted dissolution arise from the microbial tendency to oxidise elemental sulphur and the effects of microbial nutrient media toward the type of jarosite formed. The role of the surface sulphide dimer S2

2- in the sulphide oxidation has also been clarified.

1. INTRODUCTION The goal of complete copper recovery from chalcopyrite via low energy

hydrometallurgical processing is critically dependent upon the control of "passivating" overlayer structures. The body of work previously undertaken by the authors (Klauber et al., 2001; Klauber, 2003; Parker et al., 2003), concentrating on the abiotic chemistry of Mt Isa Mines (MIM) and Mt. Lyell p-type chalcopyrites, for acidic sulphate oxidative dissolution, shows related behaviour over the thermophile temperature range and describes in detail a surface specific mechanism for abiotic chalcopyrite dissolution. The typical behavioural traits are oxidative dissolution to produce iron and copper in solution, with the sequential oxidation of the sulphide to disulphide, elemental sulphur and sulphate. The latter three all contribute to the initial surface reaction phase. Eventually the sulphate dominates, with a thin ferric sulphate precursor phase forming. With continued dissolution, the ferric solution concentration reaches a point where massive hydronium jarosite layers precipitate, hindering further chalcopyrite attack. Note that varied jarosites (KFe3(SO4)2(OH)6) are possible with cation substitution of the K+ by H3O+, NH4

+ etc. The

* Corresponding author. PO Box 90, Bentley, WA 6982, Australia, e-mail: [email protected]


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initial ferric sulphate layer is observed to act as the crystallization precursor for jarosite precipitation. Iron control appears to be the key to chalcopyrite hydrometallurgy. Elemental sulphur plays a role but does not appear to be a significant candidate for hindered dissolution. Despite thorough investigation, no spectroscopic evidence was found for stable forms of the previously alluded to polysulphides (Parker et al., 2003). This work places the abiotic chemistry into context with bioleaching in the presence of two microbial strains. Particular aims were to examine the mechanism of sulphur oxidation further and jarosite composition as a function of microbially assisted leaching conditions.

2. EXPERIMENTAL

2.1 Microbial culture Sulfolobus metallicus (DSM 6482T) and Sulfobacillus thermosulfidooxidans (DSM

9293T) were obtained from the German Collection for Microorganisms And Cell Cultures (DSMZ). Sulfolobus metallicus was cultured in a nutrient medium containing (NH4)2SO4 (1.3 g/L), KH2PO4 (0.28 g/L), MgSO4

.7H2O (0.25 g/L) and S. thermosulfidooxidans in a nutrient medium containing (NH4)2SO4 (3.0 g/L), K2HPO4 (0.5 g/L), MgSO4

.7H2O (0.5 g/L), KCl (0.1 g/L), Ca(NO3)2 (0.01 g/L). Both media were adjusted to pH 2.0 with concentrated H2SO4. The media were autoclaved at 121°C (2 atm) for 15 minutes. Sulfolobus metallicus runs were also conducted in potassium free media where K2HPO4 was replaced with (NH4)2HPO4 (0.27 g/L) and (NH4)2SO4 was reduced to 1.0 g/L. Yeast extract (0.02% (w/v)) was added aseptically to both media (via passage through a 0.22 µm pore size membrane). The chalcopyrite concentrate (CuFeS2 88%, FeS2 4% and gangue minerals, mainly quartz, 8%) was wet sieved to a size fraction of +20 µm –35 µm, and then dried in an oven at 30°C. Residual xanthates were then washed from the concentrate (Klauber et al., 2001). The cultures were grown at 65°C (S. metallicus) and 50°C (S. thermosulfidooxidans) in an environmental incubator shaker operated at 180 rpm.

2.2 Leaching method The microbially assisted leaches for the concentrate and the massive chalcopyrite

were conducted in different experimental apparatus. The concentrate was leached in simple flask arrangements in an environmental incubator shaker. A 5 mL volume of solution containing S. metallicus or S. thermosulfidooxidans (late exponential phase, ~72 hours post-inoculation) was inoculated into a series of 250 mL Erlenmeyer flasks (one for each leach period), containing 95 mL of appropriate nutrient medium and 3.1 g concentrate.

Samples of the massive chalcopyrite were silver dagged to sample stubs, dry polished with 2000 grit carbide paper and dipped onto an overflowing leachant meniscus; surface tension ensured contact only with the chalcopyrite surface. The leachant cell, containing 25 mL of inoculum and 0.03% (w/v) yeast extract in 500 mL of nutrient media, was run for 12 hours with 16.0 g concentrate prior to exposure to the massive chalcopyrite.

2.3 Solution sampling and analysis

The pH and redox potential (Ag|AgCl) were measured periodically throughout the experiments. For the concentrate leaches, concentrations of ferrous ion, total sulphur, copper and iron in solution were determined after centrifugation to separate the solid material. Ferrous ion concentrations in solution were determined by spectrophotometry using the method of Wilson (1960). The total sulphur, iron and copper concentrations in


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solution were measured using Inductively Coupled Plasma - Atomic Emission Spectrometry (ICP-AES).

2.4 X-ray photoelectron spectroscopy At selected reaction intervals (12, 24, 36, 60 and 108 hours), the contents of a single

flask were filtered (0.22 µm pore size membrane) and then washed with 10 mL of acidified MilliQ water (H2SO4 to pH 2) to remove soluble metal species. A portion of the solid material retained on the filter was then pressed into a sample stub and transferred into the photoelectron spectrometer and cooled to ~150K to minimise sulphur volatilisation. For the massive chalcopyrite, the surface was examined at shorter intervals (0.5, 2, 24, 48 and 72 hours). These pre-mounted samples went directly into the spectrometer after rinsing with acidified MilliQ water (H2SO4 to pH 2). XPS spectra were obtained with either Mg Kα or Al Kα radiation (300 W) from a conventional source in a VG ESCALAB Mk II fitted with 5-channeltron electron detection. Spectra were run at base pressure of ~1x10-10 mbar, using 6 mm analyser slits and a fixed analyser pass energy of 20 eV. The S 2p, Cu 2p, Fe 2p, O 1s, N 1s, K 2p, C 1s, K 2p, valence band, Fe 3s and full survey regions were acquired in sequence. Methods of data reduction employed such as X-ray line function removal (source function and satellites), spin-orbit splitting removal (in the case of S 2p) and comments on the peak fitting are covered in detail in earlier papers in the series (Klauber et al., 2001; Klauber et al., 2003; Parker et al., 2003).


3.1 Solution conditions Metal species solution concentration and redox potential for the concentrate cases are

shown in figure 1 and are typical for the microbially assisted ferric acid leach conditions employed. As would be expected, Eh rises and solution ferrous concentration drops as the leachant activity and microbial numbers increase. Also note the low concentrations of soluble iron as compared to the leached copper (especially in the S. metallicus- assisted leach) indicating substantial jarosite formation. The pH remained constant at ~2 until extensive jarosite formation occurred. Note that a rate comparison to abiotic conditions is not simple. In abiotic dissolution the initially high ferric level contrasts to a low value in the microbial case and the microbial catalytic role of ferrous to ferric conversion is also absent.

3.2 Photoelectron spectra of the inorganic phases The matrix of four leach experiments (two microbes and two chalcopyrite types) over

a number of sampling periods and the acquisition of multiple spectral lines allows only a few of the spectra to be shown. For clarity, the composite XPS spectra have arbitrary shifts in the ordinate position to avoid overlap. Spectra at increasing time intervals are arranged from top to bottom. For thin surface reaction layers, exhibiting negligible static charge shifts, the binding energy (BE) scales are referenced to the underlying chalcopyrite lattice sulphide species S2- at 161.15 eV (Klauber et al., 2001) to correct for spectrometer drift. This is important for the S 2p analysis. Figures 2 and 3 show the typical (unprocessed) S 2p data for the initial stages of the oxidation process for the concentrate. Note that once deposition of thick non-conducting overlayers has occurred there is an onset of static charging (typically 5-7 eV to higher apparent BE values). This has not been corrected for in the figures so as to illustrate that critical layer formation. For example, the raw sulphate peak in figure 2 can be seen to shift from 169.1 to 175.7 eV.


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Figure 1. Solution concentration of copper and iron and Eh variation during concentrate leaching

Figure 2. Sulphur 2p spectra from concentrate leached in the presence of S. thermosulfidooxidans at 50°C for periods varying from 12 to 108 hours


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Figure 3. Sulphur 2p spectra from concentrate leached in the presence of S. metallicus at 65°C for periods varying from 12 to 108 hours in nutrient media with and without K+

Due to the difficulty in reaction rate control for the microbially assisted leaches, caution needs to be taken in comparing the observed times for surface phase formations. Qualitatively, the same surface phase components form on both the MIM concentrate and the Mt. Lyell massive samples, but they do not necessarily correlate in terms of time. This is a consequence of differences in the experimental and sample set-ups (high surface area concentrate versus low surface area massive). Even within individual repeat leach runs exhibiting the same sequence of phases, the onset time for a given part of the sequence could be at variance by hours. This is despite care in reproducing both inoculum activity and the chalcopyrite sample. Other differences between the concentrate and massive relate solely to particle size, e.g. the deposition of jarosite related layers. For the concentrate, the static charge shifts are sudden as the particles are small and such layers reduce electrical conductivity between particles and the particles and earth (i.e. multiple insulating layers). By comparison the same relative jarosite levels on the massive material (in fixed contact with earth) resulted in negligible shifts. This does not indicate a difference in chemistry.

As a diagnostic tool the Cu 2p spectrum is of limited value as all the spectra (e.g. figure 4) are typically representative of Cu[I] for the unoxidized chalcopyrite lattice, diminishing in intensity as the depositing overlayers attenuate the photoemission. The Fe 2p spectrum (figure 5) is more sensitive as iron species participate in the formation of coherent overlayers, starting with pre-cursor iron sulphate acting as a nucleation site for jarosite deposition. Analysis of the Fe 2p states via peak fitting is not practical, though a least squares principal component analysis (Parker et al., 2003) can reliably indicate the overlayer composition of iron containing phases.


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Figure 4. Copper 2p spectra from concentrate leached in the presence of S. metallicus at 65°C for periods varying from 12 to 108 hours in nutrient media with and without K+

Figure 5. Iron 2p spectra from concentrate leached in the presence of S. metallicus at 65°C for periods varying from 12 to 108 hours in nutrient media with and without K+

The S 2p peak remains the most useful means of understanding the oxidation process. Figure 6 shows a detailed sulphur analysis of the initial oxidation in terms of states largely already established (Klauber et al., 2001, Klauber 2003, Parker et al., 2003). Once the state of reaction has progressed to significant jarosite formation, the reduced emission


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intensity from the mixed states region below the sulphate makes their analysis less reliable. The component that evolved at a higher BE than the sulphate had not been previously observed. It appeared to be characteristic of the high initial sulphate levels for the concentrate, but not the massive. It is most likely to be sulphate emission from nucleating jarosite crystals. This may be an illustration of differential static charging on the surface or a genuine BE shift for conversion from a simple sulphate to one characteristic of a jarosite.

Also noteworthy in figure 6, although the sulphide peak has been aligned to the lattice value of 161.15 eV, as the elemental sulphur evolves the measured sulphide feature actually shifts to a higher BE by ~0.13 eV whilst the sulphate BE remains approximately constant. At the same time the loss feature (Klauber, 2003) diminishes in intensity. Both this and the sulphide shift may be real and may point to a breakdown of the near surface chalcopyrite lattice. There are two factors influencing observed BE. Based on the Muliken charge population analysis (Hamajima et al., 1981) the sulphide donates charge toward the iron, so a lattice breakdown to fully ionic S2- should cause a shift to lower BE. Countering that, the simultaneous breakdown of the conduction band would mean less relaxation and a shift to higher BE.

Figure 6. Detail of the sulphur chemical state analysis for processed spectra showing the variation in states during the initial leaching. For figure clarity neither the final fits nor the residuals are shown


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Comparison of the S 2p at 60 hours with the S. thermosulfidooxidans to that for abiotic dissolution at 50°C after 2 hours (Parker et al., 2003) shows a very similar structure. Whilst the abiotic response appears "faster" (due to a higher initial ferric concentration), the microbially assisted creates more surface sulphate and appears to allow a more concerted attack on the sulphur dimer species by the ferric. The likely reason for this is the direct microbial attack on the elemental sulphur, a pathway absent in the abiotic case.

3.3 Photoelectron spectra of the entrained biomass Any entrained or adherent biomass will contribute to the observed photoelectron

signal, principally for the C, N and O 1s peaks and to a lesser extent P 2p. In this study it exhibited mainly in the C 1s and N 1s peaks; the O 1s was overwhelmed by the inorganic phases. In terms of a microbial cell wall characterization, the use of XPS has generally been limited to a cluster analysis based on nett N/C, O/C and P/C ratios (Van der Mei et al., 2000). This is pragmatic as it circumvents applying detailed peak analysis that can be especially difficult for complex organic systems. The current microbial studies do produce some differences to the prior abiotic chemical investigations. This is due to the biomass and more especially its nutrient media, e.g. ammonium ion presence. In this work, some knowledge of the expected biomass XPS spectrum, especially the N 1s, is needed to compare the leaching regimes, although the role of XPS in directly increasing microbe knowledge is limited.

Based on the generalized chemical compositions of cells (Neidhardt et al., 1996) the N 1s signal would be expected to primarily originate from proteins (~55%) with the RNA component also a major contributor (~21%). As the backbone structures for proteins are polypeptides, the basic N 1s signal for the protein should be similar to that observed for amide linkages R-CO.NH-. Purely from a photoelectron BE viewpoint (i.e. chemical environment), synthetic polymers such as nylon are polypeptide analogues. This provides a guide to expected BE values (Beamson and Briggs, 1992): polymer R C1s amide C 1s amide N 1s amide O 1s nylon 6,6 285.00 - 286.02 eV 288.02 eV 399.81 eV 531.37 eV nylon 6 285.00 - 286.00 eV 288.01 eV 399.77 eV 531.35 eV nylon 12 285.00 - 285.90 eV 287.97 eV 399.84 eV 531.33 eV

Whilst the side groups on the polypeptide backbone define the protein type, it would be expected that the amide nitrogen atom would dominate the overall N 1s signal. Hence a BE of about 399.8 eV should be expected (confirmed by an examination of Escherichia coli biomass). Carrying the synthetic polymer analogy further, the aromatic nitrogen atoms in the heterocyclic RNA bases would be expected to have a BE value about 399.3 eV, i.e. 0.5 eV below that of the amide nitrogen atoms (Beamson and Briggs, 1992). Hence two nitrogen peaks with a small separation would be expected for the biomass.

As can be seen from figure 7, two N 1s states are observed on the chalcopyrite sample. Although the ∆BE value of 1.9 eV is well above the expected amide-RNA separation, it perfectly matches the separation between that of ammonium jarosite standard and the biomass signature from E. coli. The jarosite was BE corrected for the sulphate component to match that of jarosite on massive chalcopyrite and for E. coli the C 1s was fixed to 285.0 eV. The higher BE component was also noted to vary disproportionately in relative intensity and found to dominate at long leach times with significant jarosite deposition. This additionally indicated a non-biomass origin. Based on the BE comparison, the low BE state can be identified as polypeptide amide and the


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higher BE state as the ammonium component in ammonium jarosite (the ammonium coming from the nutrient media). Curiously, the ammonium also is present much earlier in the oxidation. The signal-to-noise is insufficient to separate the protein-RNA components. Note also that Al Kα is preferred over Mg Kα radiation as the N 1s region is otherwise complicated by a series of copper Auger peaks.

Figure 7. Nitrogen 1s spectra from Mt. Lyell massive leached in the presence of S. thermosulfidooxidans at 50°C for 24 hours compared to E. coli and ammonium jarosite. Both raw and processed spectra are shown

3.4 Peak in elemental sulphur production Both the S. metallicus and S. thermosulfidooxidans assisted oxidative leaches exhibit

what appear to be peaks in the concentration of surface adsorbed elemental sulphur in the 36-60 hour range (figure 6 compared, for example, to 108 hours in figure 3). They are both known to be elemental sulphur oxidisers (Huber and Stetter, 1991, Golovacheva and Karavaiko, 1978) and also capable of oxidising ferrous so the nett behaviour will be controlled by the availability of these species. As elemental sulphur forms, some of the population oxidise elemental sulphur to sulphate. The specifics of the microbially assisted and abiotic leaching mechanisms are too complex to reduce to consecutive, opposing and concurrent reactions and also fit the limited rate data. We can, however, simplistically view the paths involving elemental sulphur as two consecutive steps:

S2- ferric⎯ → ⎯ ⎯ S° microbial

⎯ → ⎯ ⎯ ⎯ ⎯ SO42-

For the classic case of irreversible first order consecutive reactions, the intermediate species would first increase to a maximum and then fall to zero. Neither of these steps are first order, but taking the concentrations of (solid) sulphide as approximately constant and that of ferric as slowly varying, it will mimic first order. Similarly, for approximately constant microbial activity, the oxidation of the elemental sulphur will mimic first order. This would predict a peak in elemental sulphur concentration.

For the abiotic case sulphate is proposed to form via a thiosulphate intermediate (Parker et al., 2003). Thiosulphate is known to be unstable at temperature in the presence of sulphuric acid disproportionating to elemental sulphur and sulphate (Mizoguchi et al.,


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1976). By comparison with the microbial case the abiotic case does not have a consumption route for elemental sulphur.

3.5 The disulphide intermediate in elemental sulphur production Analysis of the sulphur 2p states for both the 12-60 hour S. thermosulfidooxidans and

12-36 hour S. metallicus leaches (figure 6) shows an interesting pattern in the relative concentrations of the disulphide dimer S2

2- and elemental sulphur (assumed to be S8). We have previously proposed the role of the pyritic-like dimer as an intermediate in the chalcopyrite oxidative dissolution by ferric ions (Parker et al., 2003). From earlier work the dimer appears to be a consequence of the fracturing of a chalcopyrite lattice and thus systemic to the chalcopyrite system (Klauber, 2003). Interestingly, the detailed S 2p state analysis for S. metallicus and S. thermosulfidooxidans (repeated for S. metallicus) shows a clear interrelationship between S2

2- and S8. This is shown in figure 8. At the 36-hour point when Eh and ferric concentration rise, there is a substantial decrease in the S2

2- concentration and a stoichiometrically commensurate increase in the S8, i.e. the combined sulphur content of the two states remains constant. It is the increased ferric activity that converts S2

2- to S8. The concentrate’s small pyrite content complicates this interpretation, but the changing levels of dimer and the subsequent quantity of elemental sulphur are well in excess of what could be expected from the pyrite content. Moreover, preliminary acidic ferric leaches on a pyrite concentrate showed no evidence for adsorbed elemental sulphur being formed. Though the pyrite is reactive, oxidation is straight to sulphate and, unlike the lower density dimer forming on chalcopyrite, appears not to be able to form elemental sulphur under these conditions.

Figure 8. Decrease in surface dimer concentration is mirrored by a commensurate increase in adsorbed elemental sulphur, indicating the dimer to be the intermediate

3.6 Cation role in jarosite formation Prior abiotic work, in the absence of alkali cations, has indicated that eventual

deposition of hydronium jarosite, initiated by a ferric sulphate surface phase, hinders dissolution (Parker et al., 2003). For the microbially assisted cases, the presence of potassium enables potassium jarosite to form. Removal of the K+ content from the nutrient


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medium not only prevented potassium jarosite, but actually appeared to delay a jarosite onset. This then occurred as the alternate ammonium jarosite (Stott et al., 2001) and is qualitatively supported by the ammonium phase being observed in larger amounts, based on the higher BE N 1s component, in the absence of K+.

Though only qualitatively examined, it would seem reasonable that the control of singly charged cations may offer alternate avenues for jarosite control in addition to employing ferric control. The relative sensitivity to cation type has not been explored, though any mechanism that alters the final jarosite may also alter its structural coherence and possibly encourage spontaneous nucleation rather than formation as an inhibiting layer.

4. CONCLUSIONS In comparing earlier abiotic work on the acidic oxidative dissolution of samples of p-

type chalcopyrite (Mt Isa Mines flotation concentrate and massive Mt Lyell) with that in the presence of oxidising microbes S. thermosulfidooxidans and S. metallicus, at 50°C and 65°C we have been able to affirm that the interface chemistry is driven by the chemical conditions. Ηowever, the microbial activity and environment can have an impact upon that chemistry. The microbial oxidation of the elemental sulphur is a key to the peak in elemental sulphur concentration observed and appears to allow a more concerted attack, by the ferric ion, on the surface sulphur dimer than in the abiotic case. Nutrient media composition is also important. Due to the high concentrations of ferric, sulphate and monovalent cations at elevated temperatures, microbial leach solutions strongly favour jarosite formation. The potassium free nutrient media delayed the onset of potassium jarosite, with the less stable ammonium jarosite forming at later reaction times. Controlling the formation of jarosite and jarosite type remains the key in eliminating the "passivation", or hindered dissolution, of chalcopyrite with respect to copper release, in both abiotic and microbially assisted systems.

REFERENCES 1. Beamson, G., and Briggs, D., 1992. High Resolution XPS of Organic Polymers, The

Scienta ESCA300 Database. John Wiley & Sons, Chichester. 2. Golovacheva, R.S. and Karavaiko, G.I., 1978. A new genus of thermophilic spore-

forming bacteria, Sulfobacillus. Mikrobiologiya, 47, 658-665. 3. Hamajima, T., Kambara, T., Ken, I. and Oguchi, T., 1981. Self-consistent electronic

structures of magnetic semiconductors by a discrete variational Xα calculation. III Chalcopyrite CuFeS2, Phys. Rev. B: Condens. Matter 24, 3349-3353.

4. Huber, G., and Stetter, O., 1991. Sulfolobus metallicus, sp. nov., a novel strictly chemolithoautotrophic thermophilic archaeal species of metal-mobilizers. System. Appl. Microbiol., 14, 372-378.

5. Klauber, C., Parker, A., van Bronswijk, W. and Watling, H., 2001. Sulphur speciation of leached chalcopyrite surfaces as determined by X-ray photoelectron spectroscopy. Int. J. Min. Process., 62, 65-94.

6. Klauber, C., 2003. Fracture induced reconstruction of a chalcopyrite CuFeS2 surface. Surf. Inter. Anal. (in press).

7. Mizoguchi, T., Takei, Y. and Okabe, T., 1976, The chemical behaviour of low valence sulfur compounds. X. Disproportionation of thiosulfate, trithionate, tetrathionate and sulfite under acidic conditions. Bull. Chem. Soc. Jpn., 49, 70-75.


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8. Neidhardt, F.C., Curtiss, III, R., Ingrahamn J.L., Lin, E.C.C., Low, K.B., Magasanik, B., Reznikoff, W., Riley, M., Schaechter, M. and Umbarger, H.E., 1996. Escherichia coli and Salmonella typhimurium: Cellular and Molecular Biology, 2nd Ed. American Society of Microbiology, Washington D.C.

9. Parker, A., Klauber, C., Kougianos, A., Watling, H.R. and van Bronswijk, W., 2003. An X-ray photoelectron spectroscopy study of the mechanism of oxidative dissolution of chalcopyrite. Hydromet. (in press).

10. Stott, M.B., Watling, H.R., Franzmann, P.D. and Sutton, D.C., 2001. The effect of solution chemistry on jarosite deposition during the leaching of chalcopyrite by the thermophilic archaeon, Sulfolobus metallicus. In: Ciminelli, V.S.T. and Garcia Jr., O. (eds.), IBS’01: Biohydrometallurgy: Fundamentals, technology and sustainable development: Proceedings of the International Biohydrometallurgy Symposium IBS’01 (Ouro Preto, Brazil), Elsevier Science B.V. Amsterdam, A, pp. 207-216.

11. Van der Mei, H.C., de Vries, J. and Busscher, H.J., 2000. X-ray photoelectron spectroscopy for the study of microbial cell surfaces. Surf. Sci. Rep. 39, 1-24.

12. Wilson, A.D., 1960. The micro-determination of ferrous iron in silicate minerals by a volumetric and a colorimetric method. Analyst (Lond.), 85, 823-827.



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Analysis of chalcopyrite (CuFeS2) electrodes utilizing galvanic current in the presence of Acidithiobacillus ferrooxidans

D. Bevilaquaa*, A.V. Benedettib, C.S. Fugivarab, O. Garcia Jr.a† a Department of Biochemistry and Chemical Technology, Institute of Chemistry, São

Paulo State University, P.O. Box 355, Araraquara,SP-14.801-970, Brazil b Department of Physical Chemistry, Institute of Chemistry, São Paulo State University,

P.O. Box 355, Araraquara, SP 14.801-970, Brazil

Abstract The oxidative dissolution of chalcopyrite electrodes by Acidithiobacillus ferrooxidans

was studied utilizing a bacteriology battery, consisting of an electrochemical cell with two identical compartments in two different conditions: (a) a solution in a natural condition without any air or gas forced flux through the cell (non-air saturated) and (b) the solution in the two compartments was saturated in air and a flux of air was passed over the solution during the experiment. One compartment was sterile and the other inoculated with A. ferrooxidans. During the experiment time course (around 10 days) the potential and galvanic current was measured in a zero resistance ammeter. Results showed that the forced aeration caused a progressive increasing in the values of potential and current due the continuous bacterial activity, while a mineral dissolution limited by depletion of oxygen in the solution was observed.

1. INTRODUCTION The microbiological leaching of chalcopyrite (CuFeS2) is of great interest because of

its potential application to many CuFeS2-rich ore materials. At present, copper bioleaching processes are more efficient to other copper sulfides such as covellite, bornite, etc., because their leaching rates are much more favorable as compared to those of CuFeS2. The slow dissolution rate reflects the recalcitrance of chalcopyrite to the chemical and microbial oxidation [1, 2]. The slow kinetics of chalcopyrite dissolution is the result of the intrinsic thermodynamic properties of this mineral. The electrochemical behavior of chalcopyrite and a number of other sulfide ores have been investigated [3, 4, 5]. However the biological mechanisms and the reactions that come into play are still poorly understood.

The purpose of the present work was to study the oxidation of a research-grade chalcopyrite electrode by A. ferrooxidans monitoring changes in the rest potential and in the galvanic current during exposure of the sulfide to the bacterial attack. After the end of

* Financial support from the Fundação de Amparo à Pesquisa do Estado de São Paulo (FAPESP) for the project and for fellowship to D.B. is gratefully acknowledged. Acknowledgments are also due to the Conselho Nacional de Desenvolvimento Científico e Tecnológico (CNPq) (O.G. J. and A. V. B.) † Corresponding author ([email protected])


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experiments, the electrodes were also analyzed by scanning electronic microscopy in order to verify the bacterial behavior regarding its colonization pattern on the sulfide. This study is a part of a project designed to elucidate the fundamental basis and underlying mechanisms involved in the microbiological leaching of sulfide minerals.

2. MATERIAL AND METHODS

2.1 Bacterial strain and growth conditions Acidithiobacillus ferrooxidans strain LR was used in this work [6]. The culture was

grown in a mineral salts medium [7] at pH 1.8 using ferrous sulfate as energy source. The cell suspension for electrochemical experiment was obtained after growth for 48 hours in a shaker (150 rpm and 30ºC) by successive washing and centrifugation (5000·g) to eliminate residual ferric ion from the medium [1]. The cells were then twice washed in Milli-Q water (18 MΩ.cm) and suspended in 10 mL of the mineral salts solution, medium T&K [7], without ferrous iron in Milli-Q water. The cell suspension collected aseptically was standardized by the modified Lowry protein determination method [8].

2.2 Mineral sample and electrode preparation Research-grade chalcopyrite was obtained from Ward’s Natural Science

Establishment (Rochester, N.Y.). Samples were cut in pieces of approximately 1 cm2 using a diamond saw. One face was polished using silicon carbide emery paper with different particle size and finally polished with an alumina suspension of 0.3 µm. To eliminate impurities the samples were rinsed with acetone, ethanol and Milli-Q water (15 min each one in an ultrasonic bath), and then dried with pure Argon and kept in desiccators before using.

2.3 Electrochemical measurements A modified Tait type electrochemical cell consisting of two compartments, each one

containing a massive electrode of chalcopyrite similarly prepared as possible was used in the galvanic current and potential measurements. The working electrodes were fixed at the bottom of each compartment of the cell with a Viton O-ring and the reference electrode was fixed at the top of the cell. An area of 0.28 cm2 of each working electrode was exposed to the solution, which was maintained at 30°C. The electrodes and the electrochemical cell were sterilized by UV exposure for 2 hours before each experiment.

The electrolytic solution (10 mL) composition was (g.L-1): 0.4 (NH4)2SO4, 0.4 K2HPO4, 0.4 MgSO4

.7H20 at pH 1.8 adjusted with H2SO4 and sterilized by autoclaving (1 atm, 120°C for 20 min). After a certain immersion time, one compartment was inoculated with A. ferrooxidans-LR (≈ 8.5 109 cells.mL-1) and the other one was used as a control (non-inoculated). Both compartments were connected via a glass tube containing the electrolyte solution, and a cellophane membrane was fixed at the bottom of the glass tube to impede the contamination of the control compartment with bacteria. The galvanic current was measured at zero resistance ammeter by a short-circuiting the auxiliary and reference electrodes outputs of the potentiostat, and imposing a potential difference of 0 ± 0.3 mV between the two mineral samples used for recording the current fluctuations. Ag/AgCl/KClsat electrode connected to the solution through a Luggin capillary was used as reference. A quiescent solution was used in two conditions: (a) a solution in a natural condition without any air or gas forced flux through the cell (non-air saturated) and (b) the solution in the two compartments was saturated in air and a flux of air was passed over the solution during the experiment (air saturated). The experiments were conducted for at least


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10 days, and then the electrodes were rinsed thoroughly with Milli-Q water to remove any remaining unattached bacteria and examined using a scanning electron microscope (SEM) Topcon SM-300.

3. RESULTS AND DISCUSSION The Figure 1 shows the potential (E) and the galvanic current (Ig) vs. time curves

obtained for chalcopyrite electrode in the presence of A. ferrooxidans-LR in non-air saturated condition. A decrease of the potential probably related to the dissolution of oxides, naturally formed on the chalcopyrite surface was observed.

After addition of the cell suspension the potential increased up to 250 mV as a consequence of biofilm formation and remained almost constant (around 250 mV) up to the fifth day (region 1 of the figure). Afterward the potential decreased until 150 mV, the minimum value, and increased again up to around 250 mV (region 2). The minimum observed in the E vs. t curve can be associated with different concurrent processes: the bacterial-assisted dissolution of iron from the mineral, and depletion of oxygen in the non-air saturated condition by the bacteria (cathodic reaction), leading to a decrease in the bacterial activity with a consequent increase in the potential (region 3). This situation can be illustrated by the global reaction of chalcopyrite oxidation by A. ferrooxidans, as shown bellow: 2CuFeS2 + 8½ O2 + H2SO4 → 2CuSO4 + Fe2(SO4)3 + H2O

In a previous work, relevant iron dissolution was also detected only after 4 days of inoculation [1].

The galvanic current decreased initially and stabilized at around 1 nA up to the fifth day and then increased reaching a maximum (at about 10 nA) which corresponds to minimum potential (region 2), indicating an increasing in the mineral dissolution. Increasing the immersion time the current decreased and stabilized at around 2 nA from the eighth day to the end of the experiment. The same explanation given for the potential-time curve is valid.

Figure 1. Potential and the galvanic current versus time curves obtained for chalcopyrite electrode in non-air saturated condition. Arrows indicate the time of bacterial addition into the electrochemical cell


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The experiment with air saturated solution (condition b), the E vs. t curve (Figure 2) showed different behavior when compared with non-air saturated condition (Figure1). The potential increased about 30 mV during the first hours of immersion and no significant changes were observed until the end of the experiment, stabilizing around 340 mV. This potential-time profile can be associated with the presence of oxidative atmosphere over the solution during all the time course of the experiment, being the oxygen supplied to the solution by a diffusion process. The minimum potential and the maximum current values observed for non-aerated condition (Figure1) were not detected in the aerated experiment. The oxygen diffusion controlled process was previously observed using electrochemical impedance spectroscopy [9].

Figure 2 also depicts the galvanic current-time curve and a slow and continuous increasing was observed, probably due to the progress of the bacterial activity, since there was no restriction of oxygen supplying.

Figure 2. Potential and the galvanic current versus time (Eoc vs. t and Igalv vs. t) curves obtained for chalcopyrite electrode in air-saturated condition. Arrows indicate the time of bacterial addition into the electrochemical cell

Figure 3 (A and B) shows the adhesion of bacteria on chalcopyrite surface at the end of experiment in air-saturated solution. As it can be seen mainly in Figure 3A a great number of cells of A. ferrooxidans-LR attached on the chalcopyrite surface involved by a biofilm, probably formed by exopolymeric substances. In Figure 3B (a zoom of Fig. 3A) an unusual massive presence of diplobacilli can be observed, which is probably related with an intense metabolic activity. In the non-air saturated experiment a similar pattern of adhesion was observed (data not shown) despite of electrochemical differences detected between both conditions.

The E and Ig curves of chalcopyrite electrodes in the presence of A. ferrooxidans-LR showed significant differences depending on the test condition. For air-saturated experiment the dissolved oxygen promoted a continuous activity of bacteria whereas in non-air saturated test this activity seems to be limited, as revealed by the current peak after around four days of incubation.


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(Α) (Β)

Figure 3. Scanning electron microscopy of the A. ferrooxidans-LR on the surface of chalcopyrite electrode in the experiment with air saturated solution at the end of electrochemical measurements (12 days). The bars indicate the magnification

REFERENCES 1. D. Bevilaqua, A. L. L. C. Leite, O. Garcia Jr, O. H. Tuovinen, Process Biochem., 38

(2002) 587. 2. C.L. Brierley, Crit. Rev. Microbiol., 6 (1978) 207. 3. T. Cabral and I. Ignatiadis, Int. J. Miner. Process, 62 (2001) 41. 4. P. Velásquez, D. Leinen, J. Pascual, J. R. Ramos-Barrados, R. Cordova, H. Gómez, R.

Schrebler, J. Electroanalytical Chem., 494 (2000) 87. 5. P. Velásquez, H. Gómez, D. Leinen, J. R. Ramos-Barrados, Colloids and Surface A,

140 (1998) 177. 6. O.Garcia, Jr., Rev. Microbiol., 20 (1991) 1. 7. O.H. Tuovinen and D.P. Kelly, Arch. Mikrobiol., 88 (1973) 285. 8. E.F. Hartree, Anal. Biochem., 48 (1972) 422. 9. D. Bevilaqua, I. Diéz-Perez, C. S. Fugivara, F. Sanz, A. V. Benedetti, O. Garcia Jr.

Proceedings of 15th International Corrosion Congress, Granada, Spain (2002) 063.



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Application of the bacterial weathering of silicate minerals in the improvement of quality of non-metallics

I. Štyriakováa, I. Štyriakb∗ a Department of Biotechnology, Institute of Geotechnics of the Slovak Academy of

Sciences, Watsonova 45, 05343 Košice, Slovakia, E-mail: [email protected] b Department of Microbiology, Institute of Animal Physiology of the Slovak Academy of

Sciences, Šoltésovej 4-6, 040 01 Košice, Slovakia

Abstract Silicate mineral weathering is one of the important biogeochemical processes in

nature. Heterotrophic bacteria of Bacillus spp. are often relatively active in these processes because they are limited due to the lack of organic nutrients as well as other inhibitory factors in environment. However, bacterial activity can be significantly increased by forming of optimal conditions in bioleaching processes (pH, T, Eh, concentration of biogenic elements). These processes can be used in the improvement of quality of several non-metallics.

Our results indicate that removal of iron from non-metallics is the first positive marker, which is visible after bioleaching. Except this, we observed the release of contaminants from intergranular spaces of silicate grains, where Fe bearing minerals are often impregnated. Moreover, the enrichment of non-metallics by fine-grained fraction is the result of destruction of silicate minerals (mica, feldspars) and rounding-off of their edges.

These findings suggest important applications of Bacillus strains in biotechnology involving the quality improvement of kaolin, quartz sands, zeolite and feldspars.

Keywords: non-metallics, deferritization, silicate minerals, Bacillus spp.

1. INTRODUCTION Weathering of rocks is biogeochemical process occurring at earth's crust. While both

laboratory and field studies have demonstrated that microorganisms can play a role in mineral weathering, the magnitude of this effect is still unknown.

Recent surveys of deep subsurface aquifers (hundreds of meters) show relatively high microbial numbers (approximately 105 to 107 cells/cm3). Microbial abundance and diversity is variable but does not decrease systematically with depth [1].

∗ Acknowledgments: The authors are grateful to the Slovak Grant Agency for Science (Grant No. 2/2107/22).


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The identity of members of communities of microorganisms in natural environments, their distribution, and the combined effects of their metabolisms on mineral dissolution are currently poorly understood [2].

Weathering of minerals in soil releases major nutrients such as K, P, Fe, Ca, Mg, Si, as well as trace ions which are necessary for microbial and plant growth [3]. Weathering action may be the result of oxidative or reductive attack of appropriately reactive mineral constituents (oxyhydroxides of Fe, Mn, sulphides) of a rock or mineral [4]. Silicate bacteria are very active (at pH 7) in the release of impregnated sulphides from the silicate matrix and sulphide minerals are accessible to destruction (at pH 4) by another bacterial species of acidophilic sulfur-oxidizing bacteria [5]. In the case of metal sulphide ore deposits, these activities can be and often are industrially exploited for metal recovery without smelting in a process called biohydrometallurgy.

Feldspars, mica and quartz are the most abundant silicate minerals at the earth's crust. While the geomicrobiology of silicate mineral weathering might seem an extremely arcane subject, consider the following short list of processes affected by biogeochemically mediated cycles: soil formation and plant nutrition, stability of architectural materials, geochemistry of groundwater and movement of contaminants, long term stability of geological nuclear repositories, effects of mineral weathering on climate on a geological time scale [2]. However, in the case of silicate minerals weathering, this activity can be also industrially used for improvement of quality of non-metallic raw materials and treatment of wastes.


2.1 Granitic eluvium Granitic eluvium (GE), derived from granitic rocks (Slovakia). It is composed of

quartz (55-60%), feldspar (40-35%) and mica (10-20%).

2.2 Quartz sands Quartz sands (QS) from Vyšný Petrovec deposit composed of quartz (80-90%), mica

(5-10%), kaolinite (3-8%), rutil (2-5%) and siderite (4-6%).

2.3 Kaolin sample The kaolin sample from Vyšný Petrovec deposit composed of quartz (4-7%), mica (8-

15%), kaolinite (60-75%), rutil (1-5%). The kaolin sample from Horná Prievrana deposit composed of quartz (70-80%), mica (6-14%), kaolinite (10-20%).

The iron-bearing minerals decrease the quality of these raw materials.

2.4 Bacteria and media Two bacterial strains of Bacillus spp. (Bacillus cereus and Bacillus pumilus) were

isolated from a kaolin quarry in Horná Prievrana. They were identified by means of the Becton-Dickinson microbiology system (Becton Dickinson, USA). For the species identification, the strains were cultivated on Columbia agar plates according to recommendation of the system producer. This panel contains 29 enzymatic and biochemical substrates and a fluorescence control on tips of plastic prongs. The resulting pattern of the 29 reactions is converted into a ten-digit profile number that is used as the basis for identification of a wide variety of microorganisms. For experiment, the bacterial strains were grown in Nutrient broth No.2 (Imuna) at 37°C for 18 hours. Bacterial cells were


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subsequently centrifuged at 4000 rpm for 15 min, subsequently washed twice with saline solution (0.9% NaCl) and added in a concentration of 1010 cells per ml to modified Bromfield liquid medium. Bioleaching of the samples was carried out in 3000 ml Erlenmeyer flasks containing 2000 ml of modified Bromfield medium (NaH2PO4 – 0.5g/l, MgSO4

.7H2O - 0.5g/l, (NH4)2SO4 – 1.0g/l, NaCl – 0.2g/l, molasses – 0.3g/kg) inoculated with a mixture of both Bacillus cereus and Bacillus pumilus strains. The flasks were incubated statically for 28-120 days at 28°C. The abiotic controls were cultivated under the same conditions. After incubation, the culture solutions were separated from the biomass by means of membrane filtration. The presence of vegetative bacterial cells in Erlenmeyer flasks and their morphology were regularly examined by light microscopy after Gram staining.

2.5 Chemical analyses Quantitative changes of samples (solid and liquid phases) investigated from view of

element composition stability were evaluated by standard analytical method – atomic absorption spectrometry on a VARIAN spectrometer AA - 30 apparatus (Varian, Australia) after dissolution of the samples by standard procedure.

3. RESULTS AND DISCUSSION The theses on weathering process as the most important geo-chemical process in the

hypergene zone that causes material–structural changes of rocks and minerals, essentially come out of the numerous studies, but there are still a lot of unclear issues concerning the weathering stages, sequence of rock changes and participation of microorganisms in genesis of minerals, including clay ones. We have started the study of biogenous environmental transformation processes on the basis of similar considerations, which were also formulated by Ján Čurlík [6], which concerned the fact that so far there is not any precisely substantiating theory that could explain the genesis of eluvial processes and creation of mineral zones of weathering crusts.

Microorganisms play an important role in the dissolution of silicate structure in the rock weathering process and in the genesis of clay minerals. The activity of heterotrophic bacteria and the release of organic compounds into circulating solutions in rock environment are displayed in nature by zonal formation of a deposit. In the porcelain clay deposit in Horná Prievrana, there is possible to observe the zones where the kaolin zone and the Fe oxidation zone with its typical mineralization are evident. The relicts of the bedrock of phyllites are also evident. The mire samples from Horná Prievrana kaolinite pit are characteristic also by a high number of bacteria, especially of Bacillus species. There are found bacteria of Bacillus genus with a wide range of species as B. cereus, B. pumilus a B. megaterium and B. mycoides, which coexist together. They are relatively active under aerobic and anaerobic conditions in environment with higher humidity [7].

The dissolution activity of selected Bacillus strains from these mire samples was investigated with non-metallics materials contained feldspars, quartz, kaolinite, mica and oxyhydroxides of iron. The chemical analyses as well as other methods indicated the ability of bacteria to destruct these silicates and to release iron and other elements. This fact suggests that these processes can be used in the removing of iron from many non-metallics.

3.1 Granitic eluvium containing feldspars Experimental studies of the effect of microorganisms on mineral weathering of

granitic eluvium with feldspars were performed in closed (batch) reaction vessels. In batch


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experiments, solutions were monitored over time, and solid phases were analysed in the end of experiment.

Feldspars in granitic eluvium (GE) sample contain 1,44 wt.% iron. Iron was found especially in the form of Fe oxyhydroxides. The presence of iron minerals was identified also by visible brown-red colour of this material. The releasing of Fe was characteristic by the change of visible brown-red colour of these non-metallic materials to brown-white. The Bacillus strains used in this study markedly decreased Si content (from 73,3 to 53,5%) and partially decreased also content of potassium (from 2,48 to 2,15%) after 120 days’ bioleaching what resulted to a weak enrichment of Al content (from 11.51 to 11.73%) and destruction of feldspars. After 120 days of bioleaching, 31% Fe extraction was detected in GE.

The release of Fe from GE suggests the removal of the cementing iron oxides from the silicate surface. This partial Si and Fe removal can also improve the access of organic acids to iron-bearing minerals and enable the increase of the feldspar raw material quality by subsequent leaching with less concentrated organic acids. That is why bacterial destruction of intergranular spaces of silicate grains, often impregnated with iron minerals, should be used as pretreatment of raw material before chemical treatment of the sample by oxalic acid.

Bobos and Gomes [8] investigated K-feldspars weathering at the kaolin deposit of Sao Vicente de Pereira (Portugal). Their analyses indicated that the weathering products were typical by high content of Al and small contents of Si, Fe, and K. Moreover, altered K-feldspar grains exhibit "porous" surfaces. This fact indicates that the K-feldspar has been rapidly dissolved or corroded by solutions with an acid pH, whereas no secondary mineral phases were produced.

Mica and feldspars are the main constituents of rocks which were weathered into secondary minerals, according to the environmental conditions. Feldspars (especially K-feldspars) are often considered as relatively stable, and phyllosilicates (mica or chlorite) are the main source of clay minerals, such as vermiculite or smectite. In humid tropical climates, both micas and feldspars are completely transformed into kaolinite and/or gibbsite, and iron oxides [9]. Hydrolysis is the process on which governs weathering. If intense leaching of silicates by water occurs, it may lead to dissolution of the structure, desilicification and precipitation of new secondary minerals (generally 1:1 clay minerals or oxides and hydroxides). This process is the main source of clays in tropical areas. Feldspars and micas experimentally weathered by water show pits in their structure in addition to formation of hydroxides of Al and Fe [10]. Weathering is inextricably bound to biological processes, for organisms inhabit a wide range of nichens in surface and subsurface environments and influence various mineral transformation reactions [11,12].

3.2 Quartz sands Quartz sands are largely used as a raw material in the glass industry, however, are

often associated with iron and titanium impurities that decrease their economic value and hinder their application. Iron impurities lower an important technological factor related to the degree of whiteness [13,14, 15].

In the first stage of bioleaching of the QS sample, there were removed visible Fe-bearing minerals coated over quartz grains. Moreover, the extraction of elements was observed due to the destruction of silicate minerals and intergranular spaces of brown polymineral grains with releasing of ferrihydrite sealing siderite nodules. Chemical analyses of the bioleached QS


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sample showed the slight increase of Fe2O3 content from 0.15 to 0.19% and FeO content decrease from 0.19 to 0.1% after 3 months bioleaching.

Bacterial destruction of these polycomponent grains (fraction size 0.1-0.8 mm) caused the formation of a brown fine-grained fraction of mineral particles under 0.1 mm of size. This brown fine-grained fraction with a high Fe-bearing mineral content was removed from QS sample by simple elutriation process. Chemical analyses of QS show 47% removing of iron by bioleaching and elutriation.

Upgrading of silica sand for the production of clear flat glass requires a partial removal of iron, typically present in the form of iron oxide as discrete particles or coatings on the surface of the sand particles. Traditional technology involves leaching with heated sulphuric acid and producing a large quantity of acidic leachate, which is costly to treat and discharge [16].

Other chemical methods consist of leaching with mineral acids, and treatment with reducers, such as sodium dithionite plus aluminium sulfate or sulphur dioxide plus aluminium powder. These bleaching methods are usually suitable for achieving a higher degree of iron removal but are more expensive, have complex operating conditions, and are environmentally hazardous [17].

Our experiments confirmed that heterotrophic bacteria of Bacillus spp. play active role at biodestruction as well as at deferritization of quartz sands. Therefore, their activity should be used in the improvement of quartz sands quality by a new progressive biotechnology.

3.3 Kaolin sample The release of Fe from mica or from oxyhydroxides was used as indicator of mineral

dissolution and beneficiation of kaolin quality in our previous study [18]. Kaolin raw material composed of 0.92% - 1.43% iron, which was found especially in

the form of Fe oxyhydroxides (free Fe). The presence of Fe oxyhydroxides was identified by visible brown-red colour of this material. However, a substantial part of iron in kaolin is bound in the aluminosilicate lattice.

The Bacillus strains removed about 43% of free iron under anaerobic conditions from kaolin sample of brown-red colour within 28 days. The removal of structural Fe ions from kaolin of white colour under similar conditions was not so efficient because only about 15% of Fe was removed within 28 days. This fact can also suggest more difficult availability of iron for bacterial removal because of probable binding of iron in mica.

Probable mechanisms involved in the biological removal of iron from elutriated kaolin were examined on kaolin sample from Vyšný Petrovec.

The prolongation of bioleaching time from 1 month to 3 months and thus longer production of organic acids and metal-complexing substances by Bacillus strains caused the increase of iron removal from mica. There was observed after 3 months the 52% extraction of Fe atoms from octahedral position in mica when Al removal was only about 2%.

Kaolin mined at Vyšný Petrovec deposit is economically interesting raw material but its properties could be improved by removal of unsuitable minerals (mica, oxyhydroxides). After such an improvement, it could be a suitable raw material for more extensive use in ceramic industry.

Several strains of bacteria released by this way the cations from biotite (Si, Fe, Al), plagioclase, and feldspar (Si, Al) much more than abiotic procedures [2]. Microbial


1034

production of organics by fermentation, or reductive dissolution of Fe – Mn mineral phases can greatly accelerate weathering rates of aluminosilicate minerals [19, 20].

4. CONCLUSION The iron content of industrial minerals can be reduced by a number of physical,

physicochemical and chemical methods. The appropriate method for the removal of iron from an industrial mineral depends on the mineralogical forms and the distribution of iron in the particular ore [21]. The red and yellow pigmentations noticed in many clay deposits are mainly due to the associated oxides, hydroxides and hydrated oxides of ferric iron such as hematite (red), maghemite (reddish brown), goethite (brownish yellow), lepidocrocite (orange), ferrihydrite (brownish red), etc. These oxides and hydroxides occur either as coatings on individual grains or as discrete fine particles throughout the clay mass. Quantities as low as 0.4% of ferric iron may be sufficient enough to impart colour to the deposit. Removal of these associated impurities improves the quality of the material. The beneficiated clay finds use in paper coating, paper filling and as extenders in paints and polymers [22].

Considerable efforts have been devoted to the problem of removing ferric contaminants by physical [23] and chemical means. High intensity magnetic separation is a standard method used, removing substantial quantities of iron and titanium as mineral impurities, consequently improving the brightness of silicates. The chemical methods consist of leaching with mineral acids, and treatment with reducers, such as sodium dithionite plus aluminium sulphate, sulphur dioxide plus aluminium powder, and sulphur dioxide plus zink powder. These bleaching methods are usually suitable for achieving a higher degree of iron removal but they are more expensive, have complex operating conditions, and are environmentally hazardous [17].

This paper also reports a possibility of microbial removal of iron from non – metallics and improvement of silicate properties via the action of bacterial destruction of Bacillus spp. Our results indicate the removal of visible iron coatings from non-metallics as well as iron contaminants from intergranular spaces of silicate grains. Moreover, the enrichment of non-metallics by fine-grained fraction is the result of destruction of silicate minerals (mica, feldspars) and rounding-off of their edges. It is also possible to accelerate bioleaching process by continual monitoring of sugars consumption from molasses, which was used as the energy source during samples bioleaching.

REFERENCES

1. D.J. Balkwill, Geomicrobiology J., 7 (1989) 33. 2. W.W. Barker, S.A.Welch, S. Chu and J.F. Banfield, In: Banfield JF, Nealson KH

(Eds) Geomicrobiology: Interactions between microbes and minerals, Reviews in Mineralogy, Vol. 35, Washington, D.C., USA, 1997

3. P. Hinsinger, B. Jaillard and J.E. Dufey, Soil Soc Am J, 56 (1992) 977. 4. H.L. Ehrlich, Chem Geol, 132 (1996) 1. 5. I. Štyriaková, I. Štyriak and M. Kušnierová, In: R. Amils, A. Ballester (eds) Process

Metallurgy 9A, Elsevier, 1999. 6. Čurlík,J., 1988: Geochémia geologických procesov. Hypergénne procesy.

Prírodovedecká fakulta Univerzity Komenského, Bratislava. 7. I. Štyriaková, I. Štyriak, Mineralia Slovaca, 34 (2002) 99. 8. I. Bobos and C. Gomes, Geologica Carpathica, 51 (2000) 49. 9. R. Romero, M. Robert, F. Elsass and C. Garcia, Clay Minerals 27, (1992) 21.


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10. M. Robert, D. Tessier, In: I.P. Martini and W. Chesworth (eds.): Weathering, Soils and Paleosols. ELSEVIER, 1992.

11. J.F. Banfield and R.A. Eggleton, Clays and Clay Minerals, 36 (1988) 47. 12. P.M. Huang and M. Schnityer, Soil Science Society of America Special Publication

no.17 (1986) 606. 13. L.G. Yan, Y.J. Yu, S.S. Song, H.L Nan, Y.L. Cheng, X.M. Li, Q.W. Kong and Y.M

Dai, Proc. 2nd World Congress on Non-Metallic Minerals, Beijing, China, 1987. 14. F. Veglio and L. Toro, International Journal of Mineral Processing, 39 (1993) 87. 15. F. Veglio, Hydrometallurgy, 45 (1997) 181. 16. I.I.Tarasova, A.W.L. Dudeney and S. Pilurzu, Minerals Engineering, 14 (2001) 639. 17. L.M.S. Mesquita, T. Rodrigues and S.S. Gomes, Minerals Engineering, 9 (1996) 965. 18. I. Štyriaková and I. Štyriak, Ceramics Silikáty, 44 (2000), 135. 19. S.A.Welch, W.J. Ullman, Geochim Cosmochim Acta 60 (1996) 2939. 20. P.C. Bennett, F.K. Hiebert and W.J. Choi, Chem Geol 132 (1996) 45. 21. M. Taxiarchou, D. Panias, I. Douni, I. Paspaliaris, and A. Kontopoulos,

Hydrometallurgy, 46 (1997) 215. 22. V.R. Ambikadevi and M. Lalithambika, Applied Clay Science 16 (2000) 133. 23. S.J.F. Guimares, N. De Oliveira, F.L. De Salles, Trans Inst. Min. metall. 51 (1987) 13.



1037

Assessment of acid production potential of sulphide minerals using Acidithiobacillus ferrooxidans and microbial sulphate

reduction using Desulfotomaculum nigrificans

Evvie Chockalingam, S. Subramanian, K.A. Natarajan and J.J. Brauna

Department of Metallurgy and aIndo-French Centre for Water Sciences, Indian Institute of Science, Bangalore – 560 012, India

Abstract The acid production potential of pyrite and chalcopyrite minerals has been assessed in

the presence of Acidithiobacillus ferrooxidans in iron-free Silverman and Lundgren medium. The parameters such as pH, redox potential, bacterial cell number, ferrous, ferric and copper concentrations were monitored as a function of time. The inoculated samples show significant changes in the values of the parameters studied relative to the control samples. The pH values decrease to 1.6 after 100 days for pyrite and to 2.5 after 150 days for chalcopyrite respectively, from an initial value of about 7, after inoculation with Acidithiobacillus ferrooxidans cells. The ferrous ion concentration is found to be higher vis-à-vis that of ferric for both the mineral samples. The kinetics of copper dissolution from chalcopyrite is found to be slower compared to the dissolution of iron. Significant sulphate reduction and copper precipitation could be achieved using Desulfotomaculum nigrificans, both at pH 7 and 5.5. An increase in the cell number enhances the kinetics of sulphate reduction and copper precipitation.

Keywords: acid production potential, Acidithiobacillus ferrooxidans, pyrite, chalcopyrite, Desulfotomaculum nigrificans

1. INTRODUCTION A major environmental concern to the mining industry, both during the operational

period and after closure is the generation of acidity, resulting from the oxidation of sulphide mineral wastes, catalyzed by acidophilic microorganisms. The dearth of basic or buffering materials to neutralise the acid, results in the leach water becoming acidic, leading to dissolution and build up of metal species. This phenomenon known as acid mine drainage, needs to be mitigated by suitable strategies. The technology employed for the treatment of acid mine drainage includes pH adjustment, chemical precipitation, aeration and settling. Several active and passive treatment methods, including the use of soil and water covers have been investigated to curtail the oxidation of sulphidic wastes [1]. Biotechnological approaches using sulphate-reducing bacteria (SRB) have also gained importance in recent times [2-4]. The treatment of metal-contaminated wastes by sulphate-reducing bacteria has been investigated by several workers [5-7]. Specific technologies that have been successful in using SRB to remediate ground water contaminated with heavy metals include the Paques’ (THIOPAQ) system [8,9] and the NTBC bio-sulphide


1038

process [10]. In an ongoing research programme, environmental impact studies have been undertaken with reference to an abandoned copper mine. Keeping these objectives in mind, acid production potential studies on pyrite and chalcopyrite minerals using Acidithiobacillus ferrooxidans have been performed. Sulphate reduction studies have been carried out in modified Barr’s medium using Desulfotomaculum nigrificans, in the simultaneous presence of copper ions, as a function of pH and bacterial cell number. The variation in the sulphate and copper concentrations was monitored as a function of time.

2. EXPERIMENTAL

2.1 Materials

2.1.1 Mineral samples Pure mineral samples of pyrite and chalcopyrite were obtained from Alminrock

Indscer Fabriks, Bangalore, India. The samples were subjected to dry grinding in a porcelain ball mill and the < 74 µm size fraction was used for the experiments.

2.1.2 Bacterial strains The strain of Acidithiobacillus ferrooxidans was isolated from the ore samples of the

Hutti gold mines, Karnataka, India, and used after purification. A pure strain of Desulfotomaculum nigrificans was obtained from the National Collection of Industrial Microorganisms, National Chemical Laboratory, Pune, India.

2.2 Methods

2.2.1 Medium and growth of Acidithiobacillus ferrooxidans The bacteria Acidithiobacillus ferrooxidans were grown in 9K medium [11] which

has the following composition in grams per litre of solution: (NH4)2SO4, 3.0; MgSO4

.7H2O, 0.5; K2HPO4, 0.5; KCl, 0.1; Ca(NO3)2, 0.1 and FeSO4.7H2O, 44.2. The pH

was adjusted to 2 initially by the addition of sulphuric acid. The fully-grown culture of Acidithiobacillus ferrooxidans from the 9K medium was centrifuged to harvest the cells. The cells were then resuspended in pH 2 sulphuric acid to remove iron. This procedure was repeated twice to get iron free cell suspension. The iron-free Acidithiobacillus ferrooxidans cells thus obtained were used as inoculum for the acid production potential studies.

2.2.2 Medium and growth of Desulfotomaculum nigrificans The bacteria Desulfotomaculum nigrificans were grown in the modified Barr’s

medium (MBM) [12] which contained 0.5 g K2HPO4, 1 g NH4Cl, 1 g CaSO4, 2 g MgSO4

.7H2O, 5 g sodium lactate and 1 g yeast extract in 1000 mL distilled water. The pH of the Barr’s medium was adjusted to 7.5 and sterilized for 3 consecutive days. The pH was checked every time before sterilization and adjusted to the desired value. Before inoculation, for each 100 ml of the medium, 0.05 g of FeSO4(NH4)2SO4

.6H2O and 0.01 g of sodium thioglycollate was added and filter sterilized. The media was stored in 250 ml serum bottles stoppered with a tight fitting rubber cork. To each 90 ml of the above-mentioned media, 10 ml of an active culture was added through a sterilized syringe. Immediately after inoculation, nitrogen gas was flushed into the medium for 5 minutes through a syringe and sealed tightly with parafilm to maintain anaerobic conditions. The bottles were then incubated at 37°C in a rotary shaker at 160 rpm. The fully-grown culture


1039

was centrifuged and the supernatant was discarded. The cells were resuspended in distilled water. This process was repeated twice and the washed cells were used as inoculum for the sulphate reduction studies.

2.2.3 Assessment of acid production potential The acid production potential of pyrite and chalcopyrite minerals was evaluated

adopting the following procedure: 5 g of the mineral sample (< 74 µm) was suspended in 100 ml of 9K- medium (without ferrous sulphate) containing ≈ 3 x 108 cells/ml under natural pH condition, in a 250 ml Erlenmeyer flask. The flask was agitated at 240 rpm in a Remi rotary shaker at 30 + 2°C. The parameters such as pH, ESCE, bacterial cell number, ferrous, ferric and copper concentrations were monitored periodically. A control sample was also maintained in which 10% alcoholic thymol was added as a bactericide.

2.2.4 Sulphate reduction and copper precipitation kinetics using Desulfotomaculum nigrificans

In these experiments, the variation in the sulphate and copper concentrations was measured at pH ≈ 5.5 and 7 using different cell densities (≈ 107, 108 and 109 cells/ml), during growth of Desulfotomaculum nigrificans in the modified Barr’s medium. 10 ppm of copper was added as cupric sulphate pentahydrate to the MBM for these studies.

2.2.5 Analytical procedures The bacterial cell count was determined using a Petroff-Hauser counter attached to a

Leitz phase contrast microbiological microscope (Laborlux K Wild Mps12). The ferrous and total iron concentrations were analysed using the orthophenanthroline method [13] in a Shimadzu model UV260 uv-visible spectrophotometer. The total iron concentration was determined by reducing the ferric iron in solution to ferrous by adding hydroxylammonium chloride. The ferric concentration was then estimated as the difference between total iron and ferrous concentrations. Copper was analysed using a Video 11E Thermo-Jarell Ash atomic absorption spectrophotometer. Sulphate concentration was determined by turbidimetric method using barium chloride at a wavelength of 420nm [14] using a Shimadzu model UV260 uv-visible spectrophotometer.


3.1 Acid production potential studies The acid production potential of pyrite and chalcopyrite in the absence and presence

of Acidithiobacillus ferrooxidans was assessed by monitoring the variation in the pH, redox potential, bacterial cell number, ferrous, ferric and copper concentrations as a function of time. It must be emphasized that these experiments were carried out at a natural pH of 6-7, in order to ascertain the propensity for acid generation, with or without bacterial cells.

3.1.1 pH The variation of pH as a function of time for pyrite and chalcopyrite suspensions in

9K- medium for the inoculated and control samples is shown in Figure 1. In the case of the pyrite sample, the pH steeply decreases from the initial value of ≈ 7 to ≈ 2 in about 50 days. The pH marginally decreases thereafter to ≈ 1.6 and remains more or less constant upto about 220 days. However, in the case of the control sample for pyrite, the pH marginally decreases to ≈ 6 in 50 days and is further lowered to about 4.5 in about 100


1040

days and remains at that value for the time period investigated. With respect to chalcopyrite, the pH steadily decreases from ≈ 6 to ≈ 2.5 in about 150 days and remains more or less unaltered thereafter. In the case of the control sample for chalcopyrite, the pH remains close to the initial value of 6 upto 90 days and is subsequently decreased to ≈ 4.6 in 150 days and remains at that value upto 220 days. These results highlight that acidity is generated in the case of both pyrite and chalcopyrite after inoculation with Acidithiobacillus ferrooxidans. On a comparative basis, the acid production potential of pyrite is higher than that of chalcopyrite after bacterial interaction.

3.1.2 Redox potential Figure 2 depicts the redox potential changes as a function of time for pyrite and

chalcopyrite suspentions in 9K- medium. The ESCE value increases sharply from 100 mV to ≈ 400 mV in about 40 days for pyrite after interaction with the bacterial cells. The redox potential further increases to ≈ 425 mV in 50 days and remains unchanged upto about 220 days. In the case of the control sample for pyrite the ESCE increases from the initial value of 100 mV to ≈ 180 mV in 50 days. Beyond that period, the redox potential marginally increases and attains a value of 190 mV. The redox potential of the chalcopyrite sample after inoculation with Acidithiobacillus ferrooxidans steadily increases from 100 mV to ≈ 250 mV in about 90 days. A sharp increase is observed to a value of 500 mV in 130 days beyond which there is not much of a change. With respect to the control sample for chalcopyrite, the redox potential is more or less unaltered at 100 mV upto about 90 days. Thereafter, a steady increase is observed upto about 180 days and the redox potential attains a value of 270 mV. As can be expected, the redox potential values are higher in the case of the inoculated samples attesting to bacterial oxidation of the mineral samples. It is pertinent to mention that the redox potential values observed in these tests are lower than that commonly encountered in typical leaching systems. This is understandable as the initial pH of the system was close to neutral (pH 6-7), in contrast to a typical leaching system, wherein the pH is maintained close to 2-2.5.

0 50 100 150 2000

2

4

6

8

10

pH

Time (days)

Pyrite in 9K- medium Pyrite in 9K- medium + Af cells Chalcopyrite in 9K- medium Chalcopyrite in 9K- medium +

Af cells

0 50 100 150 2000

150

300

450

600

750

900

E SCE

(mV)

Time (days)

Pyrite in 9K- medium Pyrite in 9K- medium + Af cells Chalcopyrite in 9K- medium Chalcopyrite in 9K- medium +

Af cells

Figure 1. pH as a function of time for pyrite and chalcopyrite mineral suspensions

Figure 2. Redox potential as a function of time for pyrite and chalcopyrite mineral suspensions


1041

3.1.3 Bacterial cell number The changes in the bacterial cell number after interaction with pyrite and chalcopyrite

mineral samples as a function of time are portrayed in Figure 3. For both the samples, a decrease in the cell number is observed initially, presumably due to the attachment of the cells onto the mineral substrates. In the case of pyrite, the cell number steeply increases from 2 x 107 cells/ml to 8 x 108 cells/ml in about 60 days and thereafter remains unchanged. For chalcopyrite the cell number marginally increases from 7 x 107 to 2 x 108 cells/ml after 220 days. It is interesting to note that there is an increase in cell number after interaction with pyrite while after interaction with chalcopyrite there is a marginal decrease. The dissolution of copper ions from chalcopyrite as a function of time could plausibly lead to toxicity impeding cell growth.

0 50 100 150 200 250107

108

109

1010

Num

ber o

f cel

ls/m

l

Time (days)

Pyrite Chalcopyrite

Figure 3. Bacterial cell number as a function of time for pyrite and chalcopyrite mineral suspensions

3.1.4 Dissolution kinetics of ferrous and ferric ions The changes in the ferrous and ferric concentrations as a function of time for pyrite

are shown in Figure 4a. The ferrous concentration steeply increases initially upto about 50 days and attains a value of ≈ 6g/l in the case of the inoculated sample. Around 170 days, the ferrous concentration attains a value of 7g/l and remains at that value upto 220 days. The ferric concentration steadily increases to ≈ 3g/l in about 50 days and thereafter remains unchanged in the case of the inoculated sample. With respect to the control samples, the ferric concentration slightly increases to ≈ 0.5g/l in about 75 days and remains unaltered thereafter. The ferrous concentration in the control sample increases to about 1g/l in 75 days and further increases to about 1.5g/l between 160-220 days. It is evident that the ferrous concentration is higher in both the inoculated and control samples compared to the ferric concentrations. These results are in agreement with those reported earlier [15-17]. It is also pertinent to mention that the presence of the pyrite substrate facilitates the preferential oxidation of sulphur in the mineral, rather than the ferrous ions in solution, through the direct contact mechanism.

The dissolution of iron species from chalcopyrite is portrayed in Figure 4b. Almost a similar trend as that observed for pyrite is revealed in this case also. In the case of the inoculated sample, the ferrous concentration shows a marginal increase upto about 60 days and subsequently a steep increase to 3.5 g/l is observed upto 160 days. Beyond that period,


1042

there is not much of a variation in the ferrous concentration. The ferric concentration gradually increases to about 1 g/l in 150 days and remains at that value for the inoculated sample. The ferrous concentration in the control sample marginally increases to about 0.75 g/l in 150 days and more or less remains unchanged thereafter. On the other hand, the ferric concentration in the control sample remains more or less constant at 0.25 g/l in the time frame investigated. A closer examination of Figures 4a and 4b highlights that the dissolution rate of iron species from pyrite is higher vis-à-vis that of chalcopyrite. The lower redox potential values observed in Figure 2 are consistent with the low ferric/ferrous ratios observed (Figs. 4a and 4b).

0 50 100 150 2000

2

4

6

8

10

12 Pyrite

Fe3+

con

c. (g

/l)

Fe2+

conc

. (g/

l)

Time (days)

Fe2+in 9K- medium Fe2+in 9K- medium + Af cells Fe3+in 9K- medium Fe3+in 9K- medium + Af cells

0

2

4

6

8

10

12

0 50 100 150 2000

1

2

3

4

5

6

7Chalcopyrite

Fe2+

con

c. (g

/l)

Time (days)

Fe2+ in 9K- medium Fe2+ in 9K- medium + Af cells Fe3+ in 9K- medium Fe3+ in 9K- medium + Af cells

0

1

2

3

4

5

6

7

Fe

3+ c

onc.

(g/l)

Figure 4a. Dissolution of ferrous and ferric from pyrite as a function of time

Figure 4b. Dissolution of ferrous and ferric from chalcopyrite as a function of time

3.1.5 Dissolution kinetics of copper ions Figure 5 depicts the kinetics of dissolution of copper from chalcopyrite in the case of

the inoculated and control samples. The amount of copper dissolved is very low in the first 50 days for both the samples. The copper concentration steadily increases to about 3 g/l in the case of the inoculated sample after 220 days. However, the concentration of copper remains unchanged upto about 170 days at 0.2 g/l and thereafter shows a marginal increase to about 0.5 g/l in the case of the control sample. It becomes of interest to note from Figures 4b and 5 that the dissolution rate of ferrous is higher compared to that of copper from chalcopyrite. Similar results have been reported by other workers [15].

0 50 100 150 200

0

1

2

3

4

5Chalcopyrite

Cu c

onc.

(g/l)

Time (days)

9K- medium 9K- medium + Af cells

Figure 5. Dissolution of copper from chalcopyrite as a function of time


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3.2 Studies on sulphate reduction and copper precipitation using Desulfotomaculum nigrificans Studies have been carried out to assess the efficacy of Desulfotomaculum nigrificans

to bring about the reduction of sulphate as well as the precipitation of copper. These studies are of relevance to the mitigation of acid mine drainage. The effect of pH and cell number on the reduction kinetics was assessed. The growth of the organism was carried out in the modified Barr’s medium [12] in the presence of copper ions (≈ 10 ppm). Figure 6 shows the change in sulphate concentration as a function of time and cell number. In these tests, the initial pH was maintained at 7. It is evident from the figure that the sulphate concentration decreases rapidly upto about 9 days in all cases and gradually thereafter. Further, the rate of sulphate reduction increases with increase in the cell number. For example, the sulphate reduction increases from 18.5% to 24.5% to 31.5% as the cell number is increased from ≈ 107 to ≈ 109 cells/ml.

The change in copper concentration as a function of time and cell number was also monitored. The results are depicted in Figure 7. It is noteworthy that, complete precipitation of copper takes place by 3, 4 and 6 days respectively as the cell number is increased from ≈107 to 109 cells/ml. It was found that the pH increases from the initial value of 7 to about 7.8 after 4 days in all the samples and subsequently decreased back to 7. This may be attributed to the production of H2S by bacterial metabolism.

0 2 4 6 8 10 12 14 16900

1050

1200

1350

1500

1650 Initial pH = 7 + 0.2

Sulp

hate

con

c. (p

pm)

Time (days)

1.4 x 107 cells/ml 1.1 x 108 cells/ml 1.6 x 109 cells/ml

0 1 2 3 4 5 60

2

4

6

8

10

12 Initial pH = 7 + 0.2

Cu

conc

. (pp

m)

Time (days)


Figure 6. Bioreduction of sulphate as a function of time and bacterial cell number at pH 7

Figure 7. Bioremoval of copper as a function of time and bacterial cell number at pH 7

The feasibility of sulphate reduction at pH 5.5 was ascertained and the results are portrayed in Figure 8. It is apparent that, beyond 7 days the amount of sulphate reduced, increases with increase in cell number. A comparison of Figures 6 and 8 indicates that the rate of sulphate reduction is lower when the pH is reduced to 5.5 from 7. Prolonged experiments lasting for 57 days carried out at pH 7 and pH 5.5 respectively resulted in 95% and 91% sulphate reduction using ≈ 109 cells/ml.

The variation of copper concentration as a function of time for experiments carried out at pH 5.5 at different cell densities is shown in Figure 9. It is evident that complete precipitation of copper takes place by 6, 8 and 15 days respectively as the cell number is increased from ≈107 to 109 cells/ml. As anticipated, the copper precipitation kinetics is


1044

slower at pH 5.5 when compared to pH 7 (Figures 7 and 9). The results presented in Figures 6 to 9 unequivocally demonstrate that Desulfotomaculum nigrificans is capable of significant sulphate reduction and complete precipitation of copper. The mechanisms of sulphate reduction and metal ion precipitation are well documented [18, 19].

0 2 4 6 8 10 12 14 16

1200

1350

1500

1650Initial pH = 5.5 + 0.2

Sulp

hate

con

c. (p

pm)

Time (days)


0 2 4 6 8 10 12 14 160

2

4

6

8

10

12

14Initial pH = 5.5 + 0.2

Cu

conc

. (pp

m)

Time (days)


Figure 8. Bioreduction of sulphate as a function of time and bacterial cell number at pH 5.5

Figure 9. Bioremoval of copper as a function of time and bacterial cell number at pH 5.5

4. CONCLUSIONS From the results of the present investigation, the following conclusions can be arrived

at: 1. Acid production potential studies reveal that the pH values of pyrite and chalcopyrite

suspensions are decreased from ≈ 7 to about 1.6 and 2.5 respectively after inoculation with Acidithiobacillus ferrooxidans cells.

2. An increase in the bacterial cell number is observed in the case of pyrite while for chalcopyrite there is a marginal decrease in the cell number.

3. In the case of both pyrite and chalcopyrite minerals, the dissolution of ferrous ion is substantially higher compared to ferric ion.

4. The amount of copper dissolved from chalcopyrite is lower than that of ferrous. 5. About 95% and 91% sulphate reduction could be achieved at pH 7 and 5.5

respectively, using ≈ 109 cells/ml of Desulfotomaculum nigrificans. 6. Complete precipitation of copper could be achieved both at pH 5.5 and 7 using

Desulfotomaculum nigrificans cells, though the kinetics of precipitation is slower at pH 5.5.

7. The kinetics of sulphate reduction and copper precipitation is found to increase with increase in the bacterial cell number.

ACKNOWLEDGEMENTS The authors are grateful to the Indo-French Centre for Promotion of Advanced

Research (CEFIPRA) and the Institute for Research and Development (IRD) France, for grant of research projects to carry out this investigation.


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Vol. 14, 9 (2001) 997.



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Comparative study on pit formation and interfacial chemistry induced by Leptospirillum and Acidothiobacillus ferrooxidans

during FeS2 leaching

Helmut Tributsch and José Rojas-Chapana

Hahn-Meitner-Institut, Dept. Solare Energetik, 14109 Berlin, Germany

Abstract After it had been assumed for a long time that the bacterium Leptospirillum

ferrooxidans is only capable of oxidizing Fe2+ which then oxidizes the sulfide, high-resolution electron microscopy has recently revealed that the bacterium can also selectively disintegrate a crystallized solid sulfide such as FeS2 in order to recover Fe2+. Characteristic properties are the mineralization of its surface including the interacting capsula, where sulfide nanoparticles can be identified, and localized pit formation, which can result in a pronounced collective effort to dissolve the crystalline environment.

Acidothiobacillus ferrooxidans, on the other hand, profits from weakening and break up of chemical bonds mediated by the formation of cysteine-pyrite complexes. This might also be the case under natural conditions by the excretion of cysteine and cysteine-like metabolites.

This contribution aims at a comparison of both Fe2+ oxidizing bacteria, to reveal their adaptation to the leaching of the FeS2 interface. It is shown how Leptospirillum has adapted electrochemical while Acidothiobacillus has adapted chemical mechanisms to obtain an optimum output of energy supplying Fe2+ and sulfur species respectively.

1. INTRODUCTION Leaching of pyrite has received considerable attention especially due to its relevance

for the hydrometallurgical recovery of gold but also because pyrite, which is a component in many sulfide ores, releases sulfuric acid during the oxidation process [1-12]. The bacteria typically found in the presence of pyrite under leaching conditions are Acidothiobacillus (formerly Thiobacillus) ferrooxidans (A.f.) and Leptospirillum ferrooxidans (L.f.). As their names indicates, both are able to oxidize Fe2+ to Fe3+ in the presence of oxygen, but while A.f. is also able to oxidize sulfide species, L.f. is apparently only able to oxidize Fe2+. Numerous studies have been performed to investigate the oxidation properties of these two bacterial species [13-14]. With respect to the bacteria’s ability for Fe2+-oxidation it has been documented that L.f. is able to utilize Fe2+ as an energy source at significantly, 200 mV, more positive oxidation potentials compared with A.f. [9]. Under such conditions, correspondingly less electrochemical energy is available for bacterial carbon dioxide fixation. In technical gold leaching installations with pyrite containing ores, L.f. are the bacteria which are doing the main job especially when


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the solutions contain high concentrations of Fe3+. This contribution aims at a comparative analysis of the two bacterial mechanisms for interacting with pyrite [13].

2. EXPERIMENTAL Bacterial strain and growth conditions: Leptospirillum ferrooxidans (isolated from

commercial biooxidation tanks in South Africa) was kindly provided by G. Hansford (University of Cape Town, South Africa). L.f. was maintained on ferrous ions culture solution using the medium described by Tuovinen and Kelly for Thiobacillus ferrooxidans [15]. Bacterially mediated ferrous oxidation was followed by increasing coloration and gradual precipitation due to the low solubility of ferric iron at pH 1.6.

For the culture of L. ferrooxidans on pyrite crystals (pyrite from Mogul-Turkey; particle size ca. 5 mm and at 10% pulp density) the same medium was used but without ferrous ions. Continuous agitation is required to maintain uniform pulp density.

The cultivation of A.f. on synthetic pyrite layers was performed following the procedure described by Rojas-Chapana et al. [16-17].

Microscopy: High-resolution and analytical transmission electron microscopy (TEM) and scanning electron microscopy (SEM) were used to study the bacterial morphology, the stoichiometry and shape of the pyrite grains used as substrates as well as bacteria-pyrite assemblages formed during the bacterial life cycle. All samples for TEM-examinations were imaged in a Philips CM 12 electron microscope operating at 100 kV. High-resolution SEM-examinations were imaged in a Gemini SEM-microscope (LEO1530). The SEM microscope was operated in the 3-7 KeV range being configurated primarily for imaging (using secondary electrons) and compositional analysis (using energy-dispersive x-ray spectroscopy (EDS)).

2.1 Bacterial morphology in pyrite cultures Neither A.f. nor L.f. show peculiarities in their morphology when cultivated on Fe2+ as

the only energy source. However, when cultivated on ground-up pyrite crystal material, they both show a significant adaptation, in which they develop a pronounced capsule of polysaccharides around their bacterial cells.

Fig. 1a and 1b show A.f. bacteria surrounded by such capsules, which are dotted with nanoparticles. These nanoparticles turned out to be molecular sulfur [16] extracted from the pyrite interface. Fig. 1b shows a A.f. cell which is splitting-up into a mother- and daughter cell. It is seen, that also the capsule is separated during this process. The mechanism by which A.f. is interacting with the pyrite surface, via the organic material of the capsule has been discussed in some details in previous publications [11,16,17]. Since the main energy source for this bacterium is a sulfide species that liberates significantly more energy compared with Fe2+, the evolution of such a capsule is not surprising. The bacterium needs an organic matrix to circle the catalyst species, which are disrupting chemical bonds in the pyrite interface. Crystal-bound sulfur is extracted in the form of polysulfide to yield sulfur colloids as intermediate energy storing elements. The concentration, size and distribution of these colloids within the capsule may change during the life cycle of the bacterium and in dependence of the sulfide energy supply.

It was a surprising discovery that L.f. are generating a capsule around their cells when interacting with a pyrite energy source. Fig. 1c shows three bacterial cells characterized by their spiral growth, partially overlapping each other in the transmission electron microscope picture. The two bacterial cells on the right are shown in a slightly higher magnification in Fig. 1d. The areas where the bacterial cells plus capsules overlap are


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clearly shown as areas of more intensive darkness. All over the capsule of L.f. dark spots of nanoparticles are seen. They are typically unevenly distributed and vary in a certain range with respect to their dimension. X-ray luminescence (EDAX) analysis has clearly shown that one is dealing with FeS2 particles. The cubic crystal structure of pyrite could be confirmed by electron diffraction analysis [18]. The question now arises: why does L.f., if it only oxidizes Fe2+, need a capsule? and, even more important; why do crystalline FeS2 particles accumulate in it? There is apparently a mechanism by which L.f. bacterial cells are interacting with solid pyrite interfaces to disrupt them and to recover nanoparticles for some metabolic purpose. In order to get a better insight into the interaction of L.f. with pyrite interfaces it became necessary to perform studies of crystallized pyrite interfaces in the presence of bacterial cells.

Figure 1a. Planktonic A.f. bacteria grown on synthetic pyrite film

Figure 1b. A.f. bacteria during cell division on synthetic pyrite film

Figure 1c. Planktonic L.f. bacteria grown on pyrite crystals

Figure 1d. Enlargement of 1c. L.f. bacteria covered with pyrite nanoparticles

2.2 Leaching patterns on crystallized pyrite interfaces Our research group has a long-term experience with the physical chemistry of pyrite

interfaces and their electrochemical behavior. It was, therefore, quite easy to identify a very unusual type of leaching patterns which develop in presence of L.f.. In otherwise apparently untouched surfaces, pronounced 20 – 40 µm large, deep etching structures developed, which are shown in Fig. 2.

They partially show a hexagonal pattern (which can be explained for pyrite interfaces [18]). The characteristic spiral-shaped cells of L.f. are rarely found on the exposed surface of pyrite. However, interestingly, when looking into the etching craters, they can be localized there in high concentrations. This is demonstrated with Fig. 2c. The bacterial


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concentration is so high, that the bacterial cells are touching each other. There appears to be a correlation between the pronounced etching pit formation and the high concentration of L.f. cells in them. This discovery raises new aspects in relation to L.f. activity. These bacteria have apparently developed a technique for actively dissolving pyrite. If they can do that, they presumably have an advantage from it regardless of the established notion that they are unable to oxidize sulfur species.

Figure 2a-c. SEM micrographs of a pyrite surface leached by L.f. (pitting). The bottom picture shows a high concentration of L.f. in the corrosion craters

It is reasonable now to confront this observed leaching behavior with the leaching patterns observed with A.f. for pyrite. The pyrite leaching mechanism by these bacteria has been examined before in some detail, especially by in-situ observation of bacterial activity on 100 nm thin FeS2 synthetic layers [11,12,16]. A localized collective activity comparable to that observed to L.f. was never observed. The bacterial cells moved freely around on the surface interacting with it and leaching it away. Very homogeneous interfaces were that way leached away quite homogeneously. When the substrate, however, was inhomogeneous or when surface-active agents like Tween 80 were applied in very low concentrations the bacterial cells developed localized activities, which yielded pronounced bacterial etching pits.

Such pits could be observed in transmitted light microscopic studies as shown in Fig. 3a. Such pictures have been published before [10,16,17], but they are shown here again to contrast the surface activity of L.f. The white spots are etching pits, which bacteria created, making the thin pyrite layer more transparent for light. Chains of such etching pits are clearly recognized and evidence a local multiplication of bacterial cells. High-resolution electron microscopical studies of such pits revealed that the nanocrystalline pyrite substrate has been chemically leached as shown in Fig. 3b. Here an etching pit is


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seen on the left hand side and the shadow of a bacterial cell on the right. As explained in preceding publications, such etching patterns arise when bacterial cells with their capsule interact with the pyrite interface. Fe3+, H+, or the addition of a complex forming thiol such as cysteine are cycled as catalysts to break-up interfacial chemical bonds of pyrite [19]. A.f. can use Fe3+ for oxidation of pyrite, however, this oxidation does not lead to an essential disruption of chemical bonds because the energy bands from which electrons are extracted have a non-bonding nature being d-states of iron. When, however, a complexing agent like cysteine is added, then the surface can be disrupted. This has been shown in complementary experiments in which pyrite layers were exposed to cysteine-containing solution. It could be shown that pyrite is indeed dissolved by this complex forming thiol.

Figure 3. Optical transmission photographs of 100 nm thin synthetic pyrite layers which are dotted with corrosion pits generated by bacterial activity. a) shows chains of corrosion pits indicating cell multiplication on the pyrite surface; b) shows a transmission electron micrograph of a corrosion pit with a bacterium nearby

It can, therefore, be understood why A.f. is able to dissolve pyrite. This bacterium is specialized for the recovery of the sulfide species and is, therefore, able to disintegrate the crystallized solid material. How can, on the other hand, a bacterium such as L.f., which is only able to oxidize Fe2+, attack a pyrite interface where electron extraction does not primarily break chemical bonds due to the non-bonding character of the energy states involved (within 1 eV from the valence band edge)?

2.3 Comparative evaluation of leaching mechanisms Since electrons extracted from the pyrite valence band by Fe3+ do not generate broken

bonds, a straightforward dissolution of pyrite by Fe3+ can, therefore, not be expected [10,19]. Only when the pyrite electrode is electrochemically polarized beyond the standard redox potential of Fe2+/3+

⎟⎟⎠

⎞⎜⎜⎝

⎛−= +

+

3

2

lnFeFe

zFRTEoEh (1)

an electrochemical dissolution of pyrite will occur. It is described by equation (2), which shows that water molecules are needed for this dissolution:

+−+− +++→+ HSOFeeOHFeS 162148 24

222 (2)

This has been verified by experiments in which small concentrations of water have been added to organic electrolytes. It is known from literature that L.f. can oxidize Fe2+ as an energy source at very high concentrations of Fe3+ corresponding to potentials, which are approx. 200 mV more positive compared to the iron oxidation abilities of A.f. [9-13]. L.f., which is able to oxidize Fe2+, however not sulfur containing species, may therefore be


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able to oxidize pyrite by locally inducing a very positive Fe2+/3+ redox potential as shown in Fig. 4a.

anodic

Fe3+

nanosized-FeS 2

e-

Fe2+

Fe2+

Fe2+

Fe3+

H+

H+

e-

e-

e-e-

e-

Fe2+

Fe2+

Fe2+ Fe3+FeS2

FeS + 2x +4O2 2

Fe + 2x +2SO2+ 2-4Fe3+

S O2 32-

Fe2+

dispersed sulfurnanoparticles

(X= H +, Fe ; (eg. cysteine))+ 3+ Y

FeS2

Acidothiobacillusferrooxidans

Fe2+

FeS -mineralized2

Leptospirillum ferrooxidans

BAbiofilm

cathodic

Figure 4. Schemes explaining the mechanisms for pyrite dissolution applied by L.f. (a) and of A.f. (b)

The resulting local anodic dissolution of pyrite will liberate pyrite nanoparticles, which are indeed found in the capsule of this bacterial species (Fig.1). The thus generated 14 electrons per pyrite molecule may be driven to a cathodic site on the pyrite surface, where they may be engaged in the reduction of Fe3+ species according to equation (3).

++− →+ 23 141414 FeFee (3) When equation (2) and (3) are summed up, equation (4) results which corresponds to

the equation proposed by Boone et al. [1-2] for ferric leaching of pyrite ores: +−++ ++→++ HSOFeFeOHFeS 16215148 2

423

22 (4)

While the same global mechanism results, the strategy, which had to be adopted by L.f. bacterial cells to accomplish this reaction turned out to be quite different from the mechanism originally assumed for ferric leaching. We can use standard thermodynamic considerations as applied to semiconductor electrochemistry [20] to understand the process. When the anodic dissolution process (2), which is proposed to be driven by bacterial activity is coupled with the reaction proceeding at the standard hydrogen electrode,

271414 HeH ⇔+ −+ (5)

the following global reaction results which for simplicity is assumed to be reversible.

242422 78 HSOHFeSOOHFeS ++⇔+ (6)

Knowing the ∆G° value (125,6 kcal/M) of this reaction, one can deduce the thermodynamic decomposition potential (against NHE). It is:

VzFGEo 39.0

0

+=+∆

= (7)

The decomposition may proceed if the redox potential of positive charges or holes is more positive than the decomposition potential (7), provided there are no kinetic barriers.


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Such kinetic barriers exist for pyrite. They arise because of the electronic d-character of its valence band. The position of the band edges varies if the electric charge on the surface varies, either by adsorbed ions, charged surface complexes or by high concentrations of holes in the surface. The increase of free energy of holes by electron extraction via L.f., which determine the oxidative dissolution of pyrite is described by

⎟⎟⎠

⎞⎜⎜⎝

⎛ ∆+−=

V

L

VLFp N

ppkTEE ln (8)

(EV = position of valence band edge, NV = concentration of states in the valence band, ∆PL

= concentration of holes generated by electron extraction through L.f. beyond the existing concentration p). In this relation Fe3+ mediated electron extraction leading to an increase in ∆P

L will pull the potential of holes towards positive values. Thus process will be accompanied by a positive shift of the valence band position EV due to positive charging of the pyrite surface. L.f. has obviously adapted to drive FeS2 beyond the decomposition potential and its kinetic barrier.

A strategy was necessary which resembles the phenomenon of electrochemical corrosion in which anodic and cathodic sites develop on the same micro heterogeneous interface. Such a scenario is shown for L.f. activity and pyrite in Fig. 4a. Here high concentrations of bacterial cells are locally inducing anodic electrochemical dissolution processes while the thus liberated electrons are moving to cathodic sites where Fe3+ is reduced. The Fe2+ then diffuses to the bacteria where the oxidation of this energy source will occur. Experimental support for such a process are the pronounced large corrosion craters produced by L.f. and the high concentration of bacterial cells sitting in these craters. This special surface morphology created by bacterial activity may allow an intensive circulation of Fe2+ for the bacterial energy supply needed for CO2-fixation.

While L.f. can sustain its metabolism by utilizing Fe2+ at very positive redox potentials, approx. 200 mV more positive compared to A.f., the latter species operates in a potential region where pyrite is typically not electrochemically dissolved. Since Fe3+ in small concentrations have only a very limited effect on disruption of chemical bonds in the pyrite interface, A.f. had to adopt a different strategy for energy recovery. This strategy is shown in a simplified way in Fig. 4b (for more detail see [19]. Basically this bacterium utilized protons, Fe3+, and complexing thiols (i.e extracellular polymeric substances (EPS) containing cysteine or cysteine-containing metabolites) for a catalytic disruption of pyrite interfaces. This works in such a way that protons interacting with the interface may partially break bonds forming interfacial SH- groups. Fe3+ may extract electrons from thus created interfacial states and cysteine may form complexes with the surface thus disrupting it and extracting iron sulfur aggregates bonded to cysteine. Since protons, Fe3+, species and cysteine are recycled this process may be considered catalytic. Viewed globally, the process developed by L.f. is also catalytic (since Fe3+ is recycled, Fig. 4a), however, the individual steps are electrochemical and require a separation of anodic and cathodic sites as known from electrochemical corrosion mechanisms.

The effect of exogenous cysteine on the bacterial dissolution of pyrite has been demonstrated [12], but so far no proof has been given that cysteine is actually present in EPS secreted by A.f. in sufficient concentration. Though the EPS composition of some determined A.f. strains have been published, it cannot be extrapolated to bacteria directly interacting with the sulfide substrate. In-situ determinations of cysteine and cysteine-containing metabolites in the presence of a biofilm on solid sulfides are required. In addition, cysteine and methionine have recently been found In EPS produced by sulfate-reducing bacteria during corrosion of mild steel [21].


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These distinct leaching strategies adopted by both pyrite oxidizing bacteria species show the evolutionary adaptability of bacteria to their inorganic energy source. Leptospirillum populations have apparently "discovered" that pyrite disintegrates under very positive Fe2+/3+ potentials and have improved their ability to recover reducing power for Fe3+ reduction by organizing themselves for cooperative leaching activities within corrosion craters which apparently improve their ability to recover Fe2+ species.

Summarizing, the following conclusions can be drawn: While A.f. takes advantage of thiol complex chemistry to brake up sulfur species, their

main energy source, from the sulfide interface, which are subsequently oxidized to sulfate for energy recovery, the survival interest of L.f. is aimed at recovering a maximum amount of reduced iron from the FeS2. Its interfacial activity aims at pushing the FeS2 towards electrochemical oxidation (iron sulfate formation) so that the reducing power of electrons can be harnessed for Fe3+ reduction elsewhere on the sulfide surface. L.f. bacteria apparently induce electrochemical corrosion by locally polarizing the FeS2 interface with Fe3+ immobilized in their capsule, thus oxidizing the sulfur species while generating microscopic cathodic currents for Fe3+ reduction.

The specific leaching behavior of both species is not adapted for an efficient joint leaching activity. L.f. aims at a more positive redox potential and at sacrificing the energy rich sulfur species while utilizing electrochemically mobilized electrons for Fe3+ reduction to recover a maximum amount of Fe2+. Since very little energy is generated in the oxidation of ferrous to ferric ions, the bacteria must oxidize large amounts of iron in order to grow. A.f., on the other hand, is adapted to extract the sulfur species for its sulfur metabolism, which provides a high energy turnover and Fe2+ oxidation is just a parallel process which yields significantly less energy but supports the interfacial leaching process by providing Fe3+.

ACKNOWLEDGEMENTS The authors would like to thank Ms U. Bloeck and Mr. M. Wilhelm for their valuable

assistance in electron microscopy.

REFERENCES

1. M. Boon and J. J. Heijnen, Biohydrometallurgical Technologies, A. E. Torma, J. E. Wey and V. L. Lakshmanan (eds.), The Minerals, Metals & Materials Society, Warrendale Pennsylvania, (1993) 217.

2. M. Boon and J. J. Heijnen, Hydrometallurgy 48 (1998) 27. 3. W. Sand, K. K. Rohde, B. Sobotke and C. Zenneck, Appl. Environ. Microbiol., 58

(1992) 85. 4. T. Rohwerder, A. Schippers and W. Sand, Proceedings of the XX International

Mineral Processing Congress, Vol. 4, Solid-Liquid Separation, Hydro- und Biohydrometallurgy, H. Hoberg and H. von Blottnitz (eds.), Gesellschaft für Bergbau, Metallurgy, Rohstoff- und Umwelttechnik (GMDB), Clausthal Zellerfeld, (1997) 475.

5. T. Rohwerder, A. Schippers, W. Sand, Thermochimica Acta, 309 (1998) 79. 6. P. R. Norris and D. P. Kelly, The use of mixed microbial cultures in metal recovery.

Microbial Interactions and Communities, A. T. Bull and J. H. Slater (eds.), Academic Press London, (1982) 443.

7. P. R. Norri, Recent Progress in Biohydrometallurgy, G. Rossi and A. E. Torma (eds.), Associazione Mieraria Sarda, Rom, Italy, (1983) 83.


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8. W. Sand, T. Gehrke, P.-G. Jozsa, A. Schippers, Proceedings of IBS 99. Biohydrometallurgy and the Environment Towards Mining of the 21st Century, Elsevier, (1999) 27.

9. G. S. Hansford and T. Vargas, Proceedings of IBS 99. Biohydrometallurgy and the Environment Towards Mining of the 21st Century, Elsevier, (1999) 13.

10. H. Tributsch, Proceedings of IBS 99. Biohydrometallurgy and the Environment Towards Mining of the 21st Century, Elsevier, (1999) 51.

11. J. Rojas-Chapana and H. Tributsch, Hydrometallurgy 59 (2001) 291-300. 12. J. A. Rojas-Chapana and H. Tributsch, Process Biochemistry 35 (8) (2000) 815. 13. D. Rawlings, H. Tributsch, G. Hansford, Microbiology 145 (1999) 5-13. 14. F. Battaglia-Brunet, P. d’Hughes, T. Cabral, P. Cezac, J. L. Garcia, D. Morin, Minerals

Engineering 11 (1998) 195. 15. O. Tuovinen, D. Kelly, Arch Mikrobiol 88(1973) 285. 16. J. Rojas-Chapana, M. Giersig, H. Tributsch, Arch. Microbiol 163 (1995) 352. 17. H. Tributsch, J. Rojas-Chapana, C. Baertels, A. Ennaoui, W. Hofmann, Corrosion 54

(1998) 216. 18. J. A. Rojas-Chapana and H. Tributsch, submitted to FEMS Microbiol. 19. H. Tributsch and J. A. Rojas-Chapana, Electrochimica Acta 45 (2000) 4705. 20. H. Gerischer, J. Vac. Sci. Technol.15 (1978) 1422. 21. K. Chan, L.. Xu, H. Fang. Environ. Sci. Technol. 36 (2002) 1720



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Composition of biofilm communities in acidic mine waters as revealed by combined cultivation and biomolecular approaches

Sakurako Kimura, Kris Coupland, Kevin B. Hallberg and D. Barrie Johnson

School of Biological Sciences, University of Wales, Bangor, LL57 2UW, U.K.

Abstract Waters draining a pyritic outcrop and a disused underground copper mine in north

Wales are both acidic and metal-rich. Both also contain copious amounts of microbial biomass growing as biofilms ("acid streamers"). We have studied the microbial communities in these streamers using a combination of cultivation-based and cultivation-independent molecular techniques. Isolation of acidophiles using a range of solid media indicated that the predominant microbes at one site were iron-oxidizing microbes whereas iron oxidizers and heterotrophic acidophiles were present in similar numbers in streamers from the other site. Streamer communities were also analyzed by fluorescent in situ hybridization (FISH) using a range of oligonucleotide probes of varying specificity. FISH revealed that the streamers from both mine sites were similar in overall gross structure, and were dominated (80-90% of the total bacterial counts) by β-Proteobacteria. The next most abundant group of microbes was γ-Proteobacteria, with Acidithiobacillus ferrooxidans accounting for about 25-90% of these. Archaea, low G+C Gram-positive bacteria and At. thiooxidans were not detected in streamers either by FISH and cultivation techniques, though "Ferrimicrobium"-like bacteria were detected using both approaches. These data show that the microbial communities of streamer biofilms from two different mine sites are remarkably similar, but differ substantially from those found in an abandoned pyrite mine at Iron Mountain, California.

Keywords: acid mine drainage; acid streamers; biodiversity; FISH

1. INTRODUCTION The diversity of prokaryotic microorganisms that may live in highly acidic, metal-rich

environments, such as acid mine drainage (AMD) and leachate liquors in biomining operations is known to be extensive [1]. The most obvious manifestations of microbial life in these environments are filamentous gelatinous growths, generally referred to as "acid streamers". The earliest report of acid streamers was by Lackey [2] who observed them in AMD streams in West Virginia. He originally believed that the long, colorless or light brown structures were fungal, but microscopic examination showed that they were "bacterial masses, presumably in a zoogloeal jelly". In contrast, Leathen [3] described acid streamers as consisting of fibrous masses of sheath-like structures, and that bacterial cells were absent. Temple and Koehler’s [4] description of acid streamers consisting of non-filamentous, non-orientated bacteria embedded within a tough slime, was more similar to that of Lackey. They also noted that, inside mines, acid streamers were occasionally


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encrusted with orange-brown ferric precipitates and, in AMD streams outside mines, colonized by phototrophic Euglena spp., giving them a green appearance. Dugan et al. [5] isolated a neutrophilic Bacillus sp. from acid streamers, which produced copious amounts of extracellular slime in liquid media; they concluded that this bacterium was a significant member of the acid streamer microbial community.

Johnson et al. [6] also isolated a variety of neutrophilic bacteria (several of which produced exopolysaccharides) from acid streamers growing extensively (estimated biovolume >100 m3) in an abandoned pyrite mine in North Wales. In later work [7, 8], acidophilic chemolithotrophic and heterotrophic bacteria were also isolated from the same site. These authors concluded that these acid streamers were a mixed community of iron- and sulfur-oxidizing bacteria, which were the primary producers in this pyrite mine, and that the heterotrophic bacteria lived off exudates and lysates of the primary producers. Interestingly, one of the isolates (coded "CCH7") was a ferrous iron-oxidizing filamentous isolate that formed streamer-like growths in vitro. In contrast, Wakao et al. [9] considered that it was highly unlikely that heterotrophic bacteria could form acid streamers in AMD and concluded that such growths were composed of Acidithiobacillus ferrooxidans embedded in a gelatinous matrix.

New insights into the microbial composition of acid streamer-like communities came from the work of Bond et al. [10] who studied a 1 cm-thick slime biofilm that developed on the surface of finely disseminated pyrite ore within an AMD stream at Iron Mountain, California. Phylogenetic analysis of 16S rRNA genes cloned from this microbial community indicated that the dominant sequences were of bacteria related to the iron-oxidizer Leptospirillum. Archaeal (Thermoplasmales lineage), Acidimicrobium/ "Ferrimicrobium" and δ-proteobacterial gene sequences were also detected. On the basis of these (and other) results, a variety of gene probes were designed and the slime community further analyzed using fluorescent in situ hybridization (FISH). Whilst these findings contrasted greatly with earlier results that were based chiefly on cultivation and physiological criteria, the environment within Iron Mountain was also very different (temperature 30-50ºC; pH typically 0.5-1.0) from the cooler, less acidic (pH 2-3) AMD waters in which acid streamers have been reported elsewhere.

In this paper, we describe the microbial diversity in acid streamers from AMD at two mine sites in north Wales, as revealed by using a combination of cultivation and molecular approaches.


2.1 Mine sites and water analysis Acid streamers were sampled at two mine sites in north Wales: the former Parys

copper mine and Trefriw Spa, which is situated below the abandoned Cae Coch pyrite mine [11]. Small streamer biofilm samples (ca. 1 cm3) were removed and taken to the laboratory with one hour. On site analysis of pH, redox potential, temperature and conductivity used a Whatman Water Tester (Whatman, U.K.); dissolved oxygen (DO) was measured using a D400 DO meter (Whatman, U.K.). Ferrous iron was measured in filtered (0.2 µm pore-size) samples using the ferrozine assay [12]; other metals were analyzed by atomic absorption spectrometry, and sulfate was measured turbidometrically as barium sulfate (Hydrocheck, Cambridge, U.K.). Dissolved organic carbon was measured using a Protoc Analyzer (Pollution & Process Monitoring Ltd., U.K.).


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2.2 Isolation of bacteria Streamer samples were put into bottles containing basal salts solution [13], adjusted to

pH 3.5, and vortexed for 5 min. The suspension was then centrifuged at low speed to remove undispersed streamer material, the supernatant was serially diluted, and samples plated onto a variety of solid media. The majority of these were overlay media, which facilitate the growth of a wide range of autotrophic and heterotrophic acidophiles. Full details of the media can be found elsewhere [13, 14]. R2 agar [15] was used to enumerate neutrophilic bacteria. Inoculated plates were incubated for up to 6 weeks at 20ºC under aerobic and anaerobic conditions (the latter using AnaeroGen system; Oxoid, UK).

2.3 Phylogenetic analysis of isolates The 16S rRNA genes of streamer isolates were amplified, using protocols described

elsewhere [16]. Purified PCR products were sequenced using a BeckmanCoulter dye Terminator Cycle Sequencing kit and a CEQ8000 Genetic Analysis System (Beckman Coulter, U.K.). The resulting gene sequences were compared with those available in GenBank using BLAST [17].

2.4 Total bacterial counts Bacteria in disrupted streamer suspensions (section 2.2) were harvested, washed with

phosphate buffered saline (PBS) and the re-suspended cells fixed in 3% (v/v) paraformaldehyde in PBS. Fixed cells were washed with PBS, suspensions filtered through 25 mm black polycarbonate membranes (0.2 µm pore size) and stained with 4’, 6-diamidino-2-phenylindole (DAPI; 10 ml of a 1 µg/ml solution). Membranes were placed onto glass slides and viewed using a Nikon ECLIPSE E600 microscope. A minimum number of 100 bacteria/membrane were counted.

2.5 Fluorescent in situ hybridization (FISH) The fixed streamer cells (section 2.4) were also utilized for FISH analysis, using a

modified method of that described by Bond and Banfield [18]. The RNA probes used are listed in Table 1. Where the probe used targeted Gram-positive bacteria, fixed cells were first incubated (4°C; 10 min) with lysozyme (1027.5 units of activity/ml) prior to dehydration with ethanol. Hybridization was carried out using 25 ng of fluorescein-labeled eubacterial probe (EUB338Fl) and 25 ng of a more specific, Cy3-labeled probe at the same time. Mounting medium (70% glycerol in 100 mM sodium tetraborate, pH 9.2, containing 3 mg/ml N-propyll gallate) was applied to reduce fading of the probe signal. Various concentrations of formamide (0-50%, v/v) were tested to optimize specificity and maximize signal response for each probe, using pure cultures of target and related microorganisms (Table 1). To enumerate different groups of bacteria, Cy3-labeled cells were counted relative to those stained with the EUB338Fl probe. For the eubacterial and archaeal probes, counts of stained cells were made relative to those stained by DAPI.

Table 1. List of 16S rRNA (except where noted) oligonucleotide probes used for FISH analysis and formamide concentrations required for optimum specificity. All probes, except for EUB338Fl, were labeled with the fluorochrome Cy3.

Probe Target Probe sequence (5’ 3’) Formamide (%) Reference EUB338Fl Eubacteria GCTGCCTCCCGTAGGAGT 0 – 50 [19] ALF1B α-Proteobacteria, some δ-

Proteobacteria and most spirochetes CGTTCGYTCTGAGCCAG1 20 [20]

BET42a2 23S rRNA of most β-Proteobacteria GCCTTCCCACTTCGTTT 35 [20] GAM41a2 23S rRNA of most γ-Proteobacteria GCCTTCCCACATCGTTT 35 [20] TM1G138 Thiomonas Group 1 GCAGTTATCCCCCATCAAT 40 This study TM2G138 Thiomonas Group 2 GTAGTTATCCCCCATCACA 40 This study TF539 Acidithiobacillus ferrooxidans CAGACCTAACGTACCGCC 20 [21] ATT223 Acidithiobacillus thiooxidans AGACGTAGGCTCCTCTTC 40 This study ACM732 Acidimicrobium and “Ferrimicrobium” GTACCGGCCCAGATCGCTG 35 [18] ACM995 Acidimicrobium ferrooxidans CTCTGCGGCTTTTCCCTCCATG 10 P. Norris, unpublished LGC355 Low G+C Gram-positive bacteria GGAAGATTCCCTACTGCTG 20 This study LF655 Leptospirillum groups 1, 2 and 3 CGCTTCCCTCTCCCAGCCT 35 [18] ARCH915 Archaea GTGCTCCCCCGCCAATTCCT 40 [22]

1. Y = C or T 2. Hybridization using these two probes was carried out in the presence of a 10-fold excess of the other unlabeled oligonucleotide


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3. RESULTS

3.1 Mine water analysis Physico-chemical data from the two mine sites from where streamer biofilms were

sampled are shown in Table 2. At both sites, conditions were reducing, with both dissolved oxygen and redox potentials being relatively low, and all of the soluble iron present was in the ferrous form.

Table 2. Physicochemical data from the two AMD sites from which streamer biofilms were obtained. N.D. = not determined

Parys mine Trefriw spa pH 2.70 3.05 T (ºC) 8.8 13.3 Eh (mV) +420 +374 Ec (µS/cm) 2994 1635 DO2 (mg/L) 0.55 1.66 DOC (mg/L) 6.0 14.9 Fe2+ (mg/L) 473 217 SO4

2- (mg/L) 1552 1093 Al (mg/L) 115 N.D. Cu (mg/L) 50 N.D. Zn (mg/L) 60 N.D.

3.2 Plate and DAPI Counts Colonies that grew on overlay media were differentiated on the basis of whether or

not they became encrusted with ferric iron (i.e. iron-oxidizing autotrophs and heterotrophs), and on other morphological characteristics (color, size, etc.). Iron-oxidizing isolates were further differentiated into moderate and extreme acidophiles, depending on whether they grew on ferrous iron/thiosulfate (pH ~4) or ferrous iron (pH ~2.5) overlay plates [14]. Enumeration of bacteria in dispersed biofilms determined by plate counts (on aerobically-incubated plates) is shown in Table 3. Plating efficiencies, determined relative to total (DAPI) cell counts were 1.8% for the Parys streamers and 0.96% of the Trefriw streamers. Numbers of isolates obtained on solid media incubated under anaerobic conditions were considerably, 102- to 103-fold, fewer than obtained with plates incubated aerobically.

3.3 Phylogenetic analysis of streamer isolates Identification of representative streamer isolates, differentiated on the basis of colony

morphologies and identified from analysis of amplified 16S rRNA genes, is shown in Table 4. All of the iron-oxidizing autotrophic isolates examined in streamers at both sites were At. ferrooxidans rather than Leptospirillum spp. In addition, "Ferrimicrobium"-like isolates (Actinobacteria) were isolated from both streamers, as were Acidiphilium and Acidocella spp. (α-Proteobacteria). Parys streamer also included an isolate remotely related to Acidobacterium capsulatum, a Frateuria-like isolate (γ-Proteobacteria), and a Thiomonas-like isolate (β-Proteobacteria). Interestingly, the latter was initially classified as a heterotroph (being isolated on yeast extract solid medium) though it was later confirmed that it was capable of oxidizing thiosulfate [23].


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Table 3. Plate counts (aerobic) and total microbes in the dispersed streamer samples from the two sites (CFU per g wet weight dispersed streamer)

Parys streamers Trefriw streamers Iron-oxidizing bacteria - extreme acidophiles 9.36 x 106 1.16 x 107 - moderate acidophiles 2.42 x 103 <102 Heterotrophic acidophiles 4.40 x 105 4.80 x 106 Neutrophiles <102 <102 Total plate counts 9.80 x 106 1.64 x 107 Total counts (DAPI) 5.44 x108 1.7 x109

Table 4. Identification of acid streamer isolates based on analyses of amplified 16S rRNA genes

(i) Parys streamers

Isolate Physiological class Nearest relative (% identity) KP1 Iron-oxidizing heterotroph Uncultured eubacterium TRA2-10 (99.9)

“Ferrimicrobium acidophilum” (93.0) PK11 Iron-oxidizing autotroph At. ferrooxidans NO-37 (99.9) PK51 Heterotroph γ-proteobacterium WJ2 (100) PK40 Heterotroph Acidiphilium NO-17 (98.0) PK46 Heterotroph Acidiphilium organovorum (98.5) PK48 Heterotroph Acidiphilium acidophilum (97.5) M21 Heterotroph Acidocella NO-12 (99.4) PK35 Heterotroph Acidobacterium WJ7 (99.5) KP3 Heterotroph Uncultured eubacterium TRB82 (96.7)

Acidobacterium capsulatum (91.8) PK44 Heterotroph Thiomonas CO2 (97.4)

ii) Trefriw streamers

Isolate Physiological class Nearest relative (% identity) CS11 Iron-oxidizing heterotroph Uncultured eubacterium TRA2-10 (99.7)

“Ferrimicrobium acidophilum” (93.0) CCW10 Iron-oxidizing autotroph At. ferrooxidans NO-37 (99.5) CCW68 Iron-oxidizing autotroph At. ferrooxidans ATCC 19859 (99.0) CCW30 Heterotroph Acidocella NO-12 (99.1) CCP3 Heterotroph Acidiphilium NO-17 (99.5)

3.4 FISH analysis of dispersed streamers Analysis of microorganisms in the disrupted streamers by FISH produced data that

both confirmed and contrasted with results from cultivation techniques. The prokaryotes in the streamers were shown to be exclusively (>99.9%) bacteria; no cells were detected using the archaeal probe. Also in line with plating data, probes targeting Leptospirillum spp., At. thiooxidans and low G+C Gram-positive bacteria also failed to detect any cells.


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The Gram-positive bacteria that were detected by FISH were exclusively Actinobacteria detected by the Acidimicrobium/”Ferrimicrobium” probe but not by the Am. ferrooxidans-specific probe.

The vast majority of cells in both streamer biofilms were Proteobacteria. Whilst α-, β- and γ-Proteobacteria were all detected by FISH, the dominant group in both Parys and Trefriw streamers were β-Proteobacteria (Figure 1). The only β-proteobacterium isolated was a Thiomonas species from the Parys streamer. By FISH analysis with the corresponding probe (TM1G138), however, this microbe was detected at less than 1 cell per thousand in the Parys streamer and was not detected at all in the Trefriw streamer. A second probe (TM2G138), targeting another group of Thiomonas species [23], also failed to detect cells in either streamer.

Figure 1. Counts of microbes in the streamer samples targeted by a range of Cy3-labeled oligonucleotide probes, given as percentages of the fluorescein-labeled Eubacterial probe (EUB338Fl)

4. DISCUSSION The application of molecular techniques is increasingly becoming important in

microbial ecology [24, 25]. The power of these techniques lies in the ability to identify and enumerate microorganisms without the reliance on cultivation. Cultivation-based studies often lead to incorrect conclusions, such as the importance of Acidithiobacillus ferrooxidans in commercial bioleaching operations, which have been shown by a molecular approach to be dominated by Leptospirillum ferrooxidans and relatives [26].

In this study, we have taken the rarely used approach of combining cultivation with molecular studies to characterize the microbial populations of streamer biofilms that dominate mine water draining two metal mines in north Wales. Earlier cultivation-based studies were carried out on both sites, and the identities of the microbes present were determined solely on physiological characteristics. Since that time, improved solid media for the cultivation of acidophiles have been developed, and were employed in the present study along with 16S rRNA-based identification of the isolates. As would be expected from earlier studies, the dominant cultivatable isolates from both sites were iron-oxidizing microbes and the majority of these were found, by 16S rRNA gene sequence analysis, to be Acidithiobacillus ferrooxidans.


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The heterotrophs isolated from both streamers proved to be more diverse than previously thought, especially in the Parys streamer. The heterotrophs from both streamers included isolates from the well-known genera of acidophiles Acidiphilium and Acidocella, and amongst these were isolates that may represent new species of these genera. Some previously unrecognized microbes also inhabit the streamer from Parys Mountain. One isolate, PK51, is related to neutrophilic microbes of the genus Frateuria, PK51, and two other isolates are related to Acidobacterium capsulatum. The homology of these isolates to Ab. capsulatum and to each other is sufficiently low that they can be considered to belong to different genera. It is interesting to note that few microbes of the phylogenetic group that includes Ab. capsulatum have been cultivated, even though molecular studies show that they are the dominant microbes in many habitats [e.g. 27]. The latter two groups of microbes have also been isolated from a wetland constructed to treat acid mine drainage [14].

As with the cultivation-based study, the populations of both of the streamers appeared to be similar when analyzed with probes targeting ribosomal RNA. There was no indication of archaea present in either streamer community, as indicated by FISH and also by the absence of an amplification product following PCR with archaeal 16S rRNA gene-specific primers (data not shown). In contrast to the cultivation study, both of the streamers were dominated by β-Proteobacteria and not γ-Proteobacteria, the probe for which detects At. ferrooxidans. The only isolate obtained from either of the streamers, PK44, that belongs to the β-Proteobacteria is related to the genus Thiomonas. Using probes to target the 16S rRNA gene of bacteria in this genus, however, it was shown that they are not the dominant microbes in either streamer, in keeping with the cultivation-based studies. Further molecular characterization of these streamers, including the generation of 16S rRNA gene libraries, is currently in progress and will reveal the identity of the dominant microbes in the streamer biofilms. Identification of the unknown microbe, or microbes, that dominate these streamers will be important to understanding the biogeochemistry of these microbial growths that are very prominent in AMD.

Even though the streamers studied here were quite similar to each other, they differed substantially from similar microbial growths found in AMD at Iron Mountain, California. Analysis of 16S rRNA genes cloned from Iron Mountain streamer material showed that those streamers were dominated by Leptospirillum-like microbes [28], which have only been found there to date, and by archaea. Other microbes found were related to “Ferrimicrobium acidophilium” and Acidimicrobium ferrooxidans. The Parys and Trefriw isolates related to “Ferrimicrobium” were highly related (>99% 16S rRNA gene homology) to the Iron Mountain clone TRA2-10. FISH analysis [10] confirmed that the major microbes in the Iron Mountain streamers were the Leptospirillum-like microbes and archaea, the latter namely Ferroplasma. Neither the Acidimicrobium/”Ferrimicrobium” type bacteria, nor At. ferrooxidans, were detected by FISH, in contrast to the Parys and Trefriw streamers.

While it is difficult to say what determines the differences in the microbial constitution of these various streamers, certainly the physico-chemical parameters play a central role, with temperature (~40°C at Iron Mountain vs ~10°C) and pH (<2 at Iron Mountain vs ~3) having significant effects.

ACKNOWLEDGEMENTS Kris Coupland is grateful to the NERC (U.K.; ref. NER/S/C/2001/06450) and Rio

Tinto Technologies for provision of a research studentship. K.B.H. and D.B.J. thank the BBSRC and DTI (U.K.; ref. 5/BRM18412) for partial support of this work.


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REFERENCES 1. K. B. Hallberg and D. B. Johnson, Adv. Appl. Microbiol. 49 (2001) 37. 2. J. B. Lackey, Public Health Rep. 53 (1938) 1499. 3. W. W. Leathen (1952) Special Summary Report of Microbiological Studies of

Bituminous Coal MIne Drainage, Mellon Institute of Industrial Research, University of Pittsburg, Pittsburg, Pa.

4. K. L. Temple and W. A. Kohler (1954) In Research Bulletin No. 25, Engineering Exp. Station, University of West Virginia, Morgantown.

5. P. R. Dugan, C. B. Macmillan and R. M. Pfister, J. Bacteriol. 101 (1970) 982. 6. D. B. Johnson, W. I. Kelso and D. A. Jenkins, Environ. Pollut. 18 (1979) 107. 7. D. B. Johnson, M. A. Ghauri and M. F. Said, Appl. Environ. Microbiol. 58 (1992)

1423. 8. S. McGinness and D. B. Johnson, Sci. Total Environ. 132 (1993) 27. 9. N. Wakao, H. Tachibana, Y. Tanaka, Y. Sakurai and H. Shiota, J. Gen. Appl.

Microbiol. 31 (1985) 17. 10. P. L. Bond, G. K. Druschel and J. F. Banfield, Appl. Environ. Microbiol. 66 (2000)

4962. 11. D. A. Jenkins and D. B. Johnson, J. Russell Soc. 5 (1993) 40. 12. D. R. Lovley and E. J. P. Phillips, Appl. Environ. Microbiol. 53 (1987) 1536. 13. D. B. Johnson, J. Microbiol. Meth. 23 (1995) 205. 14. K. B. Hallberg and D. B. Johnson, Hydrometallurgy (2003) in press. 15. D. J. Reasoner and E. E. Geldreich, Appl. Environ. Microbiol. 49 (1985) 1. 16. N. Okibe, M. Gericke, K. B. Hallberg and D. B. Johnson, Appl. Environ. Microbiol. 69

(2003) 1936. 17. S. F. Altschul, T. L. Madden, A. A. Schaffer, J. Zhang, Z. Zhang, W. Miller and D. J.

Lipman, Nucleic Acids Res. 25 (1997) 3389. 18. P. L. Bond and J. F. Banfield, Microbial Ecol. 41 (2001) 149. 19. R. I. Amann, L. Krumholz and D. A. Stahl, J. Bacteriol. 172 (1990) 762. 20. W. Manz, R. Amann, W. Ludwig, M. Wagner and K. H. Schleifer, Syst. Appl.

Microbiol. 15 (1992) 593. 21. M. O. Schrenk, K. J. Edwards, R. M. Goodman, R. J. Hamers and J. F. Banfield,

Science 279 (1998) 1519. 22. D. A. Stahl and R. Amann, Nucleic Acid Techniques in Bacterial Systematics, E.

Stackebrandt and M. Goodfellow (eds.), Wiley, New York, (1991) 205. 23. K. Coupland, F. Battaglia-Brunet, K. B. Hallberg, M. C. Dictor, F. Garrido and D. B.

Johnson, This Volume 24. I. Amann, W. Ludwig and K. H. Schleifer, Microbiol. Rev. 59 (1995) 143. 25. I. M. Head, J. R. Saunders and R. W. Pickup, Microbial Ecol. 35 (1998) 1. 26. N. J. Coram and D. E. Rawlings, Appl. Environ. Microbiol. 68 (2002) 838. 27. P. Hugenholtz, B. M. Goebel and N. R. Pace, J. Bacteriol. 180 (1998) 4765. 28. P. L. Bond, S. P. Smriga and J. F. Banfield, Appl. Environ. Microbiol. 66 (2000) 3842.



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Computational fluid dynamics simulation of immobilized Acidothiobacillus ferrooxidans

B. Metodieva, L. Lilovab and D. Karamanevb

a Blue EdgeValley, 103 Gerald Cr., London, Ontario, Canada N5Z 5A4 b Department of Chemical and Biochemical Engineering, The University of Western

Ontario, London, Ontario, Canada N6A 5B9

Abstract This paper presents the results of the simulation, using computational fluid dynamics,

of the flow and ferric iron concentrations profiles around a single cell of A. ferrooxidans, immobilized on the surface of a sulfide crystal. It has been shown that there are significant concentration gradients of ferric iron concentration between the surface of the crystal and the bulk of liquid. The gradient can reach several hundred mg/L. These results can explain the difference between the chemical (sterile) and biological oxidation rates of metal sulfides in liquid media containing iron ions.

Keywords: computational fluid dynamics, Acidithiobacillus ferrooxidans, bioleaching

1. INTRODUCTION The biological oxidation of metal sulfides by Acidothiobacillus ferrooxidans is of

great importance for the hydrometallurgical production of base, heavy and noble metals such as copper, zinc, cobalt, uranium and gold. Because of the practical importance of these processes, the mechanism of sulfide biooxidation has been a subject of intensive studies in the recent decades. Two main mechanisms of the oxidation kinetics have been proposed [1]: • the direct oxidation is based on the oxidation of sulfide minerals by an immobilized

on their surface microbial cell in the absence of intermediate oxidants such as ferric iron. It is based on the following reaction (in the case of pyrite oxidation): 4FeS2 + 15O2 + 2H2O = 2Fe2(SO4)3 + 2H2SO4 (1)

• the indirect oxidation is based on the ability of A. ferrooxidans and other iron-oxidizing microorganisms to oxidize ferrous ions in aqueous solution: 2FeSO4 + H2SO4 + ½O2 = Fe2(SO4)3 + H2O (2) The ferric ions, produced as a result of this reaction, then can chemically oxidize

metal sulfides, such as pyrite: 7Fe2(SO4)3 + FeS2 + 8H2O = 15FeSO4 + 8H2SO4 (3)

The ferrous ions produced by reaction (3) are then oxidized by A. ferrooxidans to ferric ions according to reaction (2). The ferric sulfate, produced by reaction (2) is further


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reused for the chemical oxidation of pyrite (reaction 3), which closes the cycle of iron ions oxidation-reduction.

After many years of research, it seems that a consensus has been reached that the indirect mechanism is if not the only [2-4], at least the predominant [5] mechanism in sulfide biooxidation by A. ferrooxidans. However, the rate of purely chemical (sterile) oxidation of pyrite by aqueous solutions of ferric sulfate is by an order of magnitude lower than that in the presence of iron oxidizing microorganisms [6]. Different explanations have been proposed to explain this discrepancy.

Boon and Heijnen [6] explained the higher rate of the biological sulfide oxidation by the fact that the ferrous ions, produced as a result of the sulfide oxidation (3) are quickly reoxidized by microorganisms, which keeps the redox potential high; in the sterile oxidation of sulfides by ferric iron, the redox potential falls quickly because of the decrease in the ratio of ferric to ferrous ions concentration. Alternatively, Gehrke et al. [7] explained the acceleration of the bacterial process over the sterile one by the quick reoxidation of iron, which is complexed with microbial extracellular polymeric substances.

The purely hydrodynamic retention of ferric ions near the point of contact of microorganisms with the sulfide surface has not been considered yet. In this work, we are trying to explain the higher rate of bacterial sulfide oxidation, compared to the sterile one, by the hydrodynamic retention of ferric ions. Computational fluid dynamics simulation is used to determine the ferric ions concentration profiles in the vicinity of a microorganism, attached physically to the surface of a sulfide crystal.

2. NUMERICAL MODEL The concentration and velocity profiles around a single microbial cell, immobilized

on the flat surface of a sulfide crystal were modeled numerically using the FLUENT computational fluid dynamics software.

2.1 Assumptions The following assumptions were made in order to develop the mathematical model:

Geometrically, the A. ferrooxidans cell was approximated as an ellipsoid of rotation with a size along the axis of symmetry equal to 2 micrometers and a small axis, perpendicular to it, of 1 micron (Fig. 1). While the size and shape slightly differ in different strains, the one used here is a typical shape and size. The surface area of the microbial cell was 5.34x10-12 m2.

The cell is a rigid solid particle, and therefore: − It does not deform as a result of the acting hydrodynamic forces. − The cell does not deform as a result of immobilization on the solid surface,

especially at the point of contact. The cell is immobilized on a flat surface (Fig. 1). The immobilization support is chemically inert. No extacellular polymeric substances (EPS) were assumed to exist between the cell

and the sulfide surface. The available literature data shows that in some cases A. ferrooxidans forms EPS when immobilized on a sulfide surface, while in other they are absent.

The ferric ions concentration in the bulk of the liquid is equal to zero.


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The ferric ions are produced evenly over the entire surface of the cell and are completely released into the liquid. The literature data show that the process of iron oxidation takes place outside of the cell membrane and the iron ions do not permeate the membrane [8]. The iron oxidation rate was assumed to be equal to 5.2x10-8 (kg Fe3+)/(s.(m2 of cell surface)).

The approaching liquid flows in either laminar or turbulent mode, parallel to the solid surface (Fig. 2), with two bulk velocities: 0.1 and 1.0 m/s. The case for 1.0 m/s was calculated for two different cases – laminar and turbulent liquid flow. For the turbulent flow the turbulence intensity was 5% and the turbulent scale was chosen at 0.01 mm.

The real geometry of the microbial cell has been scaled up by a factor of 100 in order to avoid round-off error of computing too small numbers (note that FLUENT computes only in SI units). This required scaling down the liquid velocity by the same factor in order to keep a constant Reynolds number. The computational mesh consisted of 0.83 mill tetrahedral cells. The surface mesh of the bacteria is shown in Fig. 3.

Advance turbulent model, namely the FLUENT Shear-Stress Transport (SST) k-Ω model was applied to resolve the low-Reynolds number turbulence around the bacteria.

Figure 1. Geometry of the immobilized microbial cell


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Figure 2. Scheme of the modelling space

Figure 3. Bacteria surface mesh

2.2 Numerical results and discussion The velocity and concentration profiles in the plane x-z (side view) were obtained

(Figs. 4-6) using the model, described above.


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The ferric iron concentration profiles in liquid, at the plane of contact with the sulfide crystal surface, are shown in Fig. 7. Actually, Fig. 7 shows the top view (x-y plane) of Figs. 4, 5 and 6, assuming that the microbial cell is "transparent".

Figure 4. Velocity and ferric ions concentration profiles around the immobilized cell. Liquid velocity: 0.1 m/s (laminar flow)


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Figure 5. Velocity and ferric ions concentration profiles around the immobilized cell. Liquid velocity: 1.0 m/s (laminar flow)


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Figure 6. Velocity and ferric ions concentration profiles around the immobilized cell. Liquid velocity: 1.0 m/s (turbulent flow)


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Figure 7. Top view of the ferric ions concentrations in liquid, contacting the sulfide crystal surface

The concentration profiles in Fig. 7 show that both the free stream velocity and flow

structure (laminar vs. turbulent) of liquid affect the concentration distribution only slightly. This is due to the fact that the size of the microorganism is in the order of micrometers, thus the entire microorganism is within the laminar viscous layer of the liquid, regardless of the type of the flow – laminar or turbulent. Within the liquid viscous sublayer of 1 micrometer, the liquid velocity is very close to zero, which can be seen also from the velocity plots in Figure 4. Therefore, the upstream velocity magnitude, as well as


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the upstream turbulence does have a significant effect on the concentration distribution (Fig. 7). The tail of a relatively larger concentration downstream the bacteria is wider at smaller upstream velocity. At larger velocity the larger amount of the ferric ions are advected by the flow and the thinner tail of higher concentration is formed downstream of the microbial cell.

3. CONCLUSIONS The results of the simulation show that there is significant spatial variation in ferric

ions concentration in the close surrounding of the microbial cell (Fig. 7). While the concentration at the vicinity of the contact between the microbial cell and the crystal surface is well over 100 mg/L, it drops to zero at a distance of 1 micrometer from the cell surface (except for the tail behind the cell). Therefore, the concentration of ferric ions in the bulk of liquid, far away from the cell, is different - much lower - than that on the sulfide surface, surrounding the immobilized cell. The redox potential of the liquid, given by the Nernst equation: E = Eo + (RT/F)ln(Fe3+/Fe2+) (4) is a strong function of the ferric-to-ferrous ions concentration ratio. Therefore, since the concentration of ferric ions in the immediate cell surroundings is higher than the bulk concentration, the redox potential at the surface of the sulfide crystal will be higher that that in the bulk of liquid. Since the ferric iron concentration at the cell-sulfide interface is higher than the bulk concentration by approx. 150 mg/L, the redox potential at the cell-sulfide interface will be significantly higher than the bulk potential mainly in the cases when the bulk ferric iron concentration is below approx. 1 g/L.

ACKNOWLEDGEMENTS This work was supported financially by the Natural Sciences and Engineering

Research Council of Canada and the Premier’s Research Excellence Award (Ontario).

REFERENCES 1. Α.E. Torma, Adv. Biochem. Eng., 6 (1977) 1. 2. W. Sand, T. Gehrke, P.-G. Jozsa and A. Schippers, Hydrometallurgy, 59 (2001)159. 3. H. Tributsch, Hydrometallurgy, 59 (2001) 177. 4. G.S. Hansford and T. Vargas, Proc. Metall., 9A (1999) 13. 5. K. Lilova and D. Karamanev, (paper to appear in this volume) 6. M. Boon and J. J. Heijnen, Hydrometallurgy, 48 (1998) 27. 7. T. Gehrke, R. Hallmann, K. Kinzler and W. Sand, Water Sci. Tech., 43 (2001) 159. 8. W. J. Ingledew, Biochim. Biophys. Acta, 683 (1982) 89.



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Contribution to the quantification of the Acidithiobacillus ferrooxidans biomass concentration from the oxygen uptake

rate

D.S. Savić, V.B. Veljković, M.L. Lazić

Faculty of Technology, 16000 Leskovac, Bulevar oslobodjenja 124

Abstract A simple, rapid and reproducible method for quantifying the A. ferrooxidans biomass

concentration from measuring the oxygen uptake rate is described. The specific oxygen uptake rate, found to be 0.020 ± 0.001 mg O2/mgd.w./min, for both concentrated cell suspensions and during ferrous iron biooxidation, is used in the estimation of the viable biomass concentration. The ferrous iron concentration in samples of the culture medium, where the oxygen uptake is recorded, should be higher than 0.1 mg/L to avoid substrate limitation.

Keywords: Acidithiobacillus ferrrooxidans, qantification of viable biomass concentration

1. INTRODUCTION The role of the chemolithotroph bacterium Acidithiobacillus ferrrooxidans in the

leaching of metals from sulfide ores has been well recognized. The bacterium obtains energy for growth by the oxidation of ferrous to ferric iron using atmospheric oxygen, a desirable reaction from the hydrometallurgical point of view. As in other cases, when oxygen is closely related to the energetics of microbial growth under aerobic conditions [1], the oxygen consumption of A. ferrooxidans could be correlated to its growth rate and, therefore, used for indirect kinetic studies. A few earlier studies on oxidations by A. ferrooxidans, such as the ferrous iron oxidation rate by cell envelopes [2], the competitive inhibition of ferrous iron oxidation by cells and ferric iron [3, 4], and the relative contributions of adhered and free suspended cells to pyrite biooxidation [5], are based on the measurement of oxygen uptake. The oxygen uptake rate is also suggested for evaluating bacterial activity for metal leaching [6].

A high degree of synchronization among processes of oxygen uptake, biomass growth and ferrous iron utilization during ferrous iron oxidation by A. ferrooxidans in different bioreactors, independently of pH (1.8 or 2.0), has been observed [7]. Thus, the dissolved oxygen level could be used not only for studying the kinetics of the biooxidation process but also for determining the viable biomass concentration. Such a method, based on the measurement of the oxygen uptake rate, would be a simpler technique for quantifying the A. ferrooxidans growth during the biooxidation of ferrous iron than the standard ones, which have disadvantages in requiring considerable time, materials and manual labor (serial end-point dilution, enumeration of colonies) or refers to dead and living cells


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(direct microscope counting, the measurement of the total dry weight and protein content). The viable biomass concentration and the growth rate have been successfully determined with sufficient accuracy by measuring the oxygen uptake rate during aerobic fermentations, for instance by Saccharomyces cerevisiae, Streptomyces sp. and Thermoactinomyces sp. [8] as well as Clostridium boldinii [9].

In the present work, the growth and oxygen uptake of A. ferrooxidans cells growing on a ferrous iron medium were studied under different conditions with respect to the ferrous iron concentration, the biomass concentration and the ferric/ferrous ratios. The main aim was to reflect the relationship between microbial growth and oxygen uptake and to establish a quantitative base for determining the viable biomass concentration by measuring the oxygen uptake rate.


2.1 Microorganism and cultivation conditions A strain of A. ferrooxidans B5, isolated from the Bor Copper Mine (Serbia), was

grown on the following mineral salt medium (in g/L): (NH4)2SO4 3; KH2PO4 0.5; MgSO4

.7H2O 0.5; and KCl 0.1, the initial ferrous iron concentration being 9 g/L. The culture was grown in an Erlenmeyer flask (2 L) containing 1 L of the medium at 28°C on a rotary shaker at 200 rpm.

2.2 Biomass The biomass was determined by measuring the dry weight - mgd.w (80°C, 24 h) of a

membrane filter (0.45 µm pore diameter, Millipore, NVC 45218) after the sample (500 mL) was filtered under pressure (2 bar).

2.3 Preparation of a concentrated cell suspension The biomass was harvested at the end of the exponential growth phase; it was

therefore assumed that the biomass contained a negligible number of dead cells. A volume of the growing culture (500 mL; 25 mgd.w./L) was centrifuged at low speed (1000 min-1, 10 min) to remove the insoluble ferric iron, filtered through a membrane filter (0.45 µm pore diameter, Millipore, NVC 45218), washed twice with 0.05 M H2SO4 solution (100 mL) and once with distilled water (100 mL), and concentrated by resuspending the biomass in the iron-free salt medium (25 mL). According to the ratio of the oxygen uptake rate of the original culture and that of the concentrated suspension, about tenfold concentration was achieved (238 mgd.w./L).

2.4 Oxygen uptake rate The oxygen uptake rate was calculated from the slope of the linear part of the

relationship between the dissolved oxygen level and time. The dissolved oxygen level was measured by a Clarke-type oxygen probe (HANNA HI8043), and the oxygen solubility (6.7 mg O2/L) in the mineral salt medium was estimated [10]. The reaction mixture (5 mL) was prepared by mixing a certain volume of the concentrated cell suspension with different volumes of ferrous iron and ferric iron solutions in the mineral salt medium to achieve different ratios of ferric/ferrous iron. The oxygen uptake rate of the cell-free mineral medium was found to be negligible (0.007 mg O2/L/min).


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2.5 Analytical methods The ratio of ferrous and ferric ions was calculated from the redox potential measured

using a redox probe (HANNA HI 3932). The pH was monitored with a pH meter (HANNA HI 9025).

3. RESULTS AND DISCUSSION A summary of the effect of different biomass concentrations and ferric/ferrous ratios

on the specific oxygen uptake rate of the concentrated bacterial cells is given in Tables 1 and 2. When the biomass concentration is held constant at a level of 8.0 mgd.w./L and the ferric/ferrous ratio is varied in the range from 1/10 to 1/0.1, which causes a change in the pH of the reaction mixture from 1.9 to 1.4, the specific oxygen uptake rate does not change. A mean value of 0.0185 ± 0.001 mg O2/mgd.w./min (Table 1) was obtained. Also, if the ferric/ferrous ratio of 1/10 is held constant and the biomass concentration in the reaction mixture is varied from 4.3 to 47.7 mgd.w./L, the specific oxygen uptake rate is also constant at a level of 0.020 ± 0.003 mg O2/mgd.w./min (Table 2). At the significance level of 0.05, as evaluated by the t-test (two populations), the two means are not significantly different. As both the biomass concentration and the ferric/ferrous iron ratio do not have a significant effect, the mean value of the specific oxygen uptake rate for all the experimental conditions applied was calculated: 0,020 ± 0.002 mg O2/mgd.w./min.

Table 1. Effect of the ferric/ferrous iron ratio on the specific oxygen uptake rate (biomass concentration: 8.0 mgd.w./dm3)

Fe(III)/Fe(II) qO2, mg O2/mgd.w./min 1/10 0.019 1/7 0.019 1/5 0.018 1/2 0.018

1/0.3 0.019 1/0.1 0.018

Mean value Standard deviation

0.0185 ±0.001

Table 2. Effect of the biomass concentration on the specific oxygen uptake rate (ferric/ferrous iron ratio: 1/10)

X, mgd.w./dm3 qO2, mg O2/mgd.w./min 4.3 0.017 8.0 0.019

13.6 0.025 17.9 0.021 47.6 0.022

Mean value Standard deviation

0.020 ± 0.003

The specific oxygen uptake rate found for the concentrated cells was practically the same as that determined in our studies on ferrous iron biooxidation in a bubble column without and with pH regulation at 1.8 and 2.0 and in a stirred tank with pH regulation at 2.0, as may be seen in Table 3. The mean value of the specific oxygen uptake rate was calculated from all our data: 0.020 ± 0.001 mg O2/mgd.w./min. This value agrees quite well


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with those measured for A. ferrooxidans grown in a batch reactor using the same type of oxygen probe [11] or the manometric technique [12] while other published values are twice [13, 14] or even tenfold [2] higher. The disagreement is probably due to different strains, measuring methods or cultural conditions applied in the different studies.

Table 3. Specific oxygen uptake rate of A. ferrooxidans under different culture conditions*

Culture condition qO2, mgO2/mgd.w./min Reference

Concentrated cells, pH 1.4-2.4 0.020 ± 0.002 This work Bubble column, 9 gFe/L, 28°C without pH regulation (pH 2.4-2.7)

0.021 ± 0.001 Unpublished data

Bubble column, 9 gFe/L, 28°C with pH regulation: 1.8 2.0

0.018 ± 0.001 0.020 ± 0.001

[5]

Stirred tank, 9 gFe/L, 28°C with pH regulation: 2.0

0.021 ± 0.002

[5]

Fermentor, 9 gFe/L, 28°C, pH 3.2 0.205 (230 mmol O2/mgpr./h)c

[2]

Batch reactor, 10 gFe/dm3, 30°C, pH 1.6 0.020 (630 nmol O2/mgd.w./min)

[11]

Sample from uranium b

leaching + 0.20 gFe, 30°C 0.025

(700 mL O2/mgpr./h)c [12]

Not defined a 0.045 (20840 mL O2/mgN/h)c

[13]

Shaken flasks, 0.56 gFe/dm3, 30°C, pH 1.6 0.042 (785 nmol O2/mgpr./min)c

[14]

* In all cases, a Clarke oxygen electrode was used with the following exceptions: (a) not defined and (b) the manometric method

c Assuming that the dry biomass of A. ferrooxidans contains about 60% of proteins [16, 17] and 12.4% of nitrogen [15].

The oxygen uptake rate of A. ferrooxidans, evaluated in experiments with the concentrated cell suspensions and during ferrous iron biooxidation is shown in Figure 1 as a function of the biomass concentration. All data fit a straight line, indicating that the oxygen uptake rate of the bacterial population, OUR (in mg O2/L/min), during ferrous iron biooxidation can be correlated to the biomass concentration, X (mgd.w./L), by the well-known equation:

OUR q XO=2

where qO2is the specific oxygen uptake rate of the strain A. ferrooxidans B5 (0.020 mg

O2/mgd.w./min). This correlation allows the determination of the viable biomass concentration of A. ferrooxidans from the oxygen uptake rate measured during the biooxidation process. The only limitation for applying the above equation is the ferrous iron concentration in the culture medium where the oxygen uptake rate is measured. From the stoichiometry the ferrous iron oxidation and cell synthesis by A. ferrooxidans [15], and taking the maximum oxygen rate and the usual duration time of the oxygen uptake rate measurement as 0.70 mg O2/L/min [7] and 20 minutes, respectively, the critical ferrous iron concentration in the medium, which limits the bacterial growth, was estimated to be about 0.1 g/L. This value was verified by measuring the oxygen uptake rates in samples


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taken from the stationary growth phase, where the ferrous iron concentration was below the critical value, and in the same samples after the addition of a ferrous iron solution to adjust the ferrous iron concentration above the critical level. A summary of the oxygen uptake rate of samples taken from the stationary growth phase with and without the addition of ferrous iron is presented in Table 4. The oxygen uptake rate in the latter is lower than that in the former, indicating the limitation of oxygen uptake by ferrous iron.

Figure 1. Oxygen uptake rate (OUR) as a function of dry biomass concentration (experiments with concentrated biomass suspension: ♦; the biooxidation of ferrous iron in a bubble column without pH regulation - [unpublished data] and with pH regulation at 1.8 -

and 2.0 - ; and the biooxidation of ferrous iron in a stirred tank with pH regulation at 2.0: [7])

Table 4. The oxygen uptake rate of A. ferrooxidans in samples from the stationary growth phase in a bubble column* before and after the addition of ferrous iron (up to 9 g/L)

Oxygen uptake rate, mg O2/L./min pH

Ferrous iron concentration in the culture medium, g/L

Before the addition of ferrous iron

After the addition of ferrous iron

2.4-2.7 0.08 0.08 0.41 1.8 0.08 0.07 0.57 2.0 0.03 0.08 0.50

0.05 0.06 0.56 0.05 0.055 0.48 *Data from the experiments described in [7], but not shown there.


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4. CONCLUSIONS Quantification of the A. ferrooxidans biomass concentration using usual methods,

such as serial end-point dilution and enumeration of colonies, is uncertain and time-consuming. The measurement of dry biomass, although simpler and reproducible, can only be used for estimating the biomass concentration in large bioreactors since a volume of 500-750 mL [7] is needed as a sample.

This work describes a simple method for quantifying the A. ferrooxidans biomass concentration indirectly by measuring the oxygen uptake rate by a cell population. The advantages of this method are that the test is rapid, small samples are needed, and the results are reproducible. The method could also be appropriate for enumerating aerobic microorganisms in bioleaching experiments, since many cells are attached to the ore and could not be detected by standard methods. The following assumptions were made when the indirect method was used. Firstly, all the oxygen consumed in the bioprocess was only used for ferrous iron oxidation, the oxygen, which became part of the biomass was derived from combined oxygen found in CO2, and the oxygen required for maintenance was negligible. Secondly, autooxidation of the ferrous iron did not occur, which was verified expermentally.

REFERENCES 1. M.L. Shuler and F. Kargi, Bioprocess Engineering, Prentice Hall, Englewood Cliff,

New Jersey, 1992. 2. C. Bodo and D.G. Lundgren, Can. J. Microbiol., 20 (1974) 1647 3. H.M. Lizama and I. Suzuki, Appl. Environ. Microbiol., 55 (1989) 2588 4. I. Suzuki, H.M. Lizama, P.D. Tackaberry, Appl. Environ. Microbiol., 55 (1989) 1117 5. D. Savić, V. Veljković, M. Lazić and M. Vrvić, in: R. Amils and A. Ballester (eds.),

Biohidrometallurgy and the Environment toward the Mining of the 21st Century, Elsevier, Amsterdam, 1999, Part A, pp 625-630.

6. G.I. Karavaiko, G. Rossi, A.D. Agate, S.N. Groudev and Z.A. Avakyan, (eds.) Biogeotechnology of Metals, CIP GKNT, Moscow, 1988

7. V. Veljković, D. Savić, M. Lazić and M. Vrvić, in: R. Amilis and A. Ballester (eds.), Biohidrometallurgy and the Environment toward the Mining of the 21st Century, Elsevier, Amsterdam, 1999, Part A, pp 617-623.

8. D.W. Zabriskie and A.E. Humphrey, AIChE J., 24 (1978) 138 9. M. Reuß, R.P. Jefferis and J. Lehman, GBR Monogr., 3 (1976) 107 10. A. Schumpe, I. Adler and W. Deckwer, Biotechnol. Bioeng., 20 (1978) 145 11. J.T. Pronk, W.M. Meijer, W. Hazeu, J.P. vanDungen, P. Bos and G. Kuenen, Appl.

Environ. Microbiol., 57 (1991), 2057 12. C.L. Brierley, in: L.E. Murr, A.E. Torma and J.A. Brierley (eds.) Metallurgical

Applications of Bacterial Leaching and Related Microbiological Phenomena, Academic, New York, 1978, pp. 345-364

13. S.N. Groudev, F.N. Genchev and S.S. Gaidarjev, in: L.E. Murr, A.E. Torma and J.A. Brierley (eds.) Metallurgical Applications of Bacterial Leaching and Related Microbiological Phenomena, Academic, New York, 1978, pp. 253-272.

14. D.P. Kelly and C.A. Jones, in: L.E. Murr, A.E. Torma and J.A. Brierley (eds.) Metallurgical Applications of Bacterial Leaching and Related Microbiological Phenomena, Academic, New York, 1978, pp. 19-44.

15. J.R. Smith, R.G. Luthy and A.C. Middleton, J. WPCF, 60 (1988) 518


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16. C.A. Jones and D.P. Kelly, J. Chem. Tech. Biotechnol., 33B (1983) 241 17. A.H. Basaran and O.H. Tuovinen, Coal Prep., 5 (1987) 39



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Electrochemical and microbiological characterization of mercury in contact with mud

F. Cruza, A. Welzelb, C. Sampaioc, G.E. Englertc∗, I.L. Müllerc a Material Department - UFRGS - Brazil

b PPGEM - UFRGS - Brazil Metallurgical Department - UFRGS - Brazil

Abstract HgCl2 poured onto liquid nutritive broth was used as culture media for aerobic

bacteria, Bacillus sp. This media was used to evaluate the bacterial activity resistance to toxic metal and to activate the bacteria protective mechanism. Two interconnected flasks were projected to evaluate: i) growth of the bacteria and ii) the electrochemical behaviour of gaseous mercury trapped in acid solution.

Bacteria were cultivated and maintained in plate count agar (PCA), a solid nutritive broth. Electrochemical experiments were done in a liquid culture medium using a Potentiostat PAR EG&G 273A. Potentiodynamic curves were obtained over a platinum surface during 10 days to follow the oxygen depletion by the bacteria. Electrochemical experiments were conducted in nutritive broth with and without mercuric chlorides and bacteria. At the end of the experiments bacteria remained viable despite the contact with toxic metal while the broth pH value changed of 7 through 10.

1. INTRODUCTION It is well known that mercury is a toxic metal and any contaminated with it cannot be

buried due to its hazardous effect. This kind of residues is catalogued as belonging to the Class I. Gaseous monatomic mercury Hg° may be produced by several natural sources and then it is delivered to the atmosphere. Almost 95% of atmospheric mercury is found in the Hg° state. However the fraction value that is free as a monatomic gas or adsorbed onto the solid particles, which are suspended in the air, may change considerably (1). On the other hand, in aqueous media, Hg can be oxidized to HgII as a result of the ozone activity (1). In this case bacteria with an active methylating mechanism seem to regulate environmental conditions that control the bioavailability of HgII. It was proposed in the literature that the accumulation rate of mercury in water depends on pH, selenium concentration, presence of sulfate, dissolved organic carbon in seepage lakes and decomposable organic matter (2). One of the ways of determining the rate of oxidation of mercury is by measuring the concentration of HgII. Some researchers have studied the detection of mercury through bio-indicators but they are not specific to a given form of HgII. It is also possible to detect the presence of mercury by electrochemical methods using platinum electrodes (3).

∗ Universidade Federal do Rio Grande do Sul - Av. Osvaldo Aranha 99/615-D.90035-090 Porto Alegre, RS_Brasil. [email protected] Phone numbers 55 51 3316 3344/3212 4547


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Many researchers have been looking for a technical solution to eliminate mercury from waters using microorganisms (1, 4, 5). Since microorganisms live in environments that can be reproduced artificially, it is possible to isolate them easily. Hence, it is possible to determine the effects that environmental factors play on the availability of inorganic mercury in the presence of bacteria in aquatic environments. Such knowledge is very important from an ecological point of view (3).

2. EXPERIMENTAL PROCEDURE Bacterial strain. A portion of mud was collected from an industrial area in order to

isolate one microorganism adapted to the environmental conditions. They were placed in contact with a liquid broth appropriated for aerobic bacteria. The liquid broth consisted of a solution of peptone, yeast extract and sodium chloride in water at pH 7. During one week the culture was kept at 30°C to produce a quick growth. After this stage, a liquid portion was poured in a dish plate in order to start to the bacterial isolation. The culture was then maintained at the same temperature indicated before. Bacterial identification was carried out by morphological aspects, temperature resistance (after the broth liquid with bacteria was heated up to 80°C for 15 minutes) and biochemical tests (Table 1). The temperature resistance indicates conditions for spore formation and the biochemical tests pointed out to Bacillus sp.

Table 1. Biochemical tests to identify bacteria

Motility + Urease - Citrate + Gelatin +

Nitrate reduction + Glucose + Catalase + Oxidase +

An aqueous solution with bacteria characterized previously was taken after 48 hours of incubation and poured into a fresh portion of modified liquid broth. The modified broth consisted of 1 mL of bacteria solution plus 99 mL of sterilized water. (In order to improve the bacterial growing, oxygen was continuously poured onto the flask). The resistance of the microorganisms to mercury was evaluated adding 1mL of 0.5M HgCl2 and HgCl2 which is a toxic solution to 0.2 L of broth liquid. The later solution pH was reduced from 7 to 4 by adding some drops of 0.1M nitric acid. The viability of the microorganisms was evaluated by counting them at Neusbauer camera. Some other tests were also carried out with the same solution mentioned before at pH 7, with and without mercury.

Electrochemical parameters. Anodic and cathodic potentiodynamic curves, in broth culture with and without mercury, in the presence and in the absence of bacteria, were obtained using a potentiostat PAR G&G 273A.The saturated calomel electrode (SCE) was used as a reference and a graphite was an auxiliary electrode. The working electrode (where electrochemical reactions happened) consisted of a mini-electrode of platinum with an area of 0.00035 cm2. All the experiments started at -1.0V sce and went until +1.0V sce at a scan rate of 2 mV/s. All experiments were done at sterilized and quiescent conditions.


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3. RESULTS AND DISCUSSION Biochemical tests to characterize the bacteria indicate Bacillus sp. as the isolated

bacteria. They are considered resistant to mercury according to literature (1,4,5). In this work it was observed that those species cannot grow as well in modified media as they do in the original media, although both were contaminated by Hg. In the former media the organic concentration was low, since the solution was diluted so the bacteria were in closer contact with mercury. The morphological aspects after experimentation in modified media with Hg are shown in Figure 1, in which the stressed structure is clearly observed. Bacteria that survived in a complete media without Hg were not stressed as can be clearly seen in Figure 2. It was observed that bacteria species in contact with modified broth with Hg survived despite the oxygen concentration, which was reduced progressively after electrochemical experiments (Figure 3).

Figure 1. Bacteria after 30 days in modified media without Hg

Figure 2. Bacteria after 30 days in modified media with Hg


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1E-7 1E-6 1E-5 1E-4 1E-3 0,01 0,1 1 10 100

-1,0

-0,5

0,0

0,5

1,0

1,5 1st day after 5th day

after 10th day 10th dayE

(VSC

E)

i (A/cm2)

Figure 3. Electrochemical curves in modified media with bacteria and without mercury

It was observed that, even reducing the oxygen concentration that was measured by the reduction of the current density, the bacteria were kept alive. The turbidity of the liquid (which remained for more than 10 days without any oxygen reposition) was an evidence that bacteria were still alive and was confirmed when bacteria grew in PCA. From electrochemical point of view it is possible to observe in Figure 3 that at the 5th day of the experimentation, there was a dislocation of the anodic curve to the right hand when, apparently, all oxygen had been removed by bacteria. The shape of the polarization curve is an indicative that the hydrogen cathodic reaction was the cathodic reactive, especially because pH was low. It is well known that the main cathodic reactions from the electrochemical point of view are: the reduction of oxygen dissolved in water and the reduction of the hydrogen from the water or from the acid media. The reintroduction of oxygen to the system brought the curve to the original behaviour with lower current density values. In order to compare the experiments with and without bacteria (data not shown), the same experiment was done with the liquid broth media alone and it was possible to observe that the cathodic current values did not suffer any change. This is a strong indicative that, despite oxygen is by far continuously removed by its electrochemical reduction, remotion is not enough to dislocate the curves as can be seen in Figure 3. Finally, a tiny anodic peak was observed and it may be an indicative that some change happened when comparing the experiments at the first day and the 15th day. As microorganisms grew, the production of their by-products is high enough to interact with the platinum surface. It is known that factors such as low concentration of organic carbon dissolved, the presence of sulphate, low alkalinity and low pH, may interfere in the mechanism by which bacteria retain mercury (2). However, a high concentration of total organic carbon may interfere in the mechanism that bacteria use to reduce mercury. From the anodic curve (data not shown) obtained in the broth liquid with mud, it was not possible to confirm the presence of Hg probably due to the interference to the high organic matter concentration. It is important to highlight that previous analysis through atomic absorption confirmed the presence of Hg in the mud in a concentration of a few ppm. The


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oxygen uptake is partially responsible to the high pH value observed at the end of all experiments since the initial value was 7 and the final one was 10. The participation of the by-products may also contribute to the high pH value observed.

4. CONCLUSION Tested bacteria Bacillus sp. can survive without oxygen. They create self-mechanisms

to produce a new by-product noticed by the changes in cathodic potentiodynamic curves and also in pH values. This observation was confirmed by pH 10 registered at the end of experimentation. The electrochemical tools may be jointly used with the microbiological ones to characterize the growth of bacteria during mercury remotion from contaminated liquids.

ACKNOWLEDGEMENTS To FAPERGS and CNPq for the financial support in the form of scholarship as well

to Prof. Washington Aliaga by his help to structure this work.

REFERENCES 1. T. Smith, K. Pitts, J.A. McGarvey, A.O. Summers. Applied and environmental

microbiology, Apr. 1998, vol. 64, No 4, p. 1328-1332. 2. T. Barkay, M. Gillman, R.R. Turner. Applied and environmental microbiology, Nov.

1997, vol. 63, No 11, p. 4267-4271. 3. Ionashiro, E.Y., Souza, G.R., Milaré, E., Bebedetti, A.V., Ionashiro, M. and Fertonani,

F.L. XII Simpósio Brasileiro de Eletroquímica e eltroanalítica, April, 2001, CDrom, www.ufrgs.br/sibee.

4. A.O. Ogunseitan. Applied and environmental microbiology, Fev. 1998, vol. 64, No 2, p. 695-702.

5. M.D. Mullen, D.C. Wolf, F.G. Ferris. Applied and environmental microbiology, Dec. 1989, vol. 55, No 12, p. 3143-3149.



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Evaluating the growth of free and attached cells during the bioleaching of chalcopyrite with Sulfolobus metallicus

B. Escobar, M.J. Hevia and T. Vargas

Centro de Estudios en Hidro/Electrometalurgia, Departamento de Ingeniería de Minas/Departamento de Ingeniería Química, Universidad de Chile,

Tupper 2069, Santiago, Chile. Fax 56-2-6991084

Abstract Chalcopyrite can be dissolved at covenient rates in the presence of thermophilic

microorganisms such as Sulfolobus metallicus, a chemolitothrophic archaea that can grow autotrophically at temperatures between 65 and 80°C. The catalytic influence of this microorganism is partially related to the microbial oxidation of ferrous iron and the oxidation of residual sulfur produced during chalcopyrite dissolution. Chalcopyrite bioleaching studies, however, normally report only the growth of the planktonic population of Sulfolobus metallicus which makes difficult to assess the pattern of utilization of these two subtrates by the microorganism during the process.

The present work reports experimental studies on the bioleaching of a chalcopyrite concentrate with Sulfolobus metallicus in shake flasks at 70ºC and pH 1.8. The growth of the planktonic population was determined by cell counting, MPN and protein determination by Lowry. The population of microorganisms attached to the solids was evaluated by protein determination using a modified Lowry method.

Results showed that about 68% of the Sulfolobus metallicus present in the flask grows attached to chalcopyrite particles, while the remaining 32% grows in the solution. The percentage associated to attached growth coincides with the fraction of microorganisms which initially attaches to chalcopyrite after addition of the inoculum. Cell growth during the first 72 hours of both attached and free microorganisms is mainly based on the oxidation of ferrous iron. Bacterial oxidation of sulfur, which is conducted only by attached microorganisms, becomes relevant only after 72 hours of leaching. After 120 hours of leaching there is a net reduction of both the planktonic and attached population which evidences an important process of cell death and breakage.

Keywords: S. metallicus, bioleaching, chalcopyrite, biomass estimation, attached cells

1. INTRODUCTION Primary copper sulfides such as chalcopyrite has been demonstrated to be refractory

to bacterial leaching with mesophiles such as At. ferrooxidans in ambient conditions (1). During the last years, however, several researchers have demonstrated that chalcopyrite can be dissolved at covenient rates in the presence of thermophilic microorganisms such


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as Sulfolobus metallicus, Acidianus brierley or Metallosphera sedula (2, 3). The use of Sulfolobus metallicus, a chemolitothrophic archaea that can grow autotrophically at temperatures between 65 and 80°C, has been largely investigated in the bioleaching of chalcopyrite (4-6). The catalytic influence of S. metallicus in this process is partially related to the microbial oxidation of ferrous iron, an activity which has been evidenced by several authors (1, 7). Other works have also indicated that S. metallicus presence have also a catalytic influence on the oxidation of residual elemental sulfur to sulfate (7,8). In chalcopyrite bioleaching studies only the growth of the planktonic S. metallicus is normally monitored, which makes difficult to assess the relative importance of these two substrates.

The present work reports experimental studies on the bioleaching of a chalcopyrite concentrate with Sulfolobus metallicus in shake flasks at 70ºC and pH 1.8. An important methodological aspect, the growth of both the population of microorganisms attached to the sulfide particles and microorganisms free in solution were simultaneously monitored during these experiments. The relative importance of sulfur and ferrous iron as substrates for microbial growth was evaluated from relating bacterial growth to the evolution of dissolved metals and pH during the experiment.


2.1 Microorganism A strain of Sulfolobus metallicus was used in this study. The cells were grown in

basal medium containing 1% chalcopyrite concentrate as the energy source. The composition of the basal medium was 0.4 (NH4)2SO4, 0.5 MgSO4

.7H2O and 0.2 KH2PO4 g/l, adjusted to pH 1.6 with sulfuric acid.

2.2 Mineral A -325 +400 mesh (-0.045mm +0.038mm) sample of a high purity chalcopyrite

sample prepared from an Andina Copper concentrate was used in the experiments.

2.3 Experimental procedure Attachment experiments were conducted in 125 ml Erlenmeyer flasks containing 50

ml basal medium. One g of chalcopyrite was contacted with an initial concentration of 1.1x108 cells of S. metallicus/ml. The flasks were incubated in an environmental shaker at 70°C and 100 rpm. The rate of cell attachment was determined by periodic sampling and determination of free cells by direct counting onto the microscope.

Leaching experiments were conducted in 250 ml Erlenmeyer flasks containing 95 ml of the basal medium, 1% (w/w) chalcopyrite and 5 ml of active inoculum of Sulfolobus metallicus. The flasks, 8 in total, were incubated in an environmental shaker at 100 rpm and 70°C. Two flasks were removed at different leaching times to determine total iron, ferrous iron, copper and pH in solution, and bacterial growth in solution and solids. Total and ferrous iron in the solution were determined by o-phenanthroline colorimetric method (9); for solutions with high ferric:ferrous ratio ferrous iron was determined by the modified o-phenanthroline colorimetric method (10). The number of free cells in solution was determined by direct counting onto the microscope using a Petroff-Hausser chamber and by the MPN of iron oxidizing microorganisms, in which cells were grown at 70°C in basal medium containing 3.6 mM ferrous sulfate during 10 days. Also free cells were estimated by protein determination using the Lowry methodology (11). Attached


1093

microorganisms were also estimated by protein assay using the modified Lowry method proposed by Karan et al., (12). In order to correlate the amount of proteins determined on the solids to number of cells, a calibration curve obtained from measurements of planktonic cells with both methods, direct counting and modified Lowry was used (see Figure 1).


3.1 Adherence of S. metallicus to chalcopyrite Figure 2 shows the evolution of the population of free bateria in solution during

attachment of Sulfolobus metallicus to chalcopyrite. The attachment of S. metallicus onto chalcopyrite concentrate was found to reach a stationary value within the first 100 minutes of contacting the cells with mineral, when 68% of the initial cells were attached to the mineral.

0 .0 E + 0 0

2 .0 E + 0 8

4 .0 E + 0 8

6 .0 E + 0 8

0 1 0 2 0 3 0P ro te in (m g /m l)

Cel

l/ml

0 .0E+00

2.0E+09

4.0E+09

6.0E+09

8.0E+09

0 50 100 150

Time (min)

Tota

l Cel

l/50

ml

Figure 1. Correlation between cell number of S. metallicus and measured protein

Figure 2. Attachment of S. metallicus onto 1 g chalcopyrite concentrate

3.2 Bioleaching of chalcopyrite Figure 3 shows the rate of copper and iron dissolution, determined during the

bioleaching of chalcopyrite with Sulfolobus metallicus. The total amount of dissolved iron, 8.86 mg, was much smaller than the total amount of dissolved copper, 134 mg. Ferrous/ferric determinations indicated that iron released during chalcopyrite dissolution was quantitatively oxidized to ferric iron from the start of the experiment (results not included), which indicated an effective microbial catalytic activity. Very likely, most of the iron released during chalcopyrite dissolution was eventually precipitated, which showed in the presence of large amounts of jarosite-like compounds either in the bioleaching solution or adhered to the flasks wall. The amount of ferric iron precipitated at each time corresponds to the difference between the effective and measured iron concentration, the theoretical iron dissolution expected according to the stoichiometry: CuFeS2 + 4Fe+3 → Cu+2 + 5Fe+2 + 2S (1)

The steady increase of this difference indicates that ferric iron precipitation occurred continuously since the start of the experiment.

The evolution of the number of cells and biomass of planktonik cells of S. metallicus during the experiment is shown in Figure 4. Figure 5 shows the evolution of


1094

microorganisms attached to the solids presents in the flasks, solids which includes residual chalcopyrite together with the ferric iron precipitates which were simultaneously separated from the solution by filtration. These results show that planktonic cells gives only account of 30% of cell growth occurring during the process and, consequently, about 70% of bacterial growth occurs in the cells attached to the solids. The relative importance of these two populations is similar to the distribution of cells obtained during the initial attachment stage, shown in Figure 2.

0

2 0 0

4 0 0

6 0 0

8 0 0

1 0 0 0

1 2 0 0

1 4 0 0

1 6 0 0

0 1 0 0 2 0 0 3 0 0T im e (h o u rs )

Cu

and

Tota

l Fe

(mg/

l)

0.E+00

2.E+08

4.E+08

6.E+08

0 100 200 300

Time (hours) C

ell/m

l

0 .E+00

2.E+07

4.E+07

6.E+07

Prot

ein/

ml

Figure 3. Copper , total iron measured ∆ and effective total iron disolution in bioleaching of chalcopyrite concentrate

Figure 4. Cells of S. metallicus in bioleaching solution determined by direct count , MPN and protein concentration

0.E+00

1.E+10

2.E+10

3.E+10

4.E+10

0 100 200 300

Time (hours)

Total

Cell

and

Adhe

red

Cells

0

20

40

60

80

100

120

140

160

0 100 200 300

Time (hours)

Sulfu

r (m

g)

Figure 5. Total and adhered cells of S. metallicus in the bioleaching chalcopyrite concentrate

Figure 6. Sulfur generated (curve a) and oxidized (curve b) estimated from copper and acidity of solution

The cell population of planktonic and attached cells initially increases reaching a maximum at 72 and 120 hours, respectively. After this initial stage there is a rapid decay in the population of both free and attached microorganisms.

In the case of planktonic cells, the decrease in cellular activity, determined by MPN, is paralelled by a decrease of the number of cells, determined by counting. This indicates that the decrease in cellular activity is not due to some type of inhibition, but rather to a process of cell death and/or cell breakage. This phenomenon can be linked to the sensitivity which these microorganisms show to the presence of solids, as it is widely known (3,13). Elimination of cells by incorporation into ferric iron precipitates can be


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ruled out as a main cause of cell number reduction, as this would show in a simultaneous increase in the biomass of attached bacteria, which was not observed.

It is interesting to outline that the decrease in cell number and activity of planktonic cells is paralelled by a decrease in protein concentration. This indicates that the constituent proteins of thermophilic microorganisms, once the cell breaks, are not stable at the temperature and acidity at which bioleaching solutions operate. This fact was confirmed in a complementary experiment in which Sulfolobus metallicus digested cell proteins were maintained in a shaker in an iron-free basal solution at pH 1.8 and at 70°C. The protein concentration in the solution after 72 hours in these conditions, determined by Lowry, decreased 28% with respect to the initial value.

The decay in protein concentration observed on solids indicates that in the attached cell population also a stage is reached in which the rate of cell death and/or breakage predominates over the rate of net cell growth.

3.3 Bacterial utilization of ferrous iron and sulfur Ferric iron was not initially added in the reported experiments, hence the scarce ferric

iron present in the solutions resulted from the microbial oxidation of ferrous iron released during chalcopyrite dissolution. Bioleaching solutions presented a high Eh (data not shown), indicating a high ferrous iron oxidation activity of S. metallicus. The efficiency of this oxidative activity was also evidenced from the fact that MNP cell number determinations, which were based on the reponse of microorganisms to ferrous iron oxidation, gave similar values to those determined by direct cell counting. Attached cells, however, are exposed to two substrates, ferrous iron and surface elemental sulfur formed after reaction (1). It is interesting to assess the pattern of utilization of these two substrates during bacterial growth and see how this relates to the chalcopyrite bioleaching process.

The pH of the bioleaching solution is the result of dinamic balance linked to a series of chemical reactions which consume or produce protons. Protons are consumed during the oxidation of ferrous iron which occurrs according to: Fe+2 + 1/2O2 + 2H+ → Fe+3 + H2O (2)

On the other hand, protons are produced through: a) hydrolysis of the Fe(III) with jarosite formation according to the reaction: M+ + 3Fe+3 + 6H2O + 2(SO4)-2 → MFe3(SO4)2(OH)6 + 6H+ (3)

Where M+ is either K+, NH4+ or H3O+; and through b) oxidation of sulfur to sulfate

according to the reaction: 2S° + 3O2 + 2H2O → 2H2SO4 (4)

In order to assess the amount of elemental sulfur oxidized by Sulfolobus metallicus to sulfate, the acidity resulting from this reaction was evaluated at different times during chalcopyrite bioleaching with the expression: ∆4 = Ct - Ci - ∆3 + ∆2 (5) where ∆4 is the acid generated by reaction 4, Ct is the acid concentration determined in the solution from pH measurement, Ci is the initial acid concentration, ∆3 is the acid produced in reaction 3, and ∆2 the acid consumed in reaction 2. The amount of ferrous released at each time was calculated from the amount of copper dissolved from chalcopyrite, assuming that this follows the stoichiometry of reaction (1). Ferrous iron was assumed to be quantitatively converted into ferric according to reaction (2). The amount of ferric iron precipitated was calculated as the difference between the theoretical amount of ferrous


1096

iron dissolved and the iron experimentally determined in solution (Figure 3). The acid generation associated to ferric iron precipitation was calculated assuming formation of potassium jarosites which, according to XRD analysis, is the compound we have found that predominate in concentrates bioleached with thermophiles.

Results of this calculation are shown in Figure 6. Curve a represents the amount of elemental sulfur generated during chalcopyrite dissolution according to reaction (1), and was calculated from the amount of copper dissolved at different leaching times. Curve b represents the amount of elemental sulfur oxidized to sulfate calculated from reaction (4) using the ∆4 values as generated acid. Comparison of data in Figure 6 with the evolution of cell growth in Figures 4 and 5 indicate that the fast cell growth observed during the initial 72 hours in both the planktonic and attached microorganisms, is mainly based on the oxidation of ferrous iron. The growth is faster on the solids as the attached initial population is larger. In addition, cells attached to chalcopyrite surface should have access to larger concentrations of ferrous ions which is released in that vicinity.

The oxidation of elemental sulfur becomes significant only after the first 72 hours when already 50 mg of this compound have been accumulated on the chalcopyrite surface. At this stage the growth of the attached population continues fast for another 50 hours while the planktonic population started to decrease. It is very likely that the continued growth trend observed in the attached microorganisms is partially linked to triggering of sulfur oxidation, as these microorganisms are the only ones with can have direct access to that alternative solid substrate. On the other hand, the decrease in the planktonic population after the first 72 hours can be due to depletion of ferrous iron in the solution bulk. This can be related to the large population of attached microorganisms reached at this stage, which can consume most of the ferrous iron released from the sulfide.

Studies done by several authors about the bioleaching of pyrite and chalcopyrite with thermophilic microorganisms show differences on the cells adsorption to the mineral; in some cases, in spite of the fact that the cellular attachment was not detected by microscope, they showed that pyrite leaching required the direct contact of Acidianus brierley with the mineral (14). Nemati and Harrison (15), working with S. metallicus and pyrite, showed a low cell adsorption (5-10%) and considered the concentration of the cells in the liquid phase as indicative of total microbial growth in the reactor. On the other hand, Konishi et al., (16) showed a high attachment of cells to chalcopyrite and found that the chalcopyrite leaching with A. brierley at 65°C takes place with a direct attack by adsorbed cells on the mineral surface, the ferric leaching being insignificant. The results presented in this work show that most of the population of S. metallicus present in a bioreactor remains attached to the solids, where they grow out of ferrous iron and sulfur oxidation. Any attempt to assess in depth the performance of a bioleaching reactor should include an adequate monitoring of this population.

4. CONCLUSIONS Experimental results demonstrated that the growth of attached Sulfolobus metallicus

can be efficiently monitored using the modified Lowry method of Karan et al.,(16). During bioleaching of chalcopyrite about 68% of the Sulfolobus metallicus present in

the flask grows attached to chalcopyrite particles, while the remaining 32% grows in the solution. The percentage associated to attached growth coincides with the fraction of microorganisms which initially attaches to chalcopyrite after addition of the inoculum.

Cell growth during the first 72 hours of both attached and free microorganisms is mainly based on the oxidation of ferrous iron. Bacterial oxidation of sulfur, which is


1097

conducted only by attached microorganisms, becomes relevant only after 72 hours of leaching.

After 120 hours of leaching there is a a net reduction of both the planktonic and attached population which evidences an important process of cell death and breakage.

ACKNOWLEDGEMENTS This work was funded by BHP Billiton and Conicyt Chile under Fondef Project D 00

I 1050.

REFERENCES 1. P.R. Norris, Biomining: Theory, Microbes and Industrial Processes, D.E. Rawlings

(ed), Landes, Bioscience, Austin, TX USA. (1997) 281. 2. P.R. Norris, L. Parrot and R.M. Marsh, Workshop on Biotechnology for the Mining,

Metal-refining and Fossil Fuel Processing Industries, Biotechnol. Bioeng., Symposium Nª16, H.L. Ehrlich and D.S Holmes (eds), Wiley, New York (1986) 253.

3. D.A. Clark and P. Norris, Microbiology 142 (1996) 785. 4. P. d’Hugues, D. Morin and S. Foucher, Biohydrometallurgy: Fundamentals,

Technology and Sustainable Development, Part A, V.S.T. Ciminelli and O. Garcia (eds), Elsevier, Amsterdam (2001) 75.

5. M.L. Blázquez, A. Alvarez, A. Ballester, F. González and J.A. Muñoz, Biohydrometallurgy and the Environment toward the mining of the 21st Century, A. Ballester and R. Amils (eds.) Elsevier, Amsterdam (1999) 137.

6. Y. Rodriguez, A. Ballester, M. L. Blázquez, F. González and J.A. Muñoz, Biohydrometallurgy: Fundamentals, Technology and Sustainable Development, Part A, V.S.T. Ciminelli and O. Garcia (eds), Elsevier, Amsterdam (2001) 125.

7. R. M. Marsh, P.R. Norris and N. W. Le Roux, Progress in Biohydrometallurgy, G. Rossi and A.E. Torma (eds.) Cagliari, Italy (1983) 71.

8. D.W. Shivvers and T. D. Brock, J. Bacteriol., 114 (1973) 706. 9. M. K. Muir and T. Anderson, Metall. Trans. 8B (1977) 517. 10. L. Herrera, P. Ruiz, J.C. Aguillón and A. Fehrmann, J. Chem. Tech. Biotechnol 44

(1989) 171. 11. O.H. Lowry, N.J. Rosenborough, A.L. Farr and R.J. Randall, J. Biol. Chem., 193

(1951) 265. 12. G. Karan, K.A. Natarajan and J.M. Modak, Hydrometallurgy 42 (1996) 169. 13. B. Escobar, J.M. Casas, J. Mamani and R. Badilla-Ohlbaum Biohydrometallurgical

Technologies, A. E. Torma, J.E. Wey and V.I. Lakshmanan (eds.) The Minerals, Metals & Materials Society (1993) 195.

14. L. Larsson, G. Olsson, O. Holst and H. T. Karlsson, Biotechnol. Lett. 15 (1993) 1, 99. 15. M. Nemati and S.T.L. Harrison, Biohydrometallurgy and the Environment toward the

mining of the 21st Century, Part A, A. Ballester and R. Amils (eds) Elsevier, Amsterdam (1999) 473.

16. Y. Konishi, M. Tokushige and S. Asai, Biohydrometallurgy and the Environment toward the mining of the 21st Century, Part A, A. Ballester and R. Amils (eds) Elsevier, Amsterdam (1999) 367.



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Experimental and modeling studies on inhibition effect of solution conditions on activity of Acidithiobacillus ferrooxidans

during biooxidation of mixed sulphidic concentrates

M.N. Chandraprabhaa, Jayant M. Modakb and K.A. Natarajana

a Department of Metallurgy, Indian Institute of Science, India b Department of Chemical Engineering, Indian Institute of Science, India

Abstract In this paper, we report experimental and modeling results aimed at characterizing the

effect of solution conditions on the growth and ferrous iron oxidation ability of Acidithiobacillus ferrooxidans. The conditions we have chosen pertain to the bioleaching of mixed sulphidic (pyrite and arsenopyrite) gold ore concentrates resulting in the accumulation of ferric and arsenic ions in the solution. Apart from the dissolved metal species, the pH and CO2 concentrations in the solution are also important parameters in bioleaching processes.

The iron oxidizing efficiency of the strain increases with increasing carbon dioxide supplementation of the sparged air upto 1% (V/V), beyond which the bacterial activity reduces. Maintaining lower growth pH (1.6) reduced the formation of precipitates to a large extent.

Ferric and arsenic ions are found to be toxic to the bacterium and strains tolerant to higher concentrations of these metal ions could be obtained by adaptation. The growth kinetics of the bacterium has been modeled incorporating the ferric and arsenic inhibition. The adaptation of Acidithiobacillus ferrooxidans to higher metal concentrations is quantified by gradual reduction in inhibition constants. The simulation results were found to match well with the experimental data obtained. The applicability of the inhibition constants obtained to the batch bioleaching model of a mixed sulphidic concentrate has been analysed.

Keywords: Acidithiobacillus ferrooxidans, biooxidation, inhibition, adaptation, modeling

1. INTRODUCTION Biooxidation rate depends to a large extent on the efficiency of the bacterial strain

used. The regeneration of ferric iron by bacterial oxidation of ferrous sulphate in the solution phase is an essential step in the biooxidation process. The oxidation rate is, therefore, directly related to the activity of the bacteria in solution phase, which in turn is determined and driven by the solution conditions.

Bioleaching experiments results in the formation of ferric and arsenic hydroxides that precipitate at pH 2. One way of minimizing the formation of these precipitates is by maintaining a low process pH. Studies by several workers (Jose et al, 1997; Kelly, 1978;


1100

Nagpal et al, 1992) have clearly indicated the reduction in precipitate formation by lowering pH. Since the bacterium employed has an optimum pH of 2, it has to be suitably adapted to the low pH required in order to retain its optimal activity. Studies done by Toxiarchou et al, (1994) on the extent of arsenopyrite and pyrite oxidation at various pH values showed that although the selective oxidation of arsenopyrite is enhanced at low pH, pH values lower than 1.6 adversely affects the overall oxidation kinetics since under these conditions the oxidation of pyrite is hindered. Hence a pH of 1.6 would be optimal for reducing the precipitate formation without hindering the oxidation kinetics.

Oxygen and carbon dioxide are essential nutrients for bacterially catalyzed mineral oxidation and bacterial growth. Reduction in oxygen concentration in the aqueous phase below optimum levels leads to reduced growth rates. Similarly, reduction of aqueous phase carbon dioxide also sharply reduces the bacterial growth. Hence, the effect of CO2 enrichment on the growth of bacteria needs to be studied.

Various toxic substances, often inherently present in the biooxidation of a sulphide mineral, have been found to inhibit the oxidation ability of bacteria (Bailey and Hansford, 1993). During biooxidation process, the concentration of the released metal ions in the solution continually increases and hence there will be build-up of ferric and arsenic ions. Both ferric and arsenic ions are toxic to the bacterial activity and hence the growth of bacteria in solution phase in presence of these metal ions needs to be examined. This detrimental effect can be overcome by adapting the bacteria to higher concentrations of these metal ions (Modak and Natarajan, 1995).

The objective of our present work was to analyse the effect of solution conditions on the activity of the bacteria, model the growth kinetics of the bacterium incorporating the inhibition effects of the toxic metal ions present in the solution and to analyse the applicability of this to the batch bioleaching model of a mixed sulphidic concentrate.

2. EXPERIMENTAL

2.1 Materials The bacterial culture used was a pure strain of Acidithiobacillus ferrooxidans that was

isolated from Hutti Gold Mines (HGML) and is referred to as TfH6. The purity was ascertained by the procedure outlined by Karavaiko (1988). The bacteria were cultured in sterile 9K medium developed by Silverman and Lundgren (1959). The refractory sulphidic concentrate obtained from HGML was in finely ground form with an average particle diameter of 70 µm. The chemical composition was 31.95 wt% Fe, 29.25 wt% S and 22.5 g/MT gold.

2.2 Methods

2.2.1 Growth of bacteria The growth experiments were carried out by inoculating 10%(V/V) of active

inoculum to sterilized 9K medium. The flasks were incubated at 30°C on a rotary shaker at 240 rpm. The ferrous and the total iron concentration were monitored at regular intervals by spectroscopic method. The solution cell number was determined by microscopic counting with Petroff-Hausser counter using a Leitz phase contrast microscope (Laborlux K Wild MPS12). pH changes were also monitored simultaneously using a Systronics digital pH meter.


1101

2.2.2 Effect of carbon dioxide on growth The effect of carbon dioxide availability on the activity of bacteria was monitored by

measuring the iron oxidization rate of bacteria at different levels of carbon dioxide supplementation. These experiments were carried out in 17L bioreactors with the operating volume of 6L. The aeration rate and the carbon dioxide flow rate were monitored using suitably calibrated rotameters.

2.2.3 Growth inhibition and adaptation studies The inhibition effect of ferric iron and arsenic on the growth of bacteria was studied

by adding ferric sulphate or sodium arsenite at different concentrations to the flasks containing 9K medium solution, prior to inoculation. The iron oxidation rate of these cultures were monitored and compared with that of the control without inhibition.

Adaptation to lower pH was done by serial subculturing of the bacterial strain in 9K medium maintained at the required pH. Adaptation of the bacteria to the toxic ferric and arsenic ions was carried out by serial subculturing of bacteria in 9K medium in presence of externally added ferric sulphate and sodium arsenite as the sources of ferric and arsenic ions respectively.

2.2.4 Batch bioleaching experiments The bioleaching experiments were performed in batch mode in well-agitated and

aerated 17L bioreactor with an operating volume of 6L. The reactor was aerated at the rate of 0.5 v/v/min with 1% carbon dioxide enrichment. The experiments were started with a low pulp density of 1%(W/V), which was then increased in stepwise with periodic addition of fresh concentrate when the concentrate in the reactor was completely oxidised. Complete oxidation was considered to be achieved when the total iron in the solution approached the theoretical stoichiometric value estimated.


3.1 Growth of Acidithiobacillus ferrooxidans The growth of a typical culture of Acidithiobacillus ferrooxidans (TfH6) on ferrous

iron with time under optimum conditions of temperature (30°C) and pH (2.0) is shown in Figure 1a. No significant increase in cell density was observed during the initial lag period of 20hrs, beyond which the growth followed exponential kinetics. The growth remained in log phase till the 50th hr after which the cells entered the stationary phase. The maximum cell number was reached in 48-50 hrs.

3.2 Effect of pH Lowering of pH did not have any significant effect on the activity of the bacterial

strain studied and just after one subculturing the activity of the adapted strain matched with that of the wild strain as seen from Figure 1b. Comparing the growth curves at pH 2 and pH 1.6 (Figures 1a and 1b), we find that at the end of exponential phase when all the ferrous iron is oxidized, the ferric iron concentration in the flask maintained at low pH is 8.9 g/L, which is higher than that observed at pH 2 (6.3 g/L). This clearly indicates that the precipitation of ferric hydroxide found at pH 2 is reduced to a large extent by lowering the pH.


1102

0 10 20 30 40 50 60 700

2

4

6

8

10 (a)

Iron

conc

entra

tion

(g/L

)

Time (h)

Fe2+

Fe3+

0 10 20 30 40 50 60 700

2

4

6

8

10 (b)

Iron

conc

entra

tion

(g/L

)

Time (h)

Fe2+

Fe3+

Figure 1. Growth of Acidithiobacillus ferrooxidans at (a) pH 2 and (b) pH 1.6

3.2 Effect of CO2 on growth of organism In order to study the effect of CO2 on the growth of bacteria, the growth experiments

were conducted in 17L bucket bioreactor at different levels of CO2 supplementation (vol CO2/vol air). The results obtained are plotted in Figure 2. It is clear from the figure that the supplementation of CO2 to the sparged air resulted in enhancement of growth rate. In the absence of CO2 supplementation, the bacteria were able to oxidize the ferrous iron in 9K medium in 52 hours. With 0.5% CO2 supplementation, the time taken for complete growth was just 42 hours, which reduced to 38 hours at 1% CO2 supplementation. The cell yield also showed slight increase with increase in CO2 supplementation. However beyond 1% CO2 supplementation, there was no much benefit since the oxidation rate was less at 2% CO2 supplementation as compared to 1% CO2 supplementation.

0 10 20 30 40 50 60 700

2

4

6

8

10(a)

Ferro

us ir

on c

once

ntra

tion

(g/L

)

Time (h)

0% CO2 0.5% CO2 1% CO2 2% CO2

0 10 20 30 40 50 60 700

5

10

15

20

25(b)

Cel

l num

ber X

1010

(cel

ls/L

)

Time (h)

0% CO2 0.5% CO2 1% CO2 2% CO2

(a) Ferrous iron concentration (b) Cell density

Figure 2: Effect of CO2 supplementation on growth of Acidithiobacillus ferrooxidans


1103

Nagpal et al., (1992) have shown that bacteria grown on a sulphide ore concentrate utilize oxygen and carbon dioxide in a molar ratio of approximately 20 mol oxygen per mole CO2. This suggests that air supplemented with 1% CO2 would provide the appropriate molar ratio of the two nutrients. During their tests at 16% solids, carbon dioxide concentrations in the liquor in excess of 10mg/L were found to be inhibitory to bacterial growth. The obtained results are thereby, in good agreement with that reported.

3.4 Effect of ferric and arsenic inhibition In order to study the effect of Fe3+ on ferrous oxidation, the growth of TfH6 was

monitored with varying concentrations of Fe3+ added initially to the nutrient medium. The results of the experiments with 2, 5 and 8 g/L of initial Fe3+ is presented in Figure 3a. While 2 g/L initial Fe3+ had no significant effect on growth, an extended lag phase was observed with 5 g/L and 8 g/L initial Fe3+ concentration. This is accompanied by reduced rates of ferrous oxidation and cell growth.

The arsenic ions coming to the solution due to solubilization of arsenopyrite is in the form of arsenite [As3+] and arsenate [As5+] ions. Since arsenate ions are less toxic compared to arsenite ions, the inhibition effect of only arsenite ions is considered in our studies. Figure 3b, shows the cell growth and ferrous iron oxidation by TfH6 in presence of different initial concentrations of arsenite ions. From the figure we can observe that the wild strain is naturally tolerant upto 400 ppm of arsenite concentration. Further increase in arsenite concentration resulted in extended lag phase and decrease in cell yield. The extended lag phase was significant when the initial arsenite concentration maintained was 1000 ppm and the wild strain took more than 58hrs to oxidize 9 g/L ferrous. It was observed that the total cell yield was less at arsenite concentrations of 500 and 1000 ppm.

0 10 20 30 40 50 60 700

2

4

6

8

10(a)

Fe3+ concentration

Ferr

ous

iron

conc

entra

tion

(g/L

)

Time (h)

0 g/L 2 g/L 5 g/L 8 g/L

0 10 20 30 40 50 60 700

2

4

6

8

10(b)

As3+ concentration

Ferr

ous

iron

conc

entra

tion

(g/L

)

Time (h)

0 ppm 100 ppm 200 ppm 500 ppm 1000 ppm

Figure 3: Effect of (a) ferric and (b) arsenite ions on growth of Acidithiobacillus ferrooxidans

3.5 Bacterial adaptation studies From Figures 3, it is clear that the growth of TfH6 is inhibited in presence of high

initial ferric and arsenic concentration. However, repeated subculturing in presence of these toxic ions results in reduced lag period and increased growth rate due to its adaptation. The bacterial strain is said to be adapted when the cell growth and iron oxidation rate of the strain in presence of the externally added toxic metal ion is


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comparable with that of the control, which is the growth rate of the strain in the absence of any inhibition. The results obtained in presence of 8 g/L Fe3+ and 1000 ppm As3+ is shown in Figure 4. The growth rate of the wild bacterial strain in presence of the externally added toxic metal ion is represented as its natural tolerance. When grown in presence of 8g/L Fe3+ growth rate after two subcultures matched with that of the wild strain, thereby indicating that the strain is now adapted to 8 g/L Fe3+. Hence, with just two subcultures it was possible to obtain a strain tolerant to 8 g/L Fe3+. As evident from Figure 4b, it was possible to obtain a strain tolerant to 1000 ppm As3+ by subculturing the wild strain thrice in presence of 1000 ppm As3+.

0 10 20 30 40 50 60 70

0

5

10

15

20

25(a)

Cel

l num

ber X

1010

cel

ls/L

Time (h)

Natural tolerance 1 subculture 2 subculture No inhibition (control)

0 10 20 30 40 50 60 70

0

5

10

15

20

25(b)

Cel

l num

ber X

1010

cel

ls/L

Time (h)

Natural tolerance 1 subculture 2 subculture 3 subculture No inhibition

Figure 4. Adaptation of Acidithiobacillus ferrooxidans to (a) 8 g/L ferric; (b) 1000 ppm arsenite

3.6 Modeling of bacterial growth The model for the growth of bacteria on ferrous iron should be able to predict the

rates at which ferrous is being oxidised under the changing conditions of ferrous and ferric concentrations and the cell densities. The growth rate of bacteria in liquid is assumed to be described adequately by Monod's model (Boon et al, 1999; Jensen and Webb, 1995) with ferrous iron as the substrate, that is, the rate of substrate utilization is given by:

22

2( )m

S

FedFedt Y K Fe

µ ++

+

−=

+ (1)

where µm is the maximum specific growth rate in hr-1 and KS is the saturation constant for ferrous iron in g/L respectively.

Y is the yield coefficient given by, 2dX dFeY

dt dt

+

= − (2)

where X represents cell density in solution in cells/L. Both ferric and arsenic ions inhibit the growth of Acidithiobacillus ferrooxidans

competitively without affecting the maximum specific growth rate. The rate is thus given by the below equation (Bailey and Olis, 1986),


1105

22

3 3 2( (1 ) )m

S I a

FedFedt Y K K Fe K As Fe

µ ++

+ + +

−=

+ + + (3)

where KI is the inhibition constant for ferric iron and Ka for arsenic ion in L/g. Experimental data of growth in absence of any inhibition was used to fit the

parameters of the above equation and the simulation result obtained is shown in Figure 5. The values of µm and KS obtained are 0.07 hr-1 and 1.6 g/L respectively.

(a) Ferrous iron oxidation (b) Cell growth

Figure 5. Growth of Acidithiobacillus ferrooxidans at 30°C and pH 1.5

The experimental data obtained at different levels of CO2 supplementation were used in fitting the parameters of the Monod equation and it was found that the supplementation of CO2 affects the maximum specific growth rate. The µm obtained at different levels of supplementation are given in Table 1, from which it is clear that the optimum level of supplementation is 1% V/V of CO2.

The growth rate data obtained in presence of 8 g/L Fe3+ and 1000 ppm As3+ was used to obtain the inhibition parameters in equation 3. Since both Fe3+ and As3+ ions inhibit competitively, µm remains unaffected. A comparison of the model parameters obtained with that reported in literature is given in Table 2. As see from the table, the values obtained are in close agreement with the reported values (Jensen and Webb, 1995; Lizama and Suzuki, 1989).

Table 1. Variation in µm with CO2 supplementation

Table 2. Model parameters of bacterial growth

Percent CO2 supplementation µm, hr-1 Parameter Estimated

value Reported value

0 0.07 µm (hr-1) 0.09 0.05 - 1.3 0.5 0.0825 Ks (g/L) 1.6 1 - 2 1 0.0925 Y (cells/mol Fe2+) 1.36x1012 9x1013 - 5x1014 2 0.0785 KI (L/g) 0.14 0.5 - 2.3 Ka (L/g) 1.15 0.45 - 0.6

The comparison between model simulation using the parameters listed in Table 2 and the experimental data is shown in Figure 6. The bacteria, on continuous exposure, develop tolerance to ferric and arsenic ions as discussed earlier. The effect of bacterial adaptation


1106

to ferric and arsenic ions on the inhibition constants KI and Ka is shown in Figure 6. From the figure we observe that with repeated subculture, the value of the constants approaches zero.

Figure 6. Variation of inhibition constants with adaptation of Acidithiobacillus ferrooxidans to (a) 8 g/L ferric; (b) 1000 ppm arsenic

3.7 Application to batch leaching model A complete batch bioleaching model for a mixed sulphidic concentrate was developed

earlier by considering the various subprocesses independently to obtain the model parameters (Chandraprabha et al, 2001). In this model growth of bacteria was described by equation 1, neglecting the inhibition effect of ferric and arsenic ions coming into the solution. Here the effect of the inhibition constants on the model prediction of the batch biooxidation of 1% pd pyrite-arsenopyrite concentrate was tested and the results obtained are shown in Figure 7. The bacterial growth in solution was described by equation 3 and the parameters listed in Table 2 were used for simulation. As seen from the figure, the model overpredicts the ferric iron concentration when the inhibition effect of ferric and arsenic ions is neglected (dotted lines).

The presence of ferric and arsenic ions in the solution thereby inhibits the bacterial activity. After 5 days of leaching, when the concentrate was almost completely oxidized, the pulp density was increased from 1% to 2% by addition of fresh concentrate. Simulation done at 2% pd also showed similar behaviour (data not shown). The slurry was maintained at 2% pd for 5 days after which it was increased to 3% by addition of fresh concentrate. The results for 3% pulp density are shown in Figure 7b. During the course of biooxidation the bacteria are subjected to increasing concentrations of the toxic ferric and arsenic ions coming from the concentrate and grow in presence of these stressed conditions. When solid pulp density reached 3%, the bacteria were exposed to total of about 240 h. As revealed by our adaptation studies, the bacteria attain tolerance at 3 subcultures (approximately 200 h). Hence the later generations of bacteria are adapted strains that are tolerant to these toxic ions. This is clear from Figure 7b, which shows the simulation results obtained at 3% pd concentrate with KI and Ka values set to zero. Here we find that the prediction obtained by neglecting the inhibition effects of ferric and arsenic fits well with the experimental data points.


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Figure 7. Simulation of batch bioleaching at (a) 1% pd and (b) 3% pd concentrate. Solid lines represent model curves with ferric and arsenic inhibition and dashed lines without any inhibition. Symbols represent experimental data points

4. CONCLUSIONS The effect of solution conditions on the activity of Acidithiobacillus ferrooxidans is

reported and the following conclusions can be arrived at: Lowering of process pH resulted in reduction of the precipitate formation, thereby

enhancing bacterial oxidation rates. Supply of air supplemented with CO2 enhanced the growth kinetics of the strain used.

The iron oxidation rate increased with increase in CO2 supplementation upto 1% (V/V), beyond which the cell yield decreased.

Presence of both ferric and arsenic ions inhibited the bacterial growth. Serial subculturing in presence of these metal ions resulted in adapted strains that were tolerant to higher concentrations of the metal ions.

The growth kinetics of bacteria was modeled incorporating the inhibition constants obtained for ferric and arsenic inhibition and it was found that with repeated subculturing the value of the inhibition constant approached zero which clearly indicated the adaptation of bacteria to the toxic metal ions due to continuous exposure.

The model developed for the batch bioleaching of pyrite-arsenopyrite concentrate was found to overpredict the ferric iron concentration in the solution at 1% pd when the effects of ferric and arsenic inhibition were neglected. This was overcome by incorporating the inhibition constants in the growth equation of bacteria.

The predictions obtained at later stages at a higher pulp density were found to fit well with the experimental data even when the inhibition effects were neglected. This was due to the adaptation of the bacteria to the released metal ions in solution that was achieved due to their continuous exposure to these ions.

REFERENCES

1. Bailey, J.E. and Hansford, G.S., 1993. Factors affecting bio-oxidation of sulphide minerals at high concentrations of solids: A Review. Biotechnol. Bioengg., 42, 1164-1174.


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2. Bailey, J.E. and Olis, D.F., 1986. Biochemical engineering fundamentals, second ed. McGraw Hill, Singapore.

3. Boon, M., Ras, C., Heijnen, J. J., 1999. The ferrous iron oxidation kinetics of Thiobacillus ferrooxidans in batch cultures. Appl. Microbiol. Biotechnol., 51, 813-819.

4. Chandraprabha, M.N., Modak, J.M., Raichur, A.M. and Natarajan, K.A., 2001. Modeling of biooxidation of gold bearing pyrite-arsenopyrite concentrates by Acidithiobacilli ferrooxidans. In Ciminelli, V.S.T. and Garcia, O., (Eds.), Biohydrometallurgy: Fundamentals, Technology and Sustainable Development, Process Metallurgy 11A, Elsevier, Amsterdam, pp.43-54.

5. Jensen, A.B. and Webb, C., 1995. Ferrous sulphate oxidation using Thiobacillus ferrooxidans: A review. Process Biochem., 30, 225-236.

6. Jose, M.G., Domingo, C. and Barrie, J., 1997. Comparison of the effects of temperature and pH on iron oxidation and survival of Thiobacillus ferrooxidans (type strain) and Leptospirillum ferrooxidans. Biotechnol. Bioengg., 42, 581-589.

7. Karavaiko, G.I., 1988 Methods of isolation, evaluation and studying of microorganisms. In Karavaiko, G.I., Rossi, G., Agate, A.D., Groudev, S.N., Avakyan, Z.A., (Eds.), Biotechnology of Metals – Manual, Center for International Projects GKNT, Moscow, pp.47-79.

8. Kelly, D.P. and Jones, C.A., 1978. Factors affecting metabolism and ferrous iron oxidation in suspensions and batch cultures of Thiobacillus ferrooxidans. In: Murr, L.E., Torma, A.E. and Brierly, J.A., (Eds), Metallurgical applications of bacterial leaching and related microbiological phenomenon, Academic, New York, pp. 19-44.

9. Lizama, H.M. and Suzuki, I., 1989. Synergitic competitive inhibition of ferrous iron oxidation by Thiobacillus ferrooxidans by increasing concentrations of ferric iron and cells. Appl. Environ. Microbiol., 55, 2588-2591.

10. Modak, J.M. and Natarajan, K.A., 1995. Development of special strains of Thiobacillus ferrooxidans for enhanced bioleaching of sulphide minerals. In: Jerez, C.A., Vargas, T., Toledo, H. and Wiertz, J.V., (Eds), Biohydrometallurgical processing, Vol I, University of Chile, Santiago, Chile, pp. 34-46.

11. Nagpal, S. and Dahlstrom, D., 1992. Effect of carbon dioxide concentration on the bioleaching of a pyrite-arsenopyrite ore concentrate. Biotechnol. Bioengg., 41, 459-464.

12. Silverman, M.P. and Lundgren, D.G., 1959. Studies on chemoautotrophic iron bacterium Thiobacillus ferrooxidans. J. Bacteriol., 77, 642-646.

13. Taxiarchou, M., Adam, K. and Kontopoulos, A., 1994. Bacterial oxidation conditions for gold extraction from Olympias refractory arsenical pyrite concentrate. Hydrometallurgy 36, 169-185.



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Ferrous ion oxidation by an activated carbon cloth modified with Acidithiobacillus ferrooxidans

F. de J. Cerino-Cordova, J.P. Magnin, N. Gondrexon, P. Ozil

Laboratoire d’Electrochimie et de Physico-chimie des Matériaux et des Interfaces, UMR 5631 INPG-CNRS- UJF, Equipe Génie des Procédés, ENSEEG BP75,

38402 St Martin d’Hères, France

Abstract The formation of Acidithiobacillus ferrooxidans biofilm on an Activated Carbon

Cloth (ACC) was performed during subcultures of a biologically modified ACC in sterile ferrous iron medium. Two ways of modification were investigated: the classical culture of free bacteria on ACC, an enhanced way using a preliminary sorption of concentrated bacteria on ACC surface.

In the absence of bacteria, the ACC exhibited a low ferrous-iron oxidation capacity, 23 mg Fe2+ L-1/h. The presence of free (initial concentration 0.07 mg protein L-1) or immobilised bacteria on ACC WWP3 and WKL20 (4-8 mg protein/g carbon) increased the ferrous ion oxidation rate respectively up to 60-90 and 600 mg Fe2+ L-1/h.

The preliminary sorption step allowed to decrease the time required for biofilm stabilisation of about 30 percent, depending on the initial concentration of bacteria immobilised on ACC. No difference was observed concerning the biofilm formation and activity for the two types of ACC, excepted the final fragility of bacterial modified KL20 ACC.

Keywords: Acidithiobacillus ferrooxidans, biofilm, activated carbon cloth

1. INTRODUCTION Acidithiobacillus ferrooxidans is a chemoautotrophic acidophilic microorganism,

which gets the energy indispensable for its growth from oxidation of ions ferrous, elemental sulphur and reduced sulphur compounds including metal sulphides [1]. This ability has been used from a long time in different processes such as bioleaching [2-6], H2S removal [7] and acid mine drainage treatment [8-9].

It is well known that the free growth of A. ferrooxidans is characterised by a low rate, thus requiring long residence times and consequently large reactor volumes resulting in high investment and operating costs [10]. Therefore many research works have been focused on increasing ferrous iron oxidation rate using immobilised biomass. Various configurations were investigated using different bioreactor types (packed bed [9-13], fluidised bed [13-15] trickle bed reactor [16] and rotating biological contactor [17,18,19]) with different non-metabolisable supports (glass beads [13-14]; ion exchange particles


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[13-14], diatomeous earth [20-21], polyurethane foam [22], silica gel [23], sand [8], nickel alloy fibre [24], and activated carbon [13-14, 23]).

Activated carbon exists under various forms such as granule, fibre or cloth. Activated Carbon Clothes (ACC) offer faster kinetics than Granular Activated Carbon to treat micropolluants [26] and to sorb heavy metals [27].

Our previous studies were devoted to fixing A. ferrooxidans on Granular Activated Carbon in order to use this biological modified support as the biocathode of a bioelectroreactor [28]. Activated Carbon Cloth showed an improved biomass capacity for A. ferrooxidans immobilisation (respectively 25 mg protein /g Carbon for WWP3 and 14,9 mg protein /g Carbon for WKL20) [unpublished data]. Therefore the present paper focuses on developing a A. ferrooxidans biofilm on ACC and characterising it by its iron-oxidising activity and from Scanning Electron Microscopy observations.

2 MATERIALS AND METHODS

2.1 Culture and medium The bacterial strain Acidithiobacillis ferrooxidans DSM 583 was routinely maintained

in 9K liquid medium (33 g.L-1 FeSO4, 0.4 g L-1 MgSO4.7H2O, 0.4 g.L-1 K2HPO4, 0.4 g.L-1

(NH4)2SO4) at pH=1.4 (H2SO4) After bacterial growth at 30°C in a 5-L bioreactor, the biomass was concentrated by

cross-flow filtration before being harvested at 5000 g for 10 min. The resulting biomass was washed four times with H2SO4 (pH 1.4) and harvested at 1000 rpm during 10 min to remove jarosite precipitate. Finally, the biomass was suspended again in H2SO4 (pH 1.4) and kept at 5°C during 24 h before utilisation.

2.2 Activated Carbon Cloths The physical characteristics of ACC WWP3 and WKL20 (purchased from Actitex,

France) are recapitulated in table 1.

Table 1. Industrial characteristics of Activated Carbon Cloths WWP3 and WKL20

Characteristics WWP3 WKL20 Specific Surface (MEB) 1000-1300 m2/g 1000 m2/g Weight 110-130 g/m2 100 g/m2 Mechanical resistance 5.0 DaN/cm - Median Pore diameter 8.4 Å 7 Å Weaving Woven fabric Knitted fabric Pore < 20 Å 0.52 cc/g 0.31 cc/g Pore < 500 Å 0.65 cc/g 0.33 cc/g Microporosity >80% >90% Thickness 0.5 m 0.5 m

2.3 Analytical procedures The concentration ratio of ferrous ion to total (ferrous + ferric) ions was measured by

the o-phenantroline method [29]. The bacterial concentration in solution correlated with protein concentration was estimated by the Lowry method [30].


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2.4 Modification of ACC with bacterial biomass The ACC biological modification was performed via preliminary adsorption batch

experiments at 30°C and 100 rpm. A piece of ACC (0.1345 g) was introduced in a 0.1 L flask containing 0.02 L H2SO4 (pH=1.4) and 0.2 g protein L-1. After 2h, this piece was rinsed with sterile demineralised water to eliminate non-attached bacteria. It was next used in the ferrous ion oxidation experiments.

2.5 Bacterial subcultures in the presence of ACC cloths The A. ferrooxidans bacterial cultures with modified or unmodified ACC (WWP3 or

WKL20) were realised in 0.1 L flask containing 0.2 L of culture medium at 30°C under rotary shaker agitation (150 rpm).

In the case of non-biologically modified ACC, biomass (0.07 mg prot. L-1) was introduced in the medium. During incubation, samples were regularly collected and analysed to determine both ferrous ion oxidation activity (ratio Fe2+/ (Fe2+ + Fe3+)) and protein concentration in the medium. When 90% of ferrous ion concentration were oxidised (corresponding to a time noted T90), the ACC samples were picked up under sterile condition, rinsed with sterile demineralised and introduced into a 20-ml sterile culture medium. This procedure was repeated several times until reaching constant T90 and bacterial concentrations in solution.

2.6 Scanning Electronic Microscopic ACC samples were rinsed with sterile demineralised water and put into a Na-

cacodylate solution (4% cacodilyc acid, sodium salt trihydrate 98%) dissolved in a 25% glutaraldehyde solution (pH 7). After 4-h incubation, the samples were transferred firstly into a water solution for 1h and then successively into 25%, 50%, 75% and 100% xylene/acetone solutions for 10 minutes each. Finally, the samples were air-dried and prepared before SEM observation (Jeol 6400).

3. RESULTS AND DISCUSSION The WKL20 and WWP3 ACC under study exhibited a capacity to oxidise ferrous ion

in solution as shown in Fig. 1. Without ACC and in an uninoculated medium, no ferrous iron was oxidised. In the absence of bacteria, 90% of ferrous ions were thus oxidised after 170 and 340 h respectively with WWP3 and WKL20. This Fe2+ oxidation capacity of ferrous ion oxidation was previously observed on activated carbon [11] and carbon fibre [31].

Figure 1. Ferrous oxidation by Activated Carbon Cloth WWP3 or WKL20 under different conditions: with and without bacteria


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The presence of bacteria in solution did not change the Fe2+ oxidation rate during the first 40h for of WKL20 and during 70 h for WWP3. Ferrous ions were rapidly oxidised either by ACC samples than by A. ferrooxidans biomass. After this period, bacterial Fe2+ oxidation was more important than that observed with ACC alone. 90% of ferrous ions were thus biologically oxidised within a short time (respectively 50 h and 90 h in the presence of WKL20 and WWP3 ACC). This oxidation activity was directly correlated to the free or immobilised biomass production. The protein concentrations (5 mg.L-1 with of WKL20 and 3 mg.L-1 with WWP3) were lower than that obtained without ACC (7 mg.L-

1). As a conclusion ACC supports present an inhibition effect for iron-oxidising activity such as observed for activated carbon [23], flowers of sulphur, fluorapatite, glass bead and pyrite [5,32]

This bacterial diminution in solution was linked to the development of a biofilm onto the ACC surface associated with a jarosite precipitation (Fig. 2). No jarosite or biofilm structures were observed onto initial ACC samples (data not shown here). A. ferrooxidans modified ACC materials show a catalytic effect in ferrous ion oxidation as observed in the case of Biomass Support Particles with Activated Carbon coating [11].

Figure. 2. SEM observations of A. ferrooxidans biofilm obtained (a) onto WKL20 after 60h, (b) WWP3 after 90h

The development of a stable biofilm onto ACC was studied during subcultures of A. ferrooxidans modified ACC in Fe2+ growth medium (see Materials and Methods) (Fig. 3). The amount of biomass present within the biofilm was indirectly measured from the amount of biomass present in solution (via protein concentration) and associated to SEM observation of the biofilm. The total Fe2+ oxidation activity was estimated by determining the time required to oxidise 90% of ferrous ions (T90), resulting to the biological activity due to the biomass eventually present in solution or immobilised into the biofilm.

The evolution of protein concentration or T90 appeared to be similar for both biologically modified ACC. The ferrous activity curves involved three distinct steps. During step I, the bacterial concentration in solution after oxidation of 90% Fe2+ was greater than that obtained during a batch culture without AC Cloths (5 mg.L-1). These freely dispersed bacteria resulted from two actions: either cells were divided in the biofilm, left the biofilm, and were suspended in

medium, or / and the loss of some bacteria from the biofilm due to friction between the ACC

support and glass flask. [15] In both cases, Fe2+ oxidation rate was increased as demonstrated by the fast T90

decreasing from 85h to 10-18 h depending on the ACC biofilm. The biofilm was more developed that that obtained after the first culture and consisted of jarosite precipitates and


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bacteria (Fig 4). Bacterial cells become embedded within iron precipitates and were adsorbed through complex multilayers patterns.

Figure 3. Biological characteristics (protein concentration and time necessary to oxidise 90% of ferrous ions, T90) of subcultures of ACC WKL20 (A) and WWP3 (B) in 9K medium. A. ferrooxidans biomass (0.07 mg.L-1) being only introduced during the first subculture). The vertical arrows with dotted lines indicate the different steps in the biofilm formation (see text)

During step II, T90 was increased from 10-18 h to 40 h, thus indicating a decrease of the overall activity for ferrous ions oxidation. Such diminution of biological activity was correlated to a large bacterial concentration in solution (6-11 mg protein L-1). The biomass present in solution exhibited a very low or a non-existent ferrous ion oxidation activity. However more jarosite deposits and bacteria were observed onto biofilm (Fig. 4).

Step III led to the formation of a biofilm which is stable for its structure and its metabolic activity: no variation in the biomass concentration present in solution and in the time necessary to oxidise 90% of ferrous ions initially present in solution were observed. A weak bacterial concentration in solution (1-2 mg protein L-1) was reached after 13 subcultures. The final biofilm recovered the whole ACC surface and involved a significant jarosite deposit (Fig. 4).

Finally, a stable and active A. ferrooxidans biofilm was obtained onto both ACC under study after about 300-350 h subcultures. This biofilm oxidised five times faster the substrate (ferrous ions) than a culture with free biomass.

In order to decrease the time required to obtain a stable biofilm onto ACC, the clothes used latter were previously modified by A. ferrooxidans biomass. This modification resulted of an initial sorption of concentrated biomass onto ACC (see Materials and Methods). A 2 h contact period allowed to reach the equilibrium between non immobilised


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and immobilised cells. A better bacterial sorption capacity was obtained with WWP3 (7.8 mg protein L–1) than with WKL20 (4.6 mg protein L–1) as confirmed by SEM observations (Fig. 5).

Figure 4. SEM observations of A. ferrooxidans biofilm onto Activated Carbon Cloths WKL20 and WWP3 during subcultures in 9K medium (arrows =bacterial cell)

Figure 5. SEM observations of Activated Carbon Cloths modified with bacteria after 2h of bacterial fixation (pH 1.4, 30°C)

The development of A. ferrooxidans biofilm from these initially fixed biomasses was studied during subcultures of biologically modified ACC in 9K medium (Fig. 6A, B).

No shape difference was observed on the evolution of biological characteristics whatever the ACC nature may be. However these curves were simpler than those observed when the cloth is directly introduced into the bacterial culture (Fig. 3): only two steps were observed instead of three.


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During the first culture, the ferrous ions were rapidly oxidised by bacteria as shown by the weak T90 value (12 h). This oxidation was correlated to a desorption process confirmed by measuring protein concentration in solution and by comparing SEM observations (Fig. 7 and Fig. 4 for WWP3).

Figure 6. Biological characteristics (protein concentration and T90) of subcultures for ACC WKL20 (A) and WWP3 (B) modified with A. ferrooxidans biomass in 9K medium. The vertical arrows with dotted lines indicate the different steps in the biofilm formation (see text)

Protein concentration in solution differed according to the AA cloth nature: 12 mg.L-1 for WKL20 and 4 mg.L-1 for WWP3. During the five first subcultures, the protein concentration in solution increased because of bacterial desorption. This desorption was correlated to an increase of the time required to biologically oxidise ferrous ions: 12h for WKL20 and 16 h for WWP3. However, the biofilm was developing as shown by SEM observations on Fig. 7.

From the sixth subculture, protein concentration in solution continuously decreased down to 1-2 mg protein L-1. T90 simultaneously decreased indicating a high ferrous ion oxidising capacity. This capacity was mainly due to biofilm because of the very low bacteria concentration in solution.


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After 232-265h of total subculture, the biologically active biofilm oxidised the ferrous ion substrate within 10-12h, with a very little biomass desorption (1-2 mg protein L-1).

Figure 7. SEM observations of the formation of an A. ferrooxidans biofilm on Activated Carbon Cloth WKL20 and WWP3

4. CONCLUSION Activated Carbon Cloth has been selected as a supporting material because of its large

specific volume and high ability to adsorb A. ferrooxidans. The ACC here under study showed their capacity to oxidise ferrous ions. However, this capacity is improved by developing a A. ferrooxidans biofilm.

As observed by SEM, the biofilm consisted in jarosite deposits wherein bacterial cells are able to develop. The development of a stable iron-oxidising biofilm may be explained by the model of "biofilm with adsorbed cells" previously described by Karamanev [15]. A very high iron-oxidising activity (600 mg Fe2+ oxidised L-1/h) was obtained for biologically modified ACC. A preliminary bacterial immobilisation onto ACC allowed an enhancement in developing and stabilising the biofilm. The formation of a stable biofilm and long time incubation in acidic solution led to a loss of mechanical resistance. Therefore ACC WKL20 rapidly breaks in contrast to ACC WWP3.

ACKNOWLEDGEMENTS This work was carried out owing to a grant from "Consejo Nacional de Ciencia y

Tecnologia" (CONACYT, Mexico) accorded to Mr Felipe Jesus Cerino Cordova.


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Eng. J., 12 (2002) 43. 18. H. Olen and R.F. Unz, J. Water Pollut. Control Fed, 52 (1980) 257. 19. H. Olen and R.F. Unz, Biotechnol. Bioeng, 19 (1977) 1475. 20. T. Imaizumi, Biotechn. Bioeng Symp., 16 (1986) 363. 21. H. Magota, Y. Shiratori, C. Inoue, E. Yanagisawa, JP patent No.01215399 (1989). 22. H. Armentia and C. Webb, Appl. Microbiol. Biotechnol., 36 (1992) 697. 23. T. Kai, T. Takahashi, Y. Shirakawa, Y. Kawabata, Biotechnol. Bioeng., 36 (1990)

1105. 24. S. Sandhya and R.A. Pandey, J. Environ. Sci. Health, 2 (1992) 445. 25. J.M. Gomez, D. Cantero, C. Webb, Appl. Microbiol. Biotechnol., 54 (2000) 335 26. C. Brasquet, E. Subrenat, P. Le Cloirec, Wat. Sci. Tech. 39 (1999) 201. 27. K. Kadirvelu, C. Faur-Brasquet, P. Le Cloirec, Langmuir, 16 (2000) 8404. 28. A. Boyer, J.P. Magnin, P. Ozil, Int. Biohyd. Symp., Madrid, Spain, (1999) 211. 29. O.H. Tuoniven and D. P. Kelly, Arch. Microbiol., 88 (1973). 285. 30. M.K. Muir and T.N. Anderson, Metall. Trans. B., 88 (1977) 517. 31. M. Uchida, O. Shinohara, S. Ito, N. Kawasaki, T. Nakamura, S. Tanada, J. Colloid

Interf. Sci., 224 (2000) 347. 32. A.A. Dispirito, P.R. Dungan, O.H. Tuoniven, Biotechnol. Bioeng., 23 (1981) 2761.



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Heavy metal precipitation by off-gases from aerobic culture of Klebsiella pneumoniae M426

A.M.M. Essa, L.E. Macaskie∗ and N.L. Brown

School of Biosciences, The University of Birmingham, Edgbaston, Birmingham B15 2TT UK

Abstract In this study we have identified a new method for the removal of heavy metals from

solution by uses the off-gas produced from aerobic cultures of Klebsiella pneumoniae M426. GC-Mass Spectroscopic analysis showed that the outlet gases contain dimethyldisulfide (DMDS), which precipitates some metals from their solutions. The outlet gases precipitated mercury, cadmium, palladium, and lead as metallo-thiol complexes but did not precipitate other metals such as nickel, cobalt, zinc and arsenite. Energy dispersive X-ray microanalysis showed that the metal precipitate consisted of the metal plus carbon, sulfur and oxygen. Formation of metal carbonates was discounted since precipitation was similar at pH 2.0; the precipitate did not have the physical characteristics of metal sulfides. Bio-mineralization of heavy metals by using DMDS is a novel approach because it can be used with concentrated or physiologically incompatible solutions, and since the metal precipitate is kept separate from the bacterial biomass, the biological and chemical end products can be managed independently.

1. INTRODUCTION The discharge of heavy metals into the environment due to agricultural, industrial and

military operations and the effect of this pollution on the ecosystem and human health are growing concerns. Since the natural mineralization of metals is a slow process, pollution by metals constitutes one of the most important environmental problems. Some metals seem to serve no biologically relevant function e.g. Cd, Pb, Sn, Hg instead they cause damage due to their avidity for the sulfhydryl groups of proteins which they block and inactivate [1-6]. Different procedures for the removal of toxic metal species from contaminated environments have been developed; most of them are based on ion-exchange technologies and/or chemical precipitation of cations in an inert form. These methods are expensive and also require additional products for desorption of metals and for cleaning up of the inorganic matrix and lead to formation of concentrated secondary wastes. Remediation technologies using microorganisms are feasible alternatives to the physical cleaning of soils or the concentration of metals in polluted waters by physical or chemical means [7]. For example, sulphate-reducing bacteria have been used to precipitate metals as metal sulfides [8] but this requires control of emission of highly toxic H2S. As

∗ To whom correspondence should be addressed (Fax: +44(0)121-414-5925; e-mail: [email protected])


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an alternative S-based precipitation process organosulfur compounds have been considered, but these are too expensive for routine use [9].

Dimethyldisulfide (DMDS) has been reported to be produced from different organisms: Pseudomonas sp. [10], Streptomyces sp. [11], Aspergillus sp. [12], Clostridium sporogenes [13], Alteromonas putrefaciens, Pseudomonas fluorescens and Achromobacter sp. [14], Pseudomonas putida [15], Proteus sp. [16] and Pseudomonas fluorescens, Proteus vulgaris, and Serratia marcescens [17]. These bacteria were isolated from soil, wastewater, and fish. Some bacterial strains isolated from activated sludge [18] as Lactobacillus sp., Corynebacterium sp., Alcaligenes sp., and Pseudomonas sp. showed an ability to produce DMDS from activated sludge.

The formation of methanethiol from methionine by cell extracts of Lactococcus lactis subsp. cremoris B78 was proposed by Engels et al. [19]. The first step is the transamination during which 4-methylthio-2-ketobutyric acid (KMBA) is produced. The intermediate KMBA is probably, converted to methanethiol, which under aerobic conditions is converted rapidly to dimethyldisulfide and/or dimethyltrisulfide. In this study we describe a novel approach for the removal of some heavy metals by using the outlet gases produced during the aerobic culture of Klebsiella pneumoniae M426. These gases contain a volatile sulfur compound dimethyldisulfide (DMDS), which is proposed to precipitate heavy metals with high efficiency.


2.1 Bacterial strains and bioreactor operation Klebsiella pneumoniae M426 strain was originally obtained from the NCTC (National

Culture Type Collection) Colindale, London. Bacterial cells were grown in Luria Broth medium (LB; 20) in a shaker agitated at 200 rpm and incubated at 37°C. A batch bioreactor was constructed for the metal precipitation (Fig. 1) consisting of two chambers, one for the bacterial growth (about 1L in volume, maintained under aerobic conditions by pumping in a sterilised air, 37°C) and the other chamber (50ml) for the precipitation of metal via the pumping of the headspace gases from the culture through the metal solution. Cultures were inoculated (100 ml) using cells in an exponential growth phase (6hrs) and bacterial growth was monitored by measuring the optical density at 600nm.

2.2 Analytical methods

2.2.1 Analysis of the soluble metals Heavy metals (Hg2+, Cd2+, Pb2+, Pd2+) were prepared as aqueous solutions, and used

as the following concentrations: mercuric chloride, 0.02-10.0 mg/ml; cadmium chloride, 0.7 mg/ml; lead nitrate, 0.6 mg/ml; sodium tetrachloropalladate, 0.9 mg/ml. The metal precipitates from the precipitation chamber were removed by centrifugation at 1600g for 15 minutes. Aliquots (10 ml) of the supernatant were filtered through 0.2 µm (Nalgene Syringe filters). The filtrates (Cd2+, Pb2+, Pd2+) were analysed using a 1999 Duo HR Iris advanced inductive coupled plasma (ICP) spectrophotometer [21]. Mercuric ion concentrations in the solutions were measured with a mercury analyser (Buck Scientific, model 400A) applying cold vapour atomic absorption spectrometry [22].


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Figure 1. A bioreactor used for metal precipitation via the culture off-gas consisting of two chambers one is the bacterial growth chamber and the second is the metal precipitation chamber

2.2.2 Analytical electron microscopy Samples from the precipitation chamber were centrifuged at 1600g for 15 minutes.

The supernatant was removed and precipitates were washed in 20ml deionised water followed by centrifugation as before. This step was repeated 3 times. The precipitate was collected and dried at 30°C. The metal precipitates were examined with a JEOL JSM 5900 SEM. The elemental composition of the mineral phases was determined by energy dispersive X-ray microanalysis (EDX) using an Oxford Link ISIS System.

2.2.3 GC–mass spectral analysis The outlet gases produced from the bioreactor were collected in serum bottles (50ml)

with a rubber caps. The samples were immediately analysed by GC–MS [23], using a Hewlett-Packard 5890/5970A system, with a HP1 column (50 mm x 0.2 mm fused silica capillary column, film thickness, 0.5 mm). GC oven initial temperature was 60°C and was programmed to 220°C at a rate of 2°C/min, and finally held at 220°C during 40 min. Operating conditions of the GC were as follows: helium was used as carrier gas (0.8 ml/min); the temperature of injector and detector was 250°C. The mass spectra were determined at 70 eV in the mass range of 35-400.


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3.1 Precipitation of mercury as HgS by growing cultures During the aerobic growth of K. pneumoniae M426 a black precipitate was visible in

the extracellular matrix (Fig. 2a). EDX analysis data showed that the precipitate contained mercury and sulfur (Fig. 2b). Accordingly, it was assumed that K. pneumoniae M426 has the ability to produce hydrogen sulfide (H2S) under aerobic conditions, which can precipitate Hg2+ as insoluble HgS [24]. These results were confirmed when a suspended filter paper saturated with HgCl2 turned black after exposure to the culture outlet gases. This phenomenon was recognised previously by Aiking et al. [25], who recorded an aerobic mechanism for mercury bioremediation, in which Klebsiella aerogenes formed HgS when grown in continuous aerobic culture in the presence of HgCl2 and at the same time there was no bacterial mercury volatilization. This strain showed mercuric ion sensitivity under sulfate-limited conditions and elevated levels of total cellular sulfide were detected in cells grown in the presence of mercuric ions. It was suggested that this was due to the formation of HgS. A similar phenomenon was recorded by Wang et al. [26] who showed the ability of Pseudomonas aeruginosa to remove more than 99% of the cadmium from 5mM cadmium solution by precipitation on the cell wall as cadmium sulfide under aerobic conditions.

Figure 2. (a) Electron micrograph of aerobic culture of K. pneumoniae M426 showing an extracellular dark precipitate, (b) EDX spectrum of the dark precipitate in (a)

3.2 Mercury and cadmium bio-mineralization using culture off-gases A bioreactor for mercury bioremediation was designed according to the ability of K.

pneumoniae M426 to precipitate the soluble Hg2+ as HgS. The bacterial cells were used as a gas generator in a separate chamber and the headspace gases produced during aerobic growth were pumped into the metal solutions in the precipitation chamber through a bacterial membrane filter to avoid any bacterial contamination.

A black HgS precipitate was anticipated, but more than 99% of the soluble Hg2+ was recovered as a yellowish white precipitate within 2h (Fig. 3). A similar result was obtained


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with cadmium; within 1h there was more than 99% of Cd2+ removal from CdCl2 solution. EDX analysis showed that the precipitates consisted of mercury or cadmium plus carbon, sulfur and oxygen and there was no phosphorus in the precipitate, which confirms that there was no bacterial contamination (Fig. 4a & 4b, 5a & 5b).

Bacterial growth

(O.D600)

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0.1

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Hg2+ (mg/l)

Figure 3. A time course for bacterial growth and mercury bio-precipitation by using the outlet gases produced from the aerobic culture of K. pneumoniae M426. The data shown are for an initial Hg2+ concentration of 0.8 mg/ml, similar results were obtained using Hg2+ up to 10.0 mg/ml (Error bars represent means ± standard errors)

Figure 4. Mercury precipitation by using the volatile thiols produced by K. pneumoniae M426, (a) an electron micrograph of the mercury precipitate, (b) EDX spectrum of the mercury precipitate


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Figure 5. Cadmium precipitation by using the volatile thiols produced by K. pneumoniae M426, (a) an electron micrograph of the cadmium precipitate, (b) EDX spectrum of the cadmium precipitate

The EDX spectrum for the Hg precipitate cannot easily distinguish between Hg and sulfur due to peak overlap, but the peaks were clearly differentiated in the corresponding Cd-precipitate. It could be suggested that metal carbonates were precipitated. However similar results were obtained at pH 2.0, at which pH little HCO3

- would be present in the solution and no CO3

2- [27]. GC-Mass Spectroscopic analysis showed that the outlet gases from the bioreactor contained volatile dimethyldisulfide (DMDS). According to these results, we suggest a novel mechanism for mercury bioremediation by the production of DMDS as a metabolic by-product during the aerobic growth of K. pneumoniae M426.

This volatile species may form insoluble mercurithiol-compounds. There was more than 99% mercury precipitation by this mechanism with different concentrations of HgCl2 (from 20 µg/ml to 10 mg/ml) within 2 hrs. This bioprocess also gave high metal removal efficiency over a wide pH range (acidic, neutral or alkaline) and also at salinity levels up to 1M NaCl.

DMDS for bio-mineralization of heavy metals The head gas produced by aerobic culture of K. pneumoniae M426 did not precipitate

some metals e.g. nickel, cobalt, zinc and arsenite but showed a high efficiency of precipitation of cadmium, lead and palladium from solution with about 99% metal removal efficiency within 1h. EDX microanalysis showed that the metal precipitates consisted of metal plus carbon, sulfur and oxygen, while GC-Mass Spectrum analysis showed the presence of dimethyldisulfide (DMDS) in the culture off-gas. We attributed the precipitation to the high affinity between DMDS and some metal ions to produce metallo-thiol complexes. The formation of metal sulfides seems unlikely since mercury and cadmium deposits were white instead of black and yellow respectively and the characteristic X-ray powder diffraction spectra of metal sulfides could not be obtained (data not shown). There are some reports of the use of organosulfur compounds [2,3-Dimercaptopropanol, meso-2,3-Dimercaptosuccinic acid (DMSA), Sodium 2,3-dimercapto-1-propanesulfonate (DMPS)] as therapeutic chelating agents for the removal of some heavy metals (arsenic, cadmium, lead and mercury) from living cells. The formation of complexes with metal ions can be attributed to the presence of sulfhydryl group in these chelators, which show a great affinity for metal ions producing the metal


1125

complexes [28-35]. According to our knowledge this is the first study using a naturally produced organo-volatile sulfur gases as a chelating agent to precipitate heavy metals with such high metal removal efficiency under aerobic conditions. Current studies are aiming to quantify DMDS production and its contribution to metal precipitation.

4. CONCLUSIONS There have been attempts to engineer aerobic sulfate or thiosulfate reducing systems

for metal removal to produce more sulfide [36, 37]. This study describes a novel natural process not only for mercury bioremediation but also for some other heavy metals e.g. cadmium, lead and palladium via aerobic bio-precipitation of metal ions as a metal-thiol complex. This mechanism provides an excellent approach for bioremediation of some heavy metal and has some advantages. Firstly, since culture off-gas is used, it is applicable for concentrated or non-physiologically compatible solutions under a wide range of environmental factors such as pH and salinity. Secondly, with this bioprocess one organism, K. pneumoniae, was used to precipitate different metals from their solutions and could be used with mixed metal solution. Thirdly, and an important advantage, during metal bio-mineralization there is no direct contact between the bacterial biomass and the metal precipitate; thus keeping the biomass away from the original contaminated waste water, ensuring that there is no cross-contamination between the biomass and the final metal precipitate.

REFERENCES 1. G. M. Gadd, Accumulation of metals by microorganisms and algae. In: Biotechnology,

Ed. H. J. Rehm. VCHV, Weinheim, 1988. 2. G. M. Gadd, Microbial control of heavy metal pollution. In: Microbial Control of

Pollution, Eds J. C. Fry, G. M. Gadd, R. A. Herbert, C. W. Jones and I. Watson-Craik. Cambridge University Press, Cambridge, 1992.

3. G. M. Gadd, FEMS Microbiol. Rev. 11 (1993) 297. 4. M. D.Mullen, D. C. Wolf, F. G. Ferris, T. J. Beveridge, C. A. Flemming and G. W.

Bailey, Appl. Environ. Microbiol. 55 (1989) 3143. 5. S. G.Walker, C. A. Flemming, F. G. Ferris, T. J. Beveridge and G. W. Bailey, Appl.

Microbiol. Biotechnol. 55 (1989) 2967. 6. T. J. Beveridge, Rev. Microbiol. 43 (1989) 147. 7. M. M. Valls and V. de Lorenzo, FEMS Microbiol. Rev. 26 (2002) 327. 8. G. M. Gadd and C. White, Trends Biotechnol. 11 (1993) 353. 9. R. M. Austrain, Ph.D Thesis, Institut für Thermische Verfahrenstechnik und

Umwelttechnik, Germany, 1994. 10. R. E. Kallio and A. D. Larson, Methionine degradation by a species of Pseudomonas.

In W. D. McElroy and H. B. Glass (ed.), A symposium on amino acid metabolism. Johns Hopkins University Press, Baltimore, 1955.

11. W. Segal and R. L. Starkey, J. Bacteriol. 98 (1969) 908. 12. J. Ruiz-Herrera, and R. L. Starkey, J. Bacteriol. 99 (1969) 544. 13. W. Kreis and C. Hession, Cancer Res. 33 (1973) 1862. 14. A. Miller, R. A. Scanlan, J. S. Lee and L. M. Libbey, Appl. Microbiol. 26 (1971)18. 15. S. Ito, T. Nakamura and Y. Eguchi, J. Biochem. 79 (1976) 1263 16. N. J. Hayward, T. H. Jeavons, A. J. C. Nicholson and A. G. Thornton, J. Clin.

Microbiol. 6 (1977) 187. 17. M. Pohl, E. Bock, M. Rinken, M. Avvdin and W. A. Konig. Z, Naturforsch. Teil C. 39

(1984) 240.


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18. B. Tomita, N. Hmamura and Y. Ose. Eisei Kagaku, 27 (1981) 98. 19. W. J. M. Engels, A. C. Alting, M. M. T. G. Arntz, H. Gruppen, A. G. J.Voragen, G.

Smit and S. Visser, Int. Dairy J. 10 (2000) 443. 20. J. Sambrook, E. F. Fritsch and T. Maniatis, Molecular Cloning: a Laboratory Manual,

2nd ed., vol. 3. Cold Spring Harbor Laboratory Press, Cold Spring Harbor, N.Y. 1989. 21. M. M. Matlock, B. S. Howerton and D. A. Atwood. J. Hazardous Materials 84 (2001)

73. 22. J. S. Chang, J. Hong, Q. A. Ogunseitan and B. H. Olson, Biotechnol. Prog. 9 (1993)

526. 23. M. Mestres, O. Busto and J. Guasch, J. Chromatography 945 (2002) 211. 24. A. M. M. Essa, L. E. Macaskie and N. L. Brown, Biochem. Soc. Trans. 4 (2002) 672. 25. H. Aiking, H. Govers and J. 't Riet, Appl. Environ. Microbiol. 50 (1985) 1262. 26. C. L. Wang, P. C. Michels, S. C. Dawson, S. Kitisakkul, J. A. Bross, J. D. Keasling

and D. S. Clark, Appl. Environ. Microbiol. 63 (1997) 4075. 27. C. N. Mulligan, R. N. Yong and B. F. Gibbs, J. Hazardous Materials, 85 (2001) 111. 28. H. V. Aposhian and M. M. Aposhian, Ann. Rev. Pharmacol. Toxicol. 30 (1990) 279. 29. H. V. Aposhian, D. C. Bruce, W. Alter, R. C. Dart, K. M. Hurbult and M. M.

Aposhian, FASEB J. 6 (1992) 2472. 30. M. M. Aposhian, R. M. Maiorino, Z. Xu and H. V. Aposhian, Toxicol. 109 (1996) 49. 31. P. Asiedu, T. Moulton, C. B. Blum, E. Roldan, N. J. Lolacono and J. H. Graziano,

Environ. Health Perspect. 103 (1995) 734. 32. J. Chisolm, J. Clin. Toxicol. 30 (1992) 493. 33. J. H. Graziano, N. J. Lolacono, T. Moulton, M. E. Mitchell, V. Slavkovich and C.

Zarate, J. Pediatr. 120 (1992) 133. 34. M. M. Jones, Crit. Rev. Toxicol. 21 (1991) 209. 35. R. A. Peters, L. A. Stocken and R. H. S. Thomson, Nature. 156 (1945) 616. 36. S. W. Bang, D. S. Clark and J. D. Keasling, Appl. Environ. Microbiol. 66 (2000) 3939. 37. C. L.Wang, P. D. Maratukulam, A. M. Lum, D. S. Clark and J. D. Keasling, Appl.

Environ. Microbiol. 66 (2000) 4497.



1127

Influence of pH, Mg2+ and Mn2+ on the bioleaching of nickel laterite ore using the fungus Aspergillus niger O5

O. Cotoa, D. Gutierreza, L. Abína J. Marreroa and K. Boseckerb a Departamento de Microbiología. Facultad de Biología. Universidad de la Habana.

Calle 25 entre J e I. Plaza. Vedado. Ciudad Habana. Cuba b Federal Institute for Geosciences and Natural Resources (BGR)

Stilleweg 2, D 30655 Hannover, Germany

Abstract The main global nickel deposits are laterite deposits. Their nickel contents vary from

0.3-1.5% and are associated with a non-soluble phase that cannot be processed completely by conventional methods. The Cuban nickel deposits are recognized as one of the largest deposits in the world. Besides nickel they contain other valuable metals, such as cobalt and manganese, which in spite of low concentrations of valuable metals, are of important commercial interest.

The aim of this investigation was to study the influence of pH and Mg2+ and Mn2+ ion concentrations on the dissolution of Ni, Co, Mn, and Mg from serpentinite in the Punta Gorda deposit in Moa (Cuba). For leaching, Aspergillus niger O5 was chosen because of its ability to dissolve a high percentage of Co, Ni and Mn from the mineral substrate and for its production of citric acid. The strain produced higher concentration of citric acid (25-27 g/L) when grown in the presence of the ore than in the absence of ore (14 g/L). The production of organic acid by the fungus varied with the conditions of incubation. The initial pH of the medium was a very important variable in this process; its change during the course of the time affected the bioleaching efficiency and gave rise to biosorption processes.

Keywords: bioleaching, heterotrophic microorganisms, Aspergillus niger, Penicillium simplicissimum, citric acid, organic acids, laterite ore, oxide ore

1. INTRODUCTION Cuba has an estimated world reserve of Ni of 37% and ranks first in the world

reserves of Ni and second in world reserves of Co. The main nickel deposits in Cuba are on the east side of the island in the Moa-Nicaro area of Holguín province. In this area there are three commercial plants where Ni+Co oxides and Ni+Co sulfides are obtained by hydrometallurgy and pyrometallurgy, respectively. Both technologies require high temperature, special equipment and constitute a high environmental risk. For this reason, bioleaching with organic acid produced by fungi (1-9) might be an attractive alternative for the Cuban nickel industry in order to develop a sustainable technology.

The number of scientific papers and patents on bioleaching of sulfide ores is quite large and exceeds several times the number of papers related to bioleaching of non-sulfide


1128

ores. Bioleaching of non-sulfide ores is based on the ability of fungi to produce organic acids such as citric, oxalic, and fumaric acid, which act as leaching agents (1–9). They are excreted into the environment and dissolve heavy metals by direct displacement of metal ions from the ore matrix by hydrogen ions and by the formation of soluble metal complexes and chelates (2-3). Fungi most active in leaching are Penicillium simplicissimum and Aspergillus niger. Studies of nickel leaching have revealed that the best recovery has been obtained with citric acid (2, 3, 6, 8).

The main objective of this paper was to study the influence of pH and Mg2+ and Mn2+ ion concentrations on the dissolution of Ni, Co, Mn and Mg present in the serpentinite from the Punta Gorda deposit at Moa, Cuba.


2.1 Laterite ore Nickeliferous ore from the Punta Gorda, Moa (Cuba) deposit was used in all

experiments.

Table 1. Chemical and mineralogical composition of Cuban laterite ore

Element Ni Co Fe Mg Mn Al Cr

Concentration (%) 1.54 0.27 14.8 19.86 0.49 1.95 1.02 Phases Al2O3 Fe2O3 MnO MgO SiO2

% 1.64 17.53 0.253 29.56 35.87

2.2 Screening of microorganisms for bioleaching Sixteen strains of fungi were used in this work, 15 of them identified as Penicillium

simplicissimum (strains 2, 3, 14, 16, 22, 23, 25, 26, 42, 43, 47, 50, 56, 60, 61). They were obtained from the Federal Institute for Geosciences and Natural Resources, Hannover, Germany. The other fungus was Aspergillus niger strain O5 (6). All strains were grown on solid Saboraud medium and incubated for one week at 30°C. Culture medium 1a (100 mL) was inoculated with 2 mL containing 107-108 spores/mL in a 500 mL Erlenmeyer flask. Culture medium 1a was prepared as follows: 140 g/L sucrose; 0.3 g/L yeast extract; 0.5 g/L KH2PO4; 0.250 g/L MgSO4 and 0.125 g/L MnSO4. The pH was adjusted to 6.5. The spores were counted using a Thoma counting chamber.

All leaching experiments were carried out at 5% pulp density of laterite ore. The ore was milled to a mean particle size of 0.250 mm. The culture medium was incubated at 100 rpm at 30°C. Each time a sample was taken from the solution the same volume was replaced with sterile culture medium. One Erlenmeyer flask without inoculum was used as a control.

2.3 Culture conditions for bioleaching To study the effect of oxygen on bioleaching, one Erlenmeyer flask was placed on a

shaker for the duration of the experiment, a second one was not shaken. The influence of the culture medium was investigated using seven variants of the

growth medium. Culture medium 3a was used only for the Ni leaching experiment.


1129

Table 2. Variants of culture media

Culture media (with 140 g/L sucrose and 0.3 g/L yeast extract) 1 1a 2 2a 3 3a 3b

Mg2+ - + - - + + + Mn2+ - + - - + + + pH 6.5 6.5 5.5 3.0 5.5 5.0 3.0

2.4 Bioleaching in two-stages In the first step, culture medium 3a (100 mL in a 500 mL Erlenmeyer flask) was

inoculated with A. niger O5 and incubated at 150 rpm/min for 8 days at 30°C. In the second step, the biomass was removed by filtration; the filtrate was autoclaved and used to leach Ni ore with a 5% pulp density.

2.5 Analytical methods At regular intervals, samples were withdrawn for pH measurement and element

analysis by ICP-AES. The type and concentration of the organic acids produced were determined using a densitometer after separation by thin layer chromatography (TLC). The spots became apparent under ultraviolet light (366 nm) and after staining at visible light.

The mineralogical composition of the Cuban laterite ore and some solid residues from bioleaching was determined by X-ray diffraction (XRD). This was done using a Philips X'pert system PW1730 equipped with a Cu tube, primary and secondary Soller slits and a secondary monochromator. The main phases are chrysotile, magnesioferrite and goethite.


3.1 Screening for microbial leaching In a preliminary experiment, we tested the ability of 16 strains of fungi to leach nickel

and other metals from Cuban laterite ore. Most of of the strains were able to dissolve Ni, Co, Mn, Al, Fe, and Cr to some extent, whereas strains 42, 56, 60 and 61 dissolved less than 10% (data not shown).

The amount of metal leached by four strains of the 16 strains studied is shown in Figure 1 as a function of time. The trend is rather similar with all four strains.

In the majority of strains, the initial pH of 6.2 decreased, indicating production of organic acid by the fungus, and the most of the change in pH occurred within the first seven days.

We took Ni solubilization as the main criterion for screening programs, because Ni is the metal of the most economic importance in Cuban laterite ore. Table 3 shows the percentage of nickel recovered after one and two weeks of bioleaching, indicating that the strains can be separated into three groups: the first group with the strain 23 and O5, recovered up to 60% Ni within 14 days. Strains 2, 3, 14, 22, 25, 26 and 50 form the second group, dissolving more than 20% and less than 50% Ni. The fungi of the third group, including strains 16, 42, 43, 47, 56, 60 and 61, extracted less than 15% Ni. The best strain was A. niger O5.


1130

Figure 1. Laterite bioleaching in medium 1a: A) strain 50, B) strain 3, C) strain 2 and D) strain 42. Pulp density: 5%; carbon source: sucrose 140 g/L; temperature: 30°C; reciprocal agitation: 100 rpm.min-1. (•: Mn, : Mg, ∆: pH, : Co, : Ni, : Cr, ♦: Al and ◊: Fe)

Table 3. Solubilization of Ni laterite ore by different strains % recovery after Fungal strain

7 days 14 days P. simplicissimum

2 0.27 34.30 3 14.73 25.89

14 12.64 21.02 16 9.56 14.91 22 15.61 24.42 23 30.15 56.23 25 16.10 24.92 26 21.51 35.43 42 4.93 6.57 43 6.92 11.18 47 11.42 13.72 50 17.52 31.71 56 1.60 2.09 60 10.60 12.35 61 9.36 10.01

A. niger O5 34.15 57.8


1131

The selection of the best strain is a very important factor in the bioleaching of Ni laterite ore and the leaching efficiency depends on the amount of acid produced by the fungus (1, 2, 6, 8).

3.2 Effect of incubation conditions on bioleaching of metals The results of bioleaching and change of pH with strain O5 under shaking and static

conditions are illustrated in Figure 2. Under both conditions the pH decreased to 3.8 within 7 days and remained below 4.0 thereafter.

Manganese was leached best and within 10 days about 70% Mn was dissolved. However, the pattern of bioleaching changed in the course of time depending on the incubation conditions. Aspergillus niger O5 was the only strain that showed loss of Ni, Mn and Co from dissolution when cultivated under shaking conditions (Figures 1 and 2). This was observed after 14 days of experiment (Figure 2A) and was due to the uptake of Ni, Co, and Mn by Aspergillus niger, which is known for nickel accumulation in the biomass (2, 10). Some abiotic factors, such as pH, allow the biosorption of heavy metals by fungi (10); precipitation of metals in the presence of oxalic acid (1) or electrosorption (11) may occur as well. Besides manganese and nickel, up to 40% of cobalt and magnesium was removed from the laterite ore (Figures 1 and 2). Previous mineralogical studies of this Cuban laterite ore revealed that nickel and cobalt are associated with goethite and manganese mineral phases, respectively.

In general, there are two phases in the bioleaching of nickel laterite ore, similar to the bioleaching of chalcopyrite (12). The first phase is characterized by an increasingly high rate of metal dissolution, due to a strong decrease in the pH, which reflects a high metabolic activity of the microorganisms (7). In the second phase, the rate of metal dissolution decreases or remains almost constant.

Figure 2. Laterite bioleaching with A. niger O5 in medium 1a. A) shaking conditions (reciprocal agitation: 100 rpm.min-1) and B) static conditions. Pulp density: 5%, carbon source: sucrose 140g/L, temperature: 30°C. (•: Mn, : Mg, ∆: pH, : Co, : Ni, : Cr, ♦: Al and ◊: Fe)

3.3 Influence of pH and Mg2+ and Mn2+ ion concentrations on the bioleaching Considering that some cations, such as Mn2+, Mg2+, Fe2+, and Fe3+, can affect cell

morphology, enzymatic activity and citric acid production by Aspergillus, they are essential micronutrients in the metabolism. We used some variations of the culture


1132

medium to study the effect of these cations on the release of Ni from Cuban laterite and on the type and concentration of organic acid produced by the fungus.

The composition of the culture medium considerably influenced metal removal as well as type and amount of organic acids produced by strain O5 (Figures 3 and 4).

The highest amount of Ni was leached using medium 3a. After 18 days of incubation, 60% Ni was dissolved, and 86.6% Ni was extracted at the end of experiment (Figure 3C). Medium 3b was the worst for A. niger O5. A different behaviour was observed with Penicillium simplicissimum strain 2 (data not shown).

Regardless of the initial pH in the growth medium, all pH values were around 5 after 7 days of bioleaching, after 14 days pH values of 3.8-4.8 were measured. The pH behaviour in medium 3a was different from that of the other media. The concentration of dissolved Mn, Ni and Co increased when media 1a, 2, 2a and 3 were used. Later, a loss of dissolved metals was observed (Figure 3).

A)

0

20

40

60

80

100

120

0 5 10 15 20 25 30Time (days)

Mn

Dis

solv

ed (%

)

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020406080

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1015202530

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Figure 3. Influence of initial pH, Mg and Mn on bioleaching of Cuban laterite ore by A. niger O5 (A, B, C and D) and change of pH (E). Pulp density: 5%, carbon source: 140g/L sucrose, temperature: 30°C, reciprocal agitation: 100 rpm.min-1. Media with Mg and Mn: •: 1a, : 3, ∆: 3a, : 3b and media without Mg and Mn: : 2, ♦: 2a, x: 1


1133

As shown in Figure 4, the initial pH of the media and the concentrations of Mn2+ and Mg2+ cations influenced the amount and type of organic acid produced by fungal activity. When the strain was cultivated in medium 3a (pH 5.0), the yield of citric acid was considerably enhanced and oxalic acid production was inhibited. This is of considerable advantage, since oxalic acid inhibits Ni laterite bioleaching (1). This strain produced a higher concentration of citric acid (25-27 g/L) when grown in the presence of the ore than in the absence of ore (14 g/L) (data not shown).

0102030405060

Oxa

lic a

cid

(g/L

)

05

1015202530

Citr

ic a

cid

(g/L

)

2a3b 3 1 1a 2 3a

Figure 4. Influence of different culture media on oxalic acid (A) and citric acid (B) production by A. niger strain O5 to10 days of Cuban laterite bioleaching. Pulp density: 5%, carbon source: sucrose 140 g/L, temperature: 30°C, reciprocal agitation: 100 rpm.min-1. Medium: 1, 2, and 2a (without Mg and Mn) and 1a, 3, 3a and 3b (with Mg and Mn)

Some released metal cations inhibit citric acid synthesis by A. niger O5. Mg2+ is a cofactor of the enzyme isocitrate dehydrogenase, and some concentrations of Mn2+ act in a negative way on the citric acid cycle, and citric acid is the best leaching agent for Cuban laterite nickel ore (6). Therefore, we need to define the optimum concentration of metal ions for maximum yield of citric acid.

The X-ray diffraction analysis of some residue samples from this set of experiment demonstrated a new phase in the solid residue after bioleaching. This was identified as magnesium oxalate hydrate (Figure 5).

A

B

Figure 5. X-ray diffraction: A) Cuban laterite ore before bioleaching and B) solid residues from bioleaching with A. niger O5 in medium 3 (pH 5.5)


1134

3.4 Two-stage bioleaching The leaching agent produced in the first step was used for the leaching of the ore in

the second step. The results of the leaching of laterite ore are shown in Figure 6. The only organic acid metabolite accumulated in the filtrate was citric acid (data not shown). The percentage of Ni recovered using this method (Figure 6A) was similar to that obtained using a single stage method (Figure 3B). pH during the leaching phase is shown in Figure 6B as a function of time. Note that the increase in pH with time (Figure 6B). This behavior is due to the release of 30% Mg, which is a neutralizing agent. In the single-stage bioleaching method, the pH decreased with time (Figure 3E) as consequence of metabolite activity of the fungus.

Figure 6. (A) Chemical leaching of Cuban laterite ore for 12 days by metabolic products of A. niger O5 cultivated in medium 3a (pH 5) and (B) pH without ore. Pulp density: 5%, temperature: 30°C and reciprocal agitation: 100 rpm.min-1

4. CONCLUSIONS Cuban laterite ore was more amenable to bioleaching with Aspergillus niger O5 than

with Penicillium simplicissimum. A. niger is capable of producing oxalic and citric acid or only citric acid depending on

the culture conditions. The best leaching agent for Cuban laterite nickel ore was citric acid. The initial pH of culture medium and the composition of growth medium is an important variable for the bioleaching of non-sulfide ore. Cuban sugar cane can be a substitute for sucrose in the culture medium.

There was a high solubilization of nickel (89% after 28 days of bioleaching) in medium 3a with an initial pH of 5.0. The single-stage and two-stage bioleaching methods yielded similar results. On the basis of these results, we recommend that a preliminary engineering and economic study be made.

REFERENCES 1. K. Alibhai, D. Leak, A.W.L. Dudeney, S. Agatzini and P. Tzeferis, in: R.W. Smith and

M. Misra (eds.), Mineral Bioprocessing, The Minerals, Metals and Materials Society, TMS (1991) 191.

2. K. Bosecker, in: R.L. Lawrence, R.M.R. Branion and H.H Ebner (eds.), Fundamental and Applied Biohydrometallurgy, Elsevier, Amsterdam (1986) 367.

3. K. Bosecker, in: J. Salley, R.G.L. McCready, and P.L. Wichlaz (eds.), Biohydrometallurgy, CANMET SP89-10 (1989) 15.


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4. W. Burgstaller and F. Schinner, in: A.E. Torma, J.E. Wey and V.I. Lakshmanan (eds.), Biohydrometallurgical Technologies, Vol. 1, The Minerals, Metals & Materials Society, TMS (1993) 325.

5. W. Burgstaller and F. Schinner, J. Biotechnol. 26 (1993), 340. 6. O. Coto, N. Bruguera, L. Abín, J. Gamboa and Y. Gómez, in: V.S.T Ciminelli and O.

García Jr. (eds.), Biohydrometallurgy: Fundamentals, Technology and Sustainable Development, Part A, Elsevier, Amsterdam (2001) 175.

7. L. Abin, O. Coto, B. Marrero and J. Marrero, Ciencias Biológicas CENIC No 3 (2002) In press.

8. P. Tzeferis, Int. Journal of Mineral Processing, 42 (1994) 267. 9. L.B. Sukla and V.V. Panchanadikar, in: A.E. Torma, J.E. Wey and V.I. Lakshmanan

(eds.), Biohydrometallurgical Technologies, Vol. 1, The Minerals, Metals & Materials Society, TMS (1993) 373.

10. B. Volesky, in: R. Amils and A. Ballester (eds.), Biohydrometallurgy and the Environment towards the Mining of the 21st Century, Part B, Elsevier, Amsterdam (1999) 161.

11. M. Valix, J.Y. Tang and W.H. Cheung, Minerals Engineering., 14 (12) (2001) 1629. 12. O. Coto, A. Ballester, M.L. Blázquez and F. González, Biorecovery, 2 (3) (1993) 121.



1137

Mercury tolerance of thermophilic Bacillus sp. and Ureibacillus sp.

K.J. Glendinning∗ and N.L. Brown

School of Biosciences, The University of Birmingham, Edgbaston, Birmingham, B15 2TT, U.K.

Abstract Although resistance of microorganisms to Hg(II) salts has been widely investigated

and resistant strains have been reported from many eubacterial genera, there are few reports of mercuric ion resistance in extremophilic microorganisms. Thermophilic mercury resistant bacteria were selected by growth at 62°C on Luria agar containing HgCl2. Sequence analysis of 16S rRNA genes of two isolates showed the closest matches to be with Bacillus pallidus and Ureibacillus thermosphaericus. Minimum inhibitory concentration (MIC) values for HgCl2 were found to be 30 µg/ml and 80µg/ml for these isolates respectively, compared to 10 µg/ml for B. pallidus H12 DSM3670, a mercury-sensitive control. The best-characterised mercury-resistant Bacillus strain, B. cereus RC607, had an MIC of 60 µg/ml.

Growth of the thermophilic isolates in media containing HgCl2 resulted in the formation of a black precipitate. X-Ray Diffraction (XRD) analysis of this precipitate showed it to be HgS. Growth in the presence of radioactive 203HgCl2 at 45°C or 62°C showed progressive removal of mercury from the medium, but the mechanism by which this occurs is uncertain. Biochemical assays for mercuric reductase, which reduces 203Hg2+ to volatile 203Hg°, did not detect this enzyme. It is possible that volatilisation of mercury is occurring as a by-product of normal metabolism, through complexation into volatile compounds.

Keywords: Ureibacillus sp., Bacillus sp., thermophilic, mercury resistance

1. INTRODUCTION Compounds of mercury, such as mercuric chloride and organomercurials, have long

been known to be toxic to both eukaryotic and prokaryotic cells. These compounds can pass through biological membranes [1] and bind with high affinity to thiol (SH) groups in proteins [2], thus causing damage to membranes and inactivating enzymes. Mercury is also genotoxic; inorganic Hg(II) is capable of strong reversible interactions with the nitrogens in purines and pyrimidines, and organic mercury compounds, e.g. methylmercury, also produce irreversible damage to nucleic acids [3]. Environmental contamination with mercury compounds can have devastating effects, as mercury toxicity ∗ Corresponding author. Current address: Birmingham Women's Health Care, Metchley Park Road, Birmingham B15 2TG, UK. Email: [email protected]; Fax: +44-121-627-2711.


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is cumulative, with the highest levels of mercury compounds being found in consumers at the top of the food chain.

A number of microorganisms have evolved resistance mechanisms to deal with mercury compounds. Mercury resistance was first reported in Staphylococcus aureus [4] and since then has been described in a number of bacterial species.

One of the best-defined mercury resistance determinants is the mer operon encoded by transposon Tn501, found in Gram-negative bacteria. The functions of the minimal number of proteins required to confer full resistance are as follows [5]:

MerR is the mer regulatory protein which controls the expression of all the other proteins in the operon in response to the presence of Hg(II) salts. MerP is a periplasmic protein which binds mercuric ions entering the periplasm through the porins OmpC and OmpF, and transfers Hg(II) to the MerT transport protein located in the cytoplasmic membrane. The mercuric ions are then passed from cysteine residues in MerT which are exposed to the periplasmic compartment to cysteine residues on the cytoplasmic face of the membrane, and thence to the cytoplasmic mercuric reductase (MR, the merA gene product). Hg(II) is complexed with the active site cysteine residues of MR and reduced from Hg(II) to Hg(0) using NADPH as the reductant. Hg(0) is then lost from the cell in the gas phase.

Mercuric ion resistance has also been characterised in Gram-positive genera. The resistance determinants from Bacillus sp. RC607 [6], Streptomyces lividans [7] and Staphylococcus aureus [8] have been characterised. The mechanisms of induction, mercury transport and mercury reduction in Gram-positive genera are similar to those of the Gram-negative systems, but the transport functions differ, as might be expected from the differences in the surfaces of Gram-positive and Gram-negative cells [5].

Although resistance to Hg(II) has been widely investigated, there are few reports of mercuric ion resistance in extremophilic microorganisms [9, 10]. Given the wide spread of mercuric ion resistance/tolerance throughout the bacterial kingdom, it seems likely that many extremophilic microorganisms will encode mercury resistance. This paper describes the investigation of such resistances.

Here we report the isolation of thermophilic Bacillus sp. [11] and Ureibacillus sp. [12] and characterisation of their observed mercuric ion resistance.


2.1 Isolation and identification of organisms from compost Suspensions of garden compost were plated onto LB agar plates supplemented with 0,

15 and 20 µg/ml HgCl2, and the plates were incubated at 62°C for 48 hours. Colonies were picked from the LB agar plates containing mercury and repeatedly subcultured on LB agar containing 15 µg/ml HgCl2 to obtain two axenic cultures. Approximately 700bp of 16S rDNA gene was amplified from each organism using primers pA [13] and 530r [14]. The PCR products obtained were sequenced and subjected to BLAST analysis.

2.2 HgCl2 resistance testing Minimum Inhibitory Concentration (MIC) assays for HgCl2 were performed on

Ureibacillus sp. and Bacillus sp. The controls used were a mercuric ion sensitive strain, Bacillus pallidus H12 DSM3670, and a known mer operon-containing mercury resistant Bacillus cereus RC6067 [6]. MIC assays were performed on LB agar plates containing


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HgCl2 in 10 µg/ml increments from 0µg/ml to 100µg/ml. Plates were incubated at 37°C and 62°C.

2.3 Identification of culture precipitate It was noted that when Bacillus sp. and Ureibacillus sp. were grown in culture with

HgCl2 a black precipitate was formed. Bacillus sp. cells were grown with and without HgCl2 followed by transmission electron microscopy (TEM) to localise the precipitate. In order to test whether the precipitate could be HgS formed by the reaction between Hg(II) and microbially produced H2S [15], [16], H2S production was determined. The positive control used was Proteus vulgaris, a known aerobic H2S producer, with Escherichia coli as the negative control. P. vulgaris, E. coli and B. cereus RC607 were grown overnight at 37°C in LB broth, and B. pallidus H12 DSM 3670, Ureibacillus sp. and Bacillus sp. were grown overnight at 62°C in LB broth. Overnight cultures were diluted 1:100 into sterile glass universal bottles containing 5ml volumes of fresh LB broth with 0µg/ml and 10 µg/ml HgCl2 (only used for mercury resistant strains). Lead acetate strips (BDH) were folded over the neck of the bottles (care was taken not to wet the strips) and the caps were screwed on. These were then incubated at 37°C or 62°C (as above) and examined after 24h and 48h.

In order to identify the precipitate, Bacillus sp. cells were grown in the presence of 50 µg/ml HgCl2 for 48 hours. Cells and precipitate were harvested by centrifugation and separated by careful washing with sterile distilled water. The precipitate was dried and examined by X-ray diffraction (XRD) analysis.

2.4 Mercury volatilisation assays Volatilisation assays were performed essentially by the method of [17] with the

exception that LB broth was used instead of M9 medium and induction was performed with 2.5 µg/ml HgCl2 for mercury resistant isolates and 0.25 µg/ml for sensitive isolates.

Cultures were grown to OD600 = 0.4-0.5, then preinduced with HgCl2 for one hour. Cultures were adjusted to precisely OD600 = 0.4 followed by washing and resuspension of cells in fresh LB broth. Three 1ml aliquots of the resuspended cells were dispensed into separate 50ml conical flasks. Three control flasks each containing 1ml uninoculated LB broth were prepared. Flasks were placed in a shaking waterbath set to the desired temperature and to each flask was added 4ml pre-warmed LB broth supplemented with 1.25µM HgCl2 and 1:2000 diluted 203HgCl2 (at a concentration of 2.55mg/ml Hg; Amersham Pharmacia plc). Timing commenced upon the removal of the first sample. 100 µl samples were taken from each flask at 0, 20, 40, 60, 120, 180 and 240 minutes. These samples were immediately added to 5ml volumes of scintillation cocktail (OptiPhase "HiSafe" 2) and counted in a Packard TR1-CARB 2700 TR liquid scintillation analyser with the window set at 0.0-220 MeV.

2.4.1 Mercuric reductase assays Mercuric reductase assays were performed by the method of Lund and Brown [18]. Cells were grown overnight, then inoculated 1:25 into fresh LB broth without

selection. Cells were grown at the appropriate temperature (45°C or 62°C) with shaking to OD600 = 0.5-0.6, induced with 10 µg/ml HgCl2 for one hour, then a further 10 µg/ml HgCl2 for a further hour (mercury sensitive B. pallidus H12 DSM 3670 cells were induced with 1 µg/ml HgCl2 for one hour followed by a further 1 µg/ml HgCl2 for a further hour). Cells were harvested by centrifugation, then washed in 20 ml wash buffer (30mM Tris-HCl


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pH7.6, 20% sucrose, 1mM DTT, 1mM EDTA). The cell suspension was re-centrifuged and the supernatant discarded. Cell pellets were washed twice with 20ml resuspension buffer (100mM sodium phosphate pH7.0, 1mM β-mercaptoethanol, 0.5mM EDTA, 0.05mM PMSF, 0.1mM benzamidine) followed by resuspension in 5ml fresh buffer. The cells were lysed by sonication and cell debris was pelleted by centifugation. Supernatants were retained and the pellets discarded. The supernatant was clarified by two further rounds of centrifugation in order to remove all traces of cellular debris. Lysates were stored at 0°C overnight. The following day, 7ml aliquots of assay buffer (0.1M sodium phosphate buffer pH 7.0, 1mM β-mercaptoethanol, 0.5mM EDTA, 0.1mM NADPH) were dispensed into sterile glass universal bottles and stored on ice until use. For the assay, 70µl of cell extract was added to 7ml assay buffer, mixed well and the mixture was pre-incubated at the desired temperature for 5 minutes. 1ml was removed and an A340 reading was taken. This was immediately followed by addition of 60µl of 10mM HgCl2 stock, which gave a final concentration in the assay mix of 100µM HgCl2; 60µl sterile distilled water was added to the control assay mix. Samples were mixed well and incubated at the desired temperature for 1 minute followed by an A340 reading. Further readings were taken at 3 and 5 minutes after addition of HgCl2/sterile distilled water.


3.1 Identification of microorganisms The 16S rDNA PCR products obtained from the two cultures were sequenced and

BLAST analysis of the sequences obtained revealed 97% homology with Bacillus pallidus H12 DSM3670 [11] and >99% homology with Ureibacillus thermosphaericus [12].

3.2 HgCl2 resistance

Table 1. HgCl2 MIC values (µg/ml) for Bacillus sp. and Ureibacillus sp.

Organism HgCl2MIC (mg/ml) at 37°C HgCl2MIC (mg/ml) at 62°C

Bacillus cereus RC607 60 - Bacillus pallidus DSM 3670 10 10 Bacillus sp. 60 80 Ureibacillus sp. 30 30

No data are shown for Bacillus cereus at 62°C as this temperature was outside the growth range for this organism.

The results obtained from MIC testing showed that mercury resistance of the thermophilic isolates was not temperature dependent. The growth of Bacillus sp. was reduced on normal LB agar at 37°C; this temperature is near to the bottom end of the organism's growth range, which may help explain why mercury resistance was reduced, but not abolished.

3.3 TEM of Bacillus sp. The results of Transmission Electron Microscopy (TEM) (Figure 1) show a clear

difference in the appearance of bacterial cultures grown with and without mercury. There is a large amount of black precipitate visible when Hg(II) is present, which was thought to be HgS. The appearance of the cells in the TEM suggests that the precipitate may be forming within the cells, possibly causing cells to rupture.


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Figure 1(a). TEM (10,000 X magnification) of Bacillus sp. grown with 50µg/ml HgCl2

Figure 1(b): TEM (10,000X magnification) of Bacillus sp. grown without HgCl2

3.4 Identification of the precipitate XRD analysis was performed upon a sample of dried crystalline precipitate collected

from a culture of Bacillus sp. grown with HgCl2. The thick black lines on the XRD profile (Figure 2) show the peaks obtained for the experimental sample, and the fainter, vertical lines show the peaks obtained from a database for an ultra-pure reference sample of β-HgS. There was sufficiently good correlation between the two sets of information to make a positive identification of the experimental sample as β-HgS. The profile was very similar to one obtained for HgS precipitated by Desulfovibrio desulfuricans [19].


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Figure 2. X-Ray Dispersion Analysis of the precipitate from Bacillus sp. grown in the presence of Hg(II)

3.5 Mercury volatilisation assays The results of mercury volatilisation experiments (Figure 3) showed that Ureibacillus

sp. removed Hg(II) from the media more efficiently than Bacillus sp. at 45°C than at 62°C. This may reflect the growth temperature optima for these organisms. From 16S rRNA data, the most closely related species to Bacillus sp. was Bacillus pallidus, which is reported as having a maximum growth temperature of 70°C with a temperature optimum of 60-65°C. The closest species to Ureibacillus sp. was Ureibacillus thermosphaericus which has a maximum growth temperature of 64°C.

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Figure 3: Results of volatilisation assays on Bacillus spp. and Ureibacillus sp. at 62°C and 45°C. Results are shown as the means of triplicate readings with error bars shown as sample standard deviations (σn-1)

This suggests that Ureibacillus sp. probably has a lower optimum growth temperature than Bacillus sp., hence Ureibacillus sp. may remove 203Hg(II) more efficiently at 45°C. At 45°C, 203Hg(II) was being removed from the media at a high rate by B. cereus RC607.


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This rate was reduced at 62°C, probably due to the fact that 62°C was outside the growth range of this organism. There appeared to be no appreciable difference in the rates of removal in LB broth alone and in the presence of B. pallidus H12 DSM3670 at 45°C. There did appear to be some removal of 203Hg(II) from the media by B. pallidus H12 DSM3670 at 62°C, the reason for which is not known.

3.6 Mercuric reductase assays The results of mercuric reductase assays are shown in Figure 4. Mercuric reductase

activity was observed in B. cereus RC607 at 37°, 45° and 62°C. There appeared to be no specific NADPH-dependent mercuric reductase activity in B. pallidus H12 DSM3670, Ureibacillus sp. or Bacillus sp. This result was surprising in view of the mercuric ion resistance and results from volatilisation assays, which showed Ureibacillus sp. and Bacillus sp. as capable of removing 203Hg(II) from the medium. An alternative, non-mer mechanism of mercury detoxification and volatilisation may operate within Ureibacillus sp. and Bacillus sp., possibly as a by-product of normal metabolism.

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Figure 4. Mercuric reductase activity of Bacillus spp. and Ureibacillus sp. at 37°C, 45°C and 62°C. Errors bars show one standard deviation (σn-1) of triplicate readings

3.7 Testing for H2S production The precipitation of HgS by microorganisms is often due to a reaction of Hg(II) with

H2S gas. The strains in this study were tested for H2S production as described in section 2.3. Growth occurred in all bottles; however, blackening of the lead acetate strip only occurred in the bottle containing P. vulgaris, indicating that only this isolate produced detectable amounts of H2S. It seems likely, therefore, that any production of HgS by Ureibacillus sp. and Bacillus sp. proceeded due to complexation of Hg(II) with cellular sulphides and not with H2S.


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4. CONCLUSIONS Mercuric ion tolerant isolates of thermophilic Bacillus sp. and Ureibacillus sp. were

isolated from compost. Mercury tolerance in these isolates was apparently not due to classical mer operon-mediated mercury reduction, but by some other mechanism. Mercury tolerance was not temperature dependent, as both isolates showed elevated tolerance to mercuric ions at 37°C and 62°C.

When considering features of mercuric ion resistance in these isolates it is important to make the distinction between mercury resistance and mercury tolerance. Mercuric ion resistance can be thought of as a genetically encoded detoxification mechanism, which is specifically induced in response to mercurials. Mercury tolerance can be thought of as a detoxification mechanism, which is a by-product of normal metabolism and is not specifically induced [19]. From the results obtained it seems possible that Ureibacillus sp. and Bacillus sp. are exhibiting mercury tolerance mechanisms rather than a specific mer operon-encoded mercury resistance. The evidence to support this is the lack of detectable NADPH-dependent mercuric reductase activity, and precipitation of HgS in the medium. Several organisms have been reported as being mercury tolerant due to HgS precipitation [15, 16, 20, 21]. This mechanism has not been shown to be specifically induced in response to mercurials, suggesting that it is indeed a by-product of normal metabolism.

HgS can be formed by various mechanisms in microbial mercury cycling. The growth of Desulfovibrio desulfuricans under sulphate-reducing conditions resulted in the precipitation of HgS when the medium was spiked with Hg(II) [15, 19]. Hg(II) reacts directly with H2S produced by D. desulfuricans [15] and Clostridium cochlearium [16] to form HgS. This is thought to provide a means of mercury tolerance in these organisms. The formation of HgS from monomethylmercury (MMHg) by D. desulfuricans may occur via the reaction of MMHg with H2S to produce unstable dimethylmercurysulphide (DMHgS), which breaks down to HgS and volatile dimethylmercury (DMHg) [15]. Precipitation of HgS by aerobic microorganisms has also been suggested as a tolerance mechanism. Klebsiella aerogenes NCTC418 was thought to form HgS when grown in continuous (aerobic) culture with the addition of 2ppm (equivalent to 2µg/ml) HgCl2. No bacterial mercury volatilisation could be detected, and it was claimed mercuric ion sensitivity was increased under sulphate-limited conditions [20]. K. aerogenes is unable to produce H2S gas, therefore any precipitation of HgS presumably occurred via an alternative mechanism.

There are few reports of precipitation of HgS by bacterial cultures, which makes it of unknown importance in microbial mercury detoxification. At least part of the mercury tolerance shown by Ureibacillus sp. and Bacillus sp. may be due to precipitation of HgS. This probably occurs by reaction of Hg(II) with cellular sulphides, as H2S was not produced by these microorganisms when grown with or without HgCl2.

The results of mercuric reductase assays showed that no NADPH-dependent mercuric reductase activity was present in Ureibacillus sp. and Bacillus sp., therefore mercury removal was not due to reduction of Hg(II) to Hg(0) by mercuric reductase, unless it uniquely uses another cofacter. It is possible that Ureibacillus sp. and Bacillus sp. may produce high levels of non-specific reductases which are capable of reducing Hg(II) to Hg(0). Alternatively, there may be a different, as yet unknown, mechanism for mercury removal operating in these organisms.

A possible explanation for removal of mercury from the media by Ureibacillus sp. and Bacillus sp. may be the methylation of mercury to either MMHg or DMHg. MMHg


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and DMHg are volatile, and could be lost from the medium in a similar fashion to Hg(0) The formation of DMHg by microorganisms may represent a mercury detoxification method due to the volatility of the compound [19]. Methylation of mercury by anaerobic microorganisms has been reported [22]. At present there are no reliable reports of mercury methylation by aerobic cultures.

The mechanisms of mercury tolerance in Ureibacillus sp. and Bacillus sp. are not fully understood. Whilst precipitation of HgS in culture is one tolerance mechanism, the means by which mercury volatilisation occurs in the absence of specific mercuric reductase is unclear.

ACKNOWLEDGEMENTS This work was supported by a BBSRC CASE studentship to KJG. Thanks are due to

Dr Jon L. Hobman and Dr Ching Leang for discussion and advice in setting up assay systems.

REFERENCES 1. Gutknecht, J. J. Memb. Biol. (1981). 61:61. 2. Leach, S.J. J. Aust. Chem. (1960). 13:520. 3. Sletten, E. and Nerdal, W. Metal Ions In Biological Systems. (1997). 34:479. 4. Moore, B. Lancet. (1960). II:453. 5. Hobman, J.L. and Brown, N.L. Metal Ions In Biological Systems. (1997). 34:527. 6. Wang, Y., Moore, M.,Levinson, H.S., Silver, S., Walsh, C. and Mahler, I. J. Bacteriol.

(1989). 171:83. 7. Sedlmeier, R. and Altenbuchner, J. Mol. Gen. Genet. (1992). 236:76. 8. Laddaga, R.A., Chu, L., Misra, T.K. and Silver, S. Proc. Natl. Acad. Sci. (1987).

84:5106. 9. Olson, G.J., Porter, F.D., Rubinstein, J. and Silver, S. J. Bacteriol. (1982). 151:1230. 10. Bogdanova, E.S., Mindlin, S.Z., Kalyaeva, E.S. and Nikiforov, V.G. FEBS Lett.

(1988). 234:280. 11. Scholz, T., Demharter, W., Hensell, R. and Kandler, O. Syst. Appl. Microbiol. (1987).

9:91. 12. Fortina, M.G., Rüdiger, P., Schumann, P., Mora, D., Parini, C., Manachini, L.P. and

Stackebrandt, E. Int. J. Syst. Evol. Microbiol. (2001). 51:447 13. Edwards, U., Rogall, T., Blocker, H., Emde, M. and Bottger, E.C. Nucleic Acids Res.

(1989). 17:7843. 14. Lane, D.J., Pace, B., Olsen, G.J., Stahl, D.A., Sogin, M.L. and Pace, N.R. Proc.

Natl.Acad. Sci. (1985). 82:6955. 15. Baldi, F., Pepi, M. and Filippelli, M. Appl. Environ. Microbiol. (1993). 59:2479. 16. Pan-Hou, H. and Imura, N. Arch. Microbiol. (1981). 129:49. 17. Leang, C., PhD thesis (1999), University of Birmingham: Birmingham, UK. 18. Lund, P.A. and Brown, N.L. Gene. (1987). 52:207. 19. Baldi, F. Metal Ions In Biological Systems. (1997). 34:213. 20. Aiking, H., Govers, H. and van 't Riet, J. Appl. Environ. Microbiol. (1985). 50:1262. 21. Belliveau, B.H., Starodub, M.E. and Trevors, J.T. Can. J. Microbiol. (1991). 37:513. 22. Choi, S.C. and Bartha, R. Appl. Environ. Microbiol. (1993). 59:290.



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Reduction of Pd(II) with Desulfovibrio fructosovorans, its [Fe]-only hydrogenase negative mutant and the activity of

the obtained hybrid bioinorganic catalysts

I.P. Mikheenkoa, V.S. Baxter-Planta, M. Roussetb, S. Dementinb, G. Adryanczyk-Perrierb and L.E. Macaskiea

a School of Biosciences, the University of Birmingham, Edgbaston, Birmingham, B15 2TT, England, United Kingdom

b Bioenérgétique et Ingénierie des Protéines, CNRS, 31 Chemin Cedex 20, Joseph Aiguier, 13402 Marseille France

Abstract A novel biocatalyst was obtained via reduction of Pd(II) to Pd(0), at the expense of

hydrogen, onto the cell surface of Desulfovibrio fructosovorans and a mutant of D. fructosovorans with deactivated [Fe]-only hydrogenase. The ability of the palladium-coated biomass to reduce chromium (Cr+6 to Cr+3), to release hydrogen from sodium hypophosphite and reductively dehalogenate chlorophenol and polychlorinated biphenyl congeners was demonstrated. Dried, palladised cells of D. fructosovorans wild type and its mutant were more effective catalysts than Pd(II) reduced chemically under hydrogen or commercially available sub-micron size Pd(0) powder. Differences were observed in the catalytic activity of the wild type and the mutant strain of D. fructosovorans when compared with each other. Negligible chloride release occurred from the chlorophenol and polychlorinated biphenyl species using biomass alone. The structure of the bioinorganic catalyst was investigated using transmission electron microscopy.

1. INTRODUCTION Palladium is one of the world’s most expensive metals. Catalyst systems based on

palladium are widely used in chemical manufacturing and processing (1). The highest palladium consumer is the auto motive industry, where this metal, together with other platinum group metals, is used in automobile catalytic converters to reduce the toxicity of exhaust gases. Palladium is also extensively used in the electronic industry. Since palladium is a highly valuable metal with only limited world resources (2), developing new methods of Pd recovery and reprocessing of scrap is necessary since existing hydro- and pyrometallurgical routes are energy demanding and/or not environmentally friendly.

In the work of Lloyd et al. (3) it was shown that palladium could be effectively recovered from solution by resting cells of the sulphate-reducing bacterium Desulfovibrio desulfuricans at ambient temperature. It was concluded that Pd(II) was reduced by the activity of hydrogenase and possibly cytochrome c3.

Yong et al. (4) demonstrated that palladised D. desulfuricans biomass could be used as a catalyst without any further processing. It has also been reported (5, 6) that in the


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process of biorecovery metallic Pd with unusual magnetic properties could be obtained: ferromagnetic Pd nanocrystals were formed on the biomass. Subsequent studies showed (7) that biologically reduced palladium (Bio-Pd) can be used as a catalyst for partial dehalogenation of polychlorinated biphenyls, which are very stable recalcitrant xenobiotics which are not generally susceptible to aerobic or anaerobic microbial metabolization.

The catalytic properties of biomineralized palladium (Bio-Pd) depend on the size and the distribution pattern of metal deposits on the cell surface. To obtain a Bio-Pd catalyst with certain catalytic properties it is important to understand the process of Pd(II) reduction and Pd(0) grain formation which take place within the periplasm of sulphate reducing bacteria (5). According to Lloyd et al. (3) hydrogenase is one of the important components of the Pd(II) reducing system. Sulphate-reducers have more than one hydrogenase system, with Desulfovibrio sp. possessing up to four different hydrogenases, which differ by structure (enzymes with Fe, Fe-Ni or Fe-Ni(Se) in their active centres) and localisation within the cell (periplasmic and cytoplasmic, membrane bound and non-bound) (8). Several hydrogenase mutants of Desulfovibrio sp. have been constructed lacking different hydrogenases (9, 10). The aim of this work is to investigate the catalytic properties of Bio-Pd obtained via Pd(II) reduction by the sulphate reducer D. fructosovorans and its [Fe]-only hydrogenase mutant. The hydrogenase and cytochrome complex of the Desulfovibrio sp. has been intensively investigated (8, 11, 12 ).

2. EXPERIMENTAL

2.1 Preparation of cells D. fructosovorans and its mutant (10) were maintained anaerobically (13). Growth

medium was inoculated with 10% of culture and cells were incubated at 37 °C for three days before harvesting. The medium for D.fructosovorans culture was as in (13).

For Bio-Pd production the cells were harvested by centrifugation at 6000 rpm for 15 min at 4°C, washed three times in degassed 20 mM MOPS buffer at pH 7.0. The cells were stored refrigerated.

2.2 Preparation of Bio-Pd Aliquots of the washed cells were resuspended as a concentrated suspension (diluted

to 4-6 mg dry wt/ml for use) in MOPS buffer and transferred under N2 to 100 ml of Pd(II) (2 mM sodium tetrachloropalladate(II) Na2PdCl4 (Aldrich Chemical Company) in 0.01 M HNO3, pH 2.3) solution in 100 ml serum bottles sealed with butyl rubber stoppers and pre-equilibrated with N2. The suspension was allowed to stand for 1 h at 30°C to form nucleation sites on the biomass. The electron donor used was H2 gas, which was bubbled through the solution for 20 min. After the black precipitate had formed, the solution was centrifuged at 6000 rpm for 10 min and the precipitate was washed three times with distilled water, dried in air and then washed with acetone. Chemical palladium was prepared identically except that no biomass was added and reduction took approximately 60 min under H2 in the absence of cells.

2.3 Pd(II) and Cr(VI) assay The kinetics of Pd(II) reduction were studied by recording the A420 of the residual

Na2PdCl4 solution, withdrawn (1.5 ml) from the reaction mixture each 20 min, (Unicam 5600 UV-VIS spectrophotometer: Cambridge, UK). The maximum height of the specific


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absorption peak of [PdCl4]2- at 420 nm depends linearly on the palladium salt concentration. The concentration of Pd(II) in solution was also determined spectrophotometrically with Sn(II) (14). For the Pd(II) assay 200 µl of sample was added to 800 µl of SnCl2 solution. Readings were taken after 60 min at a wavelength of 463 nm.

The residual concentration of Cr(VI) (where Pd(0) was used as a catalyst for the reduction of Cr(VI) (below)) was measured using the diphenylcarbazide method (5). Test solution (100 µl) was added to 1.4 ml of the diphenylcarbazide solution, mixed well, left to stabilise for 5 min and read at 540 nm.

2.4 Examination of Pd-loaded cells by electron microscopy Pellets of Pd-loaded bacteria were harvested using a Heraeus Sepatech Biofuge 13

microfuge (13000 rpm, 5 min) and fixed in 2.5% (wt/vol) glutaraldehyde in 0.1 M Na-cacodylate buffer, pH 6.8. After centrifugation the pellet was resuspended in 1.5 ml of 0.1 M Na-cacodylate buffer, pH 6.8. The preparation was stained in 1% osmium tetroxide in 0.1 M phosphate buffer, pH 7.0, dehydrated and embedded in epoxy resin. Sections (100 - 150 nm thick) were cut from the resin block and placed onto a copper grid prior to analyses. Sections were viewed with a JEOL 120CX2 transmission electron microscope.

2.5 Catalytic activity measurement Three test reactions were used to characterise the catalytic activity of the Bio-Pd:

hydrogen release from sodium hypophosphite, Cr(VI) reduction and reductive dehalogenation (RD) of polychlorinated aromatic compounds.

2.5.1 Hydrogen release from sodium hypophosphite The reaction was initiated via the addition of 4 mg of Bio-Pd (Pd:dry wt biomass 1:3)

to 20 ml of a 2% solution of hypophosphite (NaH2PO2) at 30°C. The volume of liberated hydrogen was recorded at 5 min intervals starting 20 mins after the initiation of the reaction (reaction onset time). The reaction rate was calculated as the rate of H2 release from 5 replicate experiments. Chemical palladium (4 mg) was used as a control test reaction.

2.5.2 Cr(VI) reduction Tests were performed under oxygen free nitrogen (OFN) in 12 ml sealed serum

bottles. Bio-Pd (2 mg) was added to 5 ml of 7 mM Na2CrO4 in 20 mM MOPS-NaOH buffer at pH 7.0. The solution was equilibrated at 30°C for 10 min and the reaction was initiated by addition of sodium formate (electron donor) to a final concentration of 25 mM. Samples were taken from the bottle via a rubber septum and centrifuged for 5 min at 13000 rpm to remove Pd. The supernatant was analysed for residual Cr (VI). The catalytic activity was expressed as the percent of reduced Cr(VI). Chemical palladium (2 mg) was used as a control test reaction.

2.5.3 Reductive dehalogenation of chlorinated aromatic compounds The Bio-Pd (2 mg, 1:3 Pd:dry wt biomass) was placed in 12 ml serum bottles and the

chlorinated aromatic compound added to the required concentration in a carrier of 20 mM MOPS NaOH buffer pH 7.0. Experiments were initiated via the addition of formate (10 mM sodium formate, pH 7.3). The final volume was 10ml. Dehalogenation of the chlorinated aromatic substance was estimated in periodically withdrawn samples as release of chloride ion assayed spectrophotometrically by the mercury (II) thiocyanate method (15) versus sodium chloride as standard. For these tests the washing procedures


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were carried out during preparation of the Bio-Pd with omission of Cl- from all processing steps to ensure a low background level of Cl-, against which the Cl- release was measured. The limit of sensitivity was in the range 0.5-100 µg/ml Cl-.

2.5.4 Chlorinated aromatic compounds Chlorophenols were used as aqueous solutions diluted to the required concentration

(usually 5 mM) in the media. Polychlorinated biphenyls (PCBs) were dissolved in hexane and diluted to the required nominal concentrations (80 µg/ml as chloride; shaken to mix) as hexane suspensions in the media.

3. RESULTS AND DISCUSSION The bioreduction of Pd(II) by D. fructosovorans and its [Fe]-only hydrogenase mutant

was compared. Solutions (2 mM) of sodium tetrachloropalladate (Na2PdCl4) in 0.01 M HNO3 or tetraamminepalladium chloride ([Pd(NH3)4]Cl2) in 20 mM MOPS buffer, pH 7.0 were poured into 10 ml sealed vials and degassed under OFN. The resting cell suspensions of each strain were added to the Pd(II) solution (3 parts dry weight of biomass to 1 one part of Pd(II) and left for biosorption at 30°C. After 1 h samples of supernatant were analysed for remaining Pd(II). This concentration of Pd(II) was regarded as the time-zero concentration.

Following the one hour of biosorption the Pd(II) bioreduction was initiated via the addition of electron donor (20 mM sodium formate solution, pH 7.0). The reaction was carried out anaerobically at 30°C. The residual Pd concentration in the supernatant is shown in Fig. 1.

Figure 1. Reduction of Pd(II) by D. fructosovorans, wild type and [Fe]-only hydrogenase negative mutant. Mass ratio Pd(II) dry biomass is was 1:3. Solution: 2 mM Na2PdCl4, in 0.1 M HNO3. Electron donor 0.25 mM sodium formate

The rate of Pd(II) reduction was monitored via the decreasing concentration of the coloured [PdCl4]-2 complex ion (A420), and confirmed using SnCl2. Both methods gave identical results. The reduction of Pd by sodium formate alone was used as a control. Dead cells (killed by autoclaving at 121˚C for 15 mins) did not reduce Pd (not shown). Fig 1 shows that both wild type and mutant reduce Pd(II) at approximately the same rate hence the absence of [Fe]-only hydrogenase did not significantly effect the Pd(II) reduction process. It should be noted that [Fe]-only periplasmic hydrogenase is less abundant


1151

compared to the [Fe-Ni] enzyme, which constitutes a larger part of hydrogenase activity in the periplasm (8).

The catalytic activity of Bio-Pd samples obtained using D. fructosovorans and its mutant with respect to Cr(VI) reduction was substantially higher when compared to palladium powder of 39 µm particles size (Fig. 2). The Bio-Pd obtained under acidic conditions (pH 2.3) using Na2PdCl4 salt (Fig.2a) proved to be a better catalyst when compared to the catalyst obtained at pH 7.0 using [Pd(NH3)4]Cl2 salt (Fig.2b). Interestingly, Bio-Pd obtained with [Fe]-only hydrogenase negative mutant was significantly more catalytically active when compared to Bio-Pd obtained from the wild type regardless of which palladium salt or pH was used in the preparation of Bio-Pd (Fig. 2).

1 2 3

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1 - Control: commercial Pd powder ; Bio-Pd: 2 - wild type; 3 - [Fe]-only hydrogenase negative mutant.

Figure 2. The reduction of Cr(VI) to Cr(III) using D. fructosovorans and its [Fe]-only hydrogenase negative mutant Bio-Pd produced with Na2PdCl4 salt, pH 2.3 (a); and [Pd(NH3)4]Cl2 salt, pH 7.0 (b)

The Bio-Pd was also a more active catalyst than its chemical counterpart in the release of hydrogen from sodium hypophosphite (Fig. 3). The rate of hydrogen release was higher in the presence of mutant-based Bio-Pd catalyst obtained with Na2PdCl4 salt (made at pH 2.3) (Fig. 3a) when compared to wild type-based Bio-Pd. However, in the case of Bio-Pd obtained using [Pd(NH3)4]Cl2 salt made at pH 7.0 (Fig. 3b) the difference between the two types of biocatalyst was not significant.

1 2 30,00

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1 - Control: commercial Pd powder ; Bio-Pd: 2 - wild type; 3 - [Fe]-only hydrogenase negative mutant.

Figure 3. The release of hydrogen from sodium hypophosphite using D. fructosovorans and its [Fe]-only hydrogenae negative mutants produced Bio-Pd with Na2PdCl4 salt, pH 2.3 (a); and [Pd(NH3)4]Cl2 salt, pH 7.0 (b) bars. Bars indicate standard error


1152

The reductive dehalogenation of chlorinated aromatic compounds was investigated. Chloride release from 4-chlorobiphenyl, 2,3,4,5-tetrachlorobiphenyl and 2,2’,4,4’,6,6’-hexachlorobiphenyl using Desulfovibrio fructosovorans and its hydrogenase negative mutant Bio-Pd produced with Na2PdCl4 salt (pH 2.3) was tested. Throughout, no Cl- liberation was promoted by non-palladised bacteria, chemically prepared Pd(0) or commercially available finely divided Pd(0) (not shown) although the Bio-Pd promoted Cl- release from all the PCBs (Figs. 4a-c). However only small differences were seen between the parent and the mutant Bio-Pds in each case.

1 2 3 4 5

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16

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monoPCB wild type monoPCB [Fe]-negative

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HexaPCB wild type HexaPCB [Fe]-negative

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4

8

12

16

C

hlor

ide,

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Time, hours

2-chlorophenol wild type 2-chlorophenol [Fe]-negative

hydrogenase mutant

d

Figure 4. Reductive dehalogenation of 4-chlorobiphenyl (a) and 2,3,4,5-tetrachlorobiphenyl (b) of 2,2’,4,4’,6,6-hexachlorobiphenyl (c) and 2-chlorophenol (d) using D.fructosovorans and its hydrogenase negative mutants Bio-Pd

The rate of chloride release from 5 mM 2-chlorophenol was also studied under similar conditions. In this case the rate of Cl- release for the [Fe]-only hydrogenase mutant was lower than for the wild type (Fig. 4d). The controls included resting cells of the bacteria without bound Pd(0), chemically prepared Pd(0) and commercially available finely divided Pd(0). In confirmation of a previous study the chemically prepared Pd(0) and commercially available finely divided Pd(0) gave a similar rate of Cl- release to D. fructosovorans, whereas no Cl- liberation was promoted by the non-palladised bacteria (not shown).

Thus, the release of Cl- from 2-chlorophenol was achieved using both Bio-Pd preparations, with no advantage over "chemical" Pd(0). However, the reductive dehalogenation of polychlorinated biphenyls was achieved using the Bio-Pd obtained from D. fructosovorans and its hydrogenase negative mutant under conditions where the chemically produced Pd counterpart is largely ineffective, but no advantage was demonstrated using the mutant strain.

The bioreduction tests and studies using X-ray diffraction analysis (not shown) showed that D. fructosovorans and its [Fe]-only hydrogenase negative mutant reduced


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palladium to its metallic state on their surface. Electron microscopic investigation of Bio-Pd in this study showed that for D. fructosovorans and its mutant the initial reduction of Pd takes place in the same way as that previously described for in D. desulfuricans (5). It begins within the periplasmic space (the site of localisation of hydrogenases and cytochromes) and then the growing crystals protrude through the outer cell membrane and form larger crystals on the cell surface (Fig. 5). The pattern of Pd crystal deposition was very similar in both the wild type and mutant stain. The catalytic activity of Bio-Pd obtained using the [Fe]-only hydrogenase negative mutant was higher when compared to the Bio-Pd obtained using wild type strain in the reduction of Cr(VI) and in the other tests the difference varied according to the Pd(II) salt used and pH of its preparation (hydrogen release tests) or chlorinated aromatic substrate (RD tests). We assume that the reason for differences in catalytic activity is the size and, consequently, the surface area of palladium crystals formed in the process of Pd(II) reduction. It may be possible that the lack of one of the hydrogenases leads to formation of a larger quantity of Pd(0) nanoclusters via the activity of other redox-enzymes within the periplasm. These nanoparticles are not visible on the electron micrographs but the presence within the preparation can be detected through their enhanced catalytic activity. The presence of such Pd nanoclusters was shown in the Bio-Pd preparations of D. desulfuricans via the measurement of the magnetic moment of Bio-Pd powder in a varying external magnetic field (5, 6); magnetic measurements to determine the Pd-cluster size on the wild type and mutant D. fructosovorans are in progress.

Figure 5. Electron micrographs of Pd deposition on the surface of D. fructosovorans (a) and its mutant (b) cell surface. Scale bars indicate 500 nm

4 CONCLUSIONS The catalytic activity of Bio-Pd obtained under different conditions using D.

fructosovorans and its hydrogenase negative mutant vary in absolute values although shares a general trend. It was significantly higher in all tests (except RD of chlorophenol) when compared to chemical palladium. The [Fe]-only hydrogenase negative mutant showed better catalytic activity compared to the wild type in most reactions. This provides evidence that, firstly, hydrogenase is not the only enzyme involved in Pd(II) reduction, and, secondly, it shows that the mutation has an additional effect, which causes variations in Bio-Pd catalytic activity.

AKNOWLEDGEMENTS We acknowledge the financial support of the EU BIO-CAT (Grant number GRD1-

2001-40424) and the BBSRC (Grant number E13817) and EPSRC (Grant NoGR/NO8445).


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REFERENCES 1. Platinum 2002 Interim Review. Johnson Matthey. November (2002) 28. 2. P. Crowson. Minerals Handbook. Macmillan Ltd., Surrey, 1982. 3. J.R. Lloyd, P. Yong, L.E. Macaskie. Appl. Environ. Microbiol., 64 (1998) 4608. 4. P.Yong, N.A. Rowson, J.P. Farr, I.R. Harris, L.E. Macaskie. Biotechnol. Bioeng., 80

(2002) 369. 5. I. Mikheenko, P. Mikheenko, C.N.W. Darlington, C.M. Muirhead, L.E. Macaskie. In

Biohydrometallurgy: Fundamentals, Technology and Sustainable Development Eds V.S.T. Ciminelli & O Garcia, Elsevier (2001) 525.

6. I.P. Mikheenko, L.E. Macaskie, P.M. Mikheenko, C.M. Muirhead. Pd. 19th General Conference of the EPS Condensed Matter Division held jointly with CMMP 2002 – Condensed Matter and Materials Physics. Brighton. Europhysics Conference Abstracts. 26A (2002) 93.

7. V.S Baxter-Plant, I.P. Mikheenko, L.E. Macaskie, Biodegradation (2002) in press. 8. R.Cammack, M. Frey, R. Robson (eds.), Hydrogen as a Fuel. Learning from Nature.

Taylor & Francis. London, New York, 2001. 9. L. Casalot, De Luca G, Dermoun Z, Rousset M, de Philip P. J. Bacteriol.,184 (2002)

853. 10. B.K. Pohorelic, J.K. Voordouw, E. Lojou, A. Dolla, J. Harder, G. Voordouw. J.

Bacteriol., 184 (2002) 679. 11. C. Aubert, M. Brugna, A. Dolla, M. Bruschi, M-T. Giudici-Orticoni. Biochim.

Biophys. Acta - Protein Structure & Molecular Enzymology, 1476 (2000) 85. 12. C. Wawer, G Muyzer. Appl. Environ. Microbiol., 61 (1995) 2203. 13. M. Rousset, L. Casalot, B. J. Rapp-Giles, Z. Dermoun, P. de Philip, J. P. Belaich, J.

D.Wall. Plasmid, 39 (1998) 114. 14. G.C. Dasages. Absorptiometriques des elements mineraux. Ed. Masson, Paris, 1978. 15. G.H. Jeffrey, J. Bassett, J. Mendham, R.C. Denny. Vogels textbook of quantitative

chemical analysis, Fifth Edition, Bath Press, Avon, UK, 1989.



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Removal of cobalt, strontium and caesium from aqueous solutions using native biofilm of Serratia sp. and biofilm

pre-coated with hydrogen uranyl phosphate

M. Paterson-Beedle and L.E. Macaskie

School of Biosciences, The University of Birmingham, Edgbaston, Birmingham, B15 2TT, UK

Abstract Heavy metals are removed by Citrobacter sp. (NCIMB 40259) now identified as a

Serratia sp. via the activity of an acid-type phosphatase enzyme, which liberates HPO42-

from a suitable organic phosphate donor with the stoichiometric precipitation of heavy metal cations (M2+) as insoluble MHPO4 at the cell surface. Previous studies have shown that it is possible to remove Ni2+ effectively via its intercalation into pre-formed or growing crystals of hydrogen uranyl phosphate (HUP), with the "guest" metal species intercalating within the HUP matrix. Also, the effective biomineralisation of HUP from uranium mine waters has been demonstrated. Nuclear wastes contain not only uranium but also fission products like 90Sr, 137Cs, and 60Co. We have shown that in the presence of uranyl ion and glycerol 2-phosphate the deposited HUP is able to remove "cold" surrogates of the fission products via intercalation using either continuous co-crystal growth or by using cells pre-coated with HUP as inorganic ion-exchangers. Using Serratia sp. biofilm immobilised onto polyurethane reticulated foam continuous removal of fission products surrogates was obtained by intercalation techniques, but not using the phosphatase-biomineralisation route alone, in the absence of the HUP host crystal.

1. INTRODUCTION Fission of 235U yields radioactive fission products such as the isotopes 60Co, 137Cs and

90Sr. Urgency is prompted by the nuclear industry to treat radionuclide-loaded liquid wastes. Biological methods for removal of metals from nuclear wastes can succeed where traditional physico-chemical treatments fail [1, 2]. A model for phosphatase-mediated uranyl ion accumulation, with exocellular deposition of polycrystalline (i.e., multitude of small crystals [3]) hydrogen uranyl phosphate (HUO2PO4.4H2O, HUP) that is identical to the structure of HUP prepared by inorganic crystal growth has been well published [4]. "Chemical" HUP is an established ion-exchange material, with the intralamellar mobile protons freely exchangeable for other ions like Na+, Ni2+ [5] and Co2+ [6, 7]. Previous studies have demonstrated two approaches for the bioremediation of Ni2+ from dilute aqueous flows using Serratia sp. cells [8-12]. The first approach was based on a two step process: (i) enzymatically-mediated generation of a polycrystalline ‘host lattice’ (priming layer), i.e. HUP bound to the surface of the Serratia sp. cells and (ii) cation-exchange intercalation of Ni2+ into the interlamellar spaces of HUP [8-12]. The second approach [8] was to co-challenge immobilised Serratia cells in a packed-bed reactor with a Ni2+ and


1156

UO22+ solution in the presence of glycerol-2-phosphate (phosphate donor for phosphate

release and metal precipitation) giving a sustained removal of both metals. Since Co and Ni are in the same group of the periodic table of the elements they should behave similarly. Therefore, the purpose of this study was to evaluate the methods described above using Serratia sp. biofilm immobilised onto polyurethane foam for the removal of Co2+ from solution. Previous studies showed the uranyl phosphate precipitate to be, more likely, NaUO2PO4.4H2O [13], therefore, Cs+ should be similarly removed as CsUO2PO4 and, in the presence of both fission products surrogates, bio-driven formation of mixed metal uranyl phosphate is distinctly feasible. Previous studies showed removal of Sr2+ as the phosphate precipitate [14] and removal of Sr2+ as a co-crystal was also explored.


2.1 Support, organism and biofilm production Citrobacter sp. (NCIMB 40259) now identified as Serratia sp. [15] was grown as

biofilm onto polyurethane reticulated foam Filtren TM30 (supplied by Recticel, Belgium) in an air-lift fermenter [16] and the phosphatase specific activity of the cells from the fermenter outflow was determined as described previously [16]; the steady state specific activity was ~ 3000 units (nmol p-nitrophenol released from p-nitrophenyl phosphate min-

1 mg protein-1).

2.2 Preparation of packed-bed reactor systems for metal bioaccumulation Cubes of polyurethane foam (88, 125 mm3) coated with Serratia sp. biofilm were

packed in a cylindrical glass column (length 9.0 cm and internal diameter 1.5 cm) of working volume of ca. 13 ml. The total amount of protein immobilised onto the foam was ca. 95 mg per reactor, measured by the bicinchoninic acid/CuSO4 method (Sigma protein test kit TPRO-562) using bovine serum albumin as standard [17]. Considering that the amount of protein is ca. 50% of the dry weight of biomass [unpublished], the total biomass per reactor was, therefore, ca. 190 mg. The reactors were challenged with an upward flow (ca. 10 ml h-1) via an external peristaltic pump (Watson-Marlow). All tests were done at ambient temperature. Challenge solutions comprised sodium citrate buffer (2 mM), pH 6.0 and sodium glycerol-2-phosphate (5 mM) supplemented with Co(NO3)2.6H2O (1 mM), CsCl (1 mM), Sr(NO3)2 (1 mM) or UO2(NO3)2.6H2O (1 mM) as specified in individual experiments.

2.3 Metal biosorption by packed-bed reactors Reactor systems were prepared, similar to those described in section 2.2, and

challenged with metal (1 mM) in the presence of citrate buffer (2 mM), pH 6.0, but in the absence of glycerol-2-phosphate.

2.4 Co-crystallisation of metals by packed-bed reactors not previously primed with HUP Reactor systems were prepared, similar to those described in section 2.2, and

challenged with solutions comprising UO2(NO3)2.6H2O (1 mM), sodium citrate buffer (2 mM), pH 6.0 and sodium glycerol-2-phosphate (5 mM) supplemented with Co(NO3)2.6H2O (1 mM), CsCl (1 mM) or Sr(NO3)2 (1 mM) as specified in individual experiments.


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2.5 Use of reactor packed-bed systems previously primed with HUP Reactor systems were "primed" (deposition of HUP onto the biomass) using a

solution of UO2(NO3)2.6H2O (1 mM), sodium glycerol-2-phosphate (5 mM) and sodium citrate buffer (2 mM), pH 6.0. First step: four reactors were primed for 24.3 h at a flow rate of 24 ml.h-1 giving a removal of ca. 65% of the input uranyl ion (total uranium loaded was ca 139 mg). Second step: two reactors were challenged with a solution comprising Co(NO3)2

.6H2O (1 mM) in the presence of glycerol-2-phosphate (5 mM), pH 6.0 and the other two were challenged with a solution comprising Co(NO3)2

.6H2O (1 mM) in the presence of glycerol-2-phosphate (5 mM) and sodium citrate buffer (2 mM), pH 6.0.

2.6 Spectrophotometric analysis of Co2+ The Co2+ contents of reactor outflows were estimated using the method described by

Onishi [18], with a slight modification. The sample or standard containing Co2+ (30 µl, 1-10 µg) was transferred to a test tube, citric acid solution (0.2 M, 250 µl) and phosphate-boric acid buffer (6.2 g of boric acid, 35.6 g of disodium phosphate dehydrated, and 500 ml of 1 M sodium hydroxide in a total volume of 1 l, 300 µl) were added. The pH of the solution should be close to 8.0. Nitroso-R salt (supplied by Fluka, UK) solution (0.2%, 125 µl) was added while stirring. The test tubes containing the samples were covered, transferred to a water bath at 100°C and left for 1 min. Concentrated HNO3 (250 µl) was added and the samples left for a further 1 min. Samples were cooled in the dark and then deionised water was added (325 µl). The transmittance of the solution was measured at 420 nm.

2.7 Spectrophotometric analyses of Sr2+ and/or UO22+

A method was developed, similar to that described by Yong et al. [19], for the simultaneous determination of uranium and strontium in mixed solutions using arsenazo III. Michaylova and Kouleva [20] showed that the complex formation of strontium with arsenazo III is maximum at pH 9-10. To each test tube containing 30 µl of a target metal (or metal mixture) was added 1.97 ml of one of the two solutions, as appropriate to the metal under test (0.1 M HCl for uranium and 0.1 M borate buffer, pH 9.0 for strontium). Metal was visualised by the addition of 0.1 ml of 0.15% (w/v) arsenazo III, with estimation of the blue-violet complex at 652 nm and 649 nm for uranium and strontium, respectively. The blue-violet complex develops in 25 min with strontium [20].

2.8 Spectrophotometric analysis of Cs+ The Cs+ contents of reactor outflows were estimated using the method described by

Huey and Hargis [21] with modifications. Sample or standard containing Cs+ (60 µl, 1-50 µg) was transferred to an Eppendorf tube (1.5 ml), perchloric acid solution (6 M, 84 µl) was added and diluted with deionised water (192 µl). Phosphomolybdic acid (supplied by Riedel-deHaën) solution (14%, w/v, 60 µl) was added, mixed and allowed to stand for 30 min. The suspension was centrifuged at 16,060 g for 30 min, the supernatant discarded using a glass Pasteur pipette and the precipitate washed with perchloric acid (1.2 M, 300 µl). The suspension was centrifuged at 16,060 g for 30 min, the supernatant discarded and the precipitate was dissolved in borate buffer pH 9.0 (0.05 M, 360 µl) during 30 min. Borate buffer (840 µl) was added to give a final volume of 1.2 ml. The absorbance was measured at 226 nm.


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3.1 Biosorption and bioaccumulation of Co2+ by Serratia sp. biofilm The first experiment was designed to test the background adsorption/biosorption of

Co2+ from aqueous flows by polyurethane foam coated with Serratia sp. biofilm (Fig. 1). Approximately 11% of the Co2+ was removed initially (up to 21 ml, or ca. 2 column volumes), decreasing rapidly thereafter. Biosorption of Co2+ by the biomass was negligible (less than 1% of the biomass dry weight). Bioaccumulation of Co2+ by Serratia sp. in the presence of glycerol 2-phosphate was also negligible (Fig. 1) even though excess phosphate was produced (not shown). These results are similar to those obtained using immobilised Serratia sp. cells for the removal of Ni2+ [8, 11], and show that phosphatase-mediated metal bioprecipitation is not appropriate for these metals.

Figure 1. Co2+ removal by reactors containing Serratia sp. biofilm immobilised onto polyurethane foam and challenged with: , Co(NO3)2

.6H2O (1 mM) in the presence of citrate buffer (2 mM), pH 6.0 (biosorption); , Co(NO3)2

.6H2O (1 mM) in the presence of glycerol-2-phosphate (5 mM) and citrate buffer (2 mM), pH 6.0 (bioaccumulation)

3.2 Co-crystallisation of Co2+ and HUP Previous experiments [8] showed that Serratia sp. cells, when challenged with a

solution of uranyl nitrate (0.5 mM) and nickel nitrate (0.5 mM) in the presence of glycerol-2-phosphate and citrate buffer, accumulated both metals. The flow rate was set to give a removal of UO2

2+ of 56.3% and the corresponding removal of Ni2+ from solution was maintained at 27.3% (during 42 h, i.e. 25 column volumes). The proportion of metal removed suggest the formation of Ni(UO2PO4)2

.7H2O; the identity of the material accumulated by the cells was confirmed by X-ray powder diffraction [8]. We tested cells immobilised onto polyurethane foam challenged with a solution of UO2(NO3)2.6H2O (1 mM) and Co(NO3)2

.6H2O (1 mM) in the presence of glycerol-2-phosphate with a flow rate set at 10 ml.h-1. The removal of UO2

2+ was ca. 99% but the corresponding removal of Co2+ was less than 10% (Fig. 2).

3.3 Uptake of Co2+ onto/into HUP previously accumulated onto the biomass In this study the reactors were primed with HUP as described in section 2.5 which

resulted in ca. 90 mg uranium deposited per reactor (ca. 47% of bacterial dry weight). The initial challenge with Co2+ (Fig. 3) resulted in a slowly decreasing removal of Co2+ over 226 ml, with no sharp breakthrough after saturation of the HUP host. This pattern is similar to that obtained using Serratia biofilm immobilised onto ceramic support, primed


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with HUP, and initially challenged with Ni2+ solution [11]. The theoretical capacity of a reactor containing 90 mg uranium (in the form of HUP) would be 11 mg of cobalt at a molar ratio of Co:U of 1:2, i.e. for the formation of Co(UO2PO4)2. A loss of Co-removing capacity after 226 ml corresponds to ca. 10 mg of deposited cobalt which, therefore, is attributable to column saturation. As expected the presence of citrate in the challenge solution reduced the amount of Co2+ deposited (Fig. 3), since the metal would be presented to the cells as the citrate complex; citrate was shown previously to remove Ni2+ from its position within the HUP host crystal. [12].

Figure 2. Metal removal by reactors containing Serratia sp. biofilm immobilised onto polyurethane foam and co-challenged with a solution of Co(NO3)2

.6H2O (1 mM) and UO2(NO3)2

.6H2O (1 mM) in the presence of glycerol-2-phosphate (5 mM) and citrate buffer (2 mM), pH 6.0. , UO2

2+ removed. , Co2+ removed

Figure. 3. Co2+ removal by reactors containing Serratia sp. biofilm immobilised onto polyurethane foam, primed with HUP and challenged with: , a solution of Co(NO3)2.6H2O (1 mM) in the presence of glycerol-2-phosphate (5 mM), pH 6.0; , a solution of Co(NO3)2.6H2O (1 mM) in the presence of glycerol-2-phosphate (5 mM) and citrate buffer (2 mM), pH 6.0.

3.4 Co-crystallisation of Cs+ and HUP Serratia sp. cells immobilised onto polyurethane foam were co-challenged with

UO2(NO3)2.6H2O (1 mM) and CsCl (1 mM) in the presence of glycerol-2-phosphate at a

flow rate of ca. 10 ml.h-1. The removal of UO22+ was ca. 96% and the corresponding

removal of Cs+ was ca. 58% (Fig. 4). The removal of Cs+ using the phosphatase-biomineralisation route alone, in the absence of the HUP host crystal, was less than 20%. Thus it was concluded that although co-crystallization did not promote removal of Co2+ (above) this approach has potential for the removal of Cs+. It should be noted that since glycerol 2-phosphate was provided as the sodium salt (5 mM) Na+ would be present to a five-fold excess over Cs+ and in this respect it can be suggested that co-crystal formation


1160

favours formation of the CsUO2PO4 precipitate. However, it is not easy to distinguish between HUO2PO4 and NaUO2PO4 by X-ray powder diffraction because they both have very similar crystal lattice [13].

Figure 4. Metal removal by reactors containing Serratia sp. biofilm immobilised onto polyurethane foam and co-challenged with a solution of CsCl (1 mM) and UO2(NO3)2.6H2O (1 mM) in the presence of glycerol-2-phosphate (5 mM) and citrate buffer (2 mM), pH 6.0. , UO2

2+ removed. , Cs+ removed.

3.5 Co-crystallisation of Sr2+ and HUP Serratia sp. cells immobilised onto polyurethane foam were co-challenged with

UO2(NO3)2.6H2O (1 mM) and Sr(NO3)2 (1 mM) in the presence of glycerol-2-phosphate at

a flow rate of ca. 10 ml.h-1. The steady-state removal of UO22+ was more than 92% and the

corresponding removal of Sr2+ was in the range of 50-56% (Fig. 5). The removal of Sr2+ using the phosphatase-biomineralisation route alone, in the absence of the HUP host crystal, was negligible. The molar ratio accumulated was in accordance with the deposition of Sr(UO2PO4)2.

Figure 5. Metal removal by reactors containing Serratia sp. biofilm immobilised onto polyurethane foam and co-challenged with a solution of Sr(NO3)2 (1 mM) and UO2(NO3)2

.6H2O (1 mM) in the presence of glycerol-2-phosphate (5 mM) and citrate buffer (2 mM), pH 6.0. , UO2

2+ removed. , Sr2+ removed

4. CONCLUSIONS We have shown that using the continuous co-challenge system, i.e. in the presence of

uranyl ion and glycerol 2-phosphate, the deposited HUP is able to promote removal of ca. 58% of Cs+ and 50-56% of Sr2+ (during 53 column volumes) but was not able to remove the Co2+ (less than 10%). However, using cells pre-coated with HUP as inorganic ion-exchangers it was possible to remove the Co2+ until the column reached saturation corresponding to a molar ratio of U:Co of ~2:1 in accordance with a crystal of


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Co(UO2PO4)2. Further work to be reported will confirm the identity of the crystals formed, using X ray powder diffraction analysis and EXAFS. However, the preliminary data reported here show that removal of Co2+, Sr2+ and Cs+ as, respectively, Co(UO2PO4)2, Sr(UO2PO4)2 and CsUO2PO4 is feasible since no removal of the surrogate fission products occurred in the absence of the uranyl ion. Such uranyl phosphate-driven co-deposited or intercalation is entirely feasible for the treatment of real nuclear wastes since the fission products residues often occur with an excess of residual uranium [1, 2] and, furthermore, the metal precipitating phosphatase is highly radioresistent [22].

ACKNOWLEDGMENTS This work is supported by the Biotechnology and Biological Sciences Research

Council. The authors thank Recticel (Belgium) for the polyurethane reticulated foam.

REFERENCES 1. L. E. Macaskie, CRC Crit. Rev. Biotechnol. 11 (1991) 41-112. 2. J. R. Lloyd and L. E. Macaskie, In: Environmental Microbe-Metal Interactions, (D. R.

Lovely, ed.), ASM Press, Washington, (2000) 277-327. 3. B. K. Vainshtein, Fundamentals of Crystals, Modern Crystallography, Vol. 1, 2nd edn.,

Springer, Berlin, 1994. 4. L. E. Macaskie, R. M. Empson, A. K. Cheetham, C. P. Grey and J. Skarnulis, Science

257 (1992) 782-784. 5. A. Clearfield, Chem. Rev. 88 (1988) 125-148. 6. R. Pozas-Tormo, L. Moreno-Real, M. Martinez-Lara and S. Bruque-Gamez, Can. J.

Chem. 64 (1986) 30-34. 7. R. Pozas-Tormo, S. Bruque-Gamez, M. Martinez-Lara and L. Moreno-Real, Can. J.

Chem. 66 (1988) 2849-2854. 8. K. M. Bonthrone, G. Basnakova, F. Lin and L. E. Macaskie, Nature Biotechnol. 14

(1996) 635-638. 9. G. Basnakova and L. E. Macaskie, Biotechnol. Bioeng. 54 (1997) 319-328. 10. G. Basnakova, A. J. Spencer, E. Palsgard, G. Grime and L. E. Macaskie, Environ. Sci.

Technol., 32 (1998) 760-765. 11. G. Basnakova, J. A. Finlay and L. E. Macaskie, Biotechnol. Lett., 20 (1998) 949-952. 12. G. Basnakova and L. E. Macaskie, Biotechnol. Lett., 23 (2001) 67-70. 13. P. Yong and L. Macaskie, J. Chem. Tech. Biotechnol. 63 (1995) 101-108. 14. L. E. Macaskie and A. C. R. Dean, Biotechnol. Lett., 7 (1985) 627-630. 15. P. Pattanapipitpaisal, A. N. Mabbett, J. A. Finlay, A. J. Beswick, M. Paterson-Beedle,

A. Essa, J. Wright, M. R. Tolley, U. Badar, N. Ahmed, J. L. Hobman, N. L. Brown and L. E. Macaskie, Environ. Technol., 23(7) (2002) 731-45.

16. J. A. Finlay, V. J. M. Allan, A. Conner, M.E. Callow, G. Basnakova and L. E. Macaskie, Biotechnol. Bioeng., 63 (1999) 87-97.

17. K. P. Nott, M. Paterson-Beedle, L. E. Macaskie and L. D. Hall, Biotechnol. Lett., 23 (2001) 1749-1757.

18. H. Onishi, Photometric Determination of Traces of Metals, Part IIA: Individual Metals, Aluminum to Lithium, 4th edn., John Wiley, New York (1986) 454-459.

19. P. Yong, H. Eccles and L. E. Macaskie, Anal. Chim. Acta, 329 (1996) 173-179. 20. V. Michaylova and N. Kouleva, Talanta, 21 (1974) 523-532. 21. F. Huey and L. G. Hargis, Anal. Chem., 39(1) (1967) 125-127. 22. B. C. Jeong, Studies on the atypical phosphatase of a heavy metal accumulating

Citrobacter sp., D. Phil. Thesis, University of Oxford, UK (1992).



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Removal of soluble manganese from mine waters using a fixed-bed column bioreactor

D. Barrie Johnson, Helen Miller, Sandra Ukermann and Kevin B. Hallberg


Abstract Acid mine drainage waters vary considerably in the range and concentration of heavy

metals they contain. Besides iron, aluminum and manganese are frequently present at elevated concentrations in waters draining both coal and metal mines. Passive treatment systems (aerobic wetlands and compost bioreactors) are designed to remove iron by biologically induced oxidation/precipitation, and aluminum by precipitation as Al(OH)3 as a result of biologically-generated alkalinity. Manganese, however, is problematic as it does not readily form sulfidic minerals (in compost bioreactors) and requires elevated pH (>8) for oxidation of Mn(II) to insoluble Mn(IV). As a result, manganese removal is often less effective than iron and aluminum removal, such as at the pilot passive plant treating water draining the former Wheal Jane tin mine in Cornwall, U.K. We have sought to devise a novel microbiological approach for effective removal of manganese from mine waters at pH 5-7. Ferromanganese nodules (about 2 cm diameter) were collected from an abandoned mine adit in the Gwydyr forest, north Wales, and used to inoculate a fixed bed bioreactor (working volume ca. 700 ml). Pumice-like porous beads, made from inert re-cycled glass, were used as carrier materials for microbial biofilms. Following colonization of the beads, the aerated reactor was tested for removal of soluble manganese in synthetic and actual mine waters, using a continuous plug-flow mode of operation. Data from preliminary experiments show that the bioreactor is highly efficient at catalyzing the removal of manganese from the mine waters via oxidation of soluble Mn (II) and precipitation of the resultant Mn (IV) in the bioreactor. Such an approach appears to be a suitable alternative to current bioremediation strategies employed for manganese removal from mine waters.

Keywords: acid mine drainage; bioreactors; bioremediation; manganese

1. INTRODUCTION Soluble manganese (Mn (II)) is often found in considerably greater amounts in AMD

than in unpolluted streams and groundwater [1]. Even though there are uncertainties regarding the toxicity of manganese, the removal of this metal from surface- and groundwaters is desirable for several reasons. As with iron and aluminum, manganese also contributes to the total mineral acidity of mine waters when it is oxidized. Additionally, the presence of manganese in water sources for human consumption is undesirable because it can impart a metallic taste to water, it will stain laundry and water fixtures and, as it precipitates readily as Mn (IV), manganese tends to block water distribution


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networks. Due to these concerns, the U.S. Environmental Protection Agency (EPA) has set a secondary maximum contaminant level for Mn of 0.05 mg/l. The EPA has also established guidelines limiting the concentration of Mn in acidic waters discharging from mines at maximum of 4 mg/l, as long as average discharges for a 30-day period to not exceed 2 mg/l.

To address such concerns, and to meet the potentially more stringent limits that are being discussed in Europe and in the U.S., effective means of removal of Mn from mine waters are required. Although manganese readily precipitates as Mn(IV), little oxidation of Mn(II) occurs in solutions below pH 8, in contrast to iron, and the kinetics for manganese oxidation is slow, relative to that of ferrous iron. In addition, biological oxidation of Mn(II) does not proceed rapidly in the presence of Fe(II), and thus it is not removed significantly in wetlands where the concentration of iron exceeds 1 mg/l [2]. The lack of affinity of manganese for sulfide prohibits any significant removal as a sulfide in compost wetlands and bioreactors. Therefore, alternative approaches are needed for the removal of manganese from mine drainage waters.

The treatment systems currently used are typically applied at the end of any treatment process scheme, to ensure the all iron has been removed first. An abiotic Mn-removal system consisting of columns filled with limestone (to significantly raise the pH for Mn oxidation and subsequent precipitation) has been described [3]. Biological approaches that have been proposed for removing manganese from solution, by causing solution pH to rise above 8 via oxygenic photosynthesis, include columns of immobilized cyanobacteria [4] and small pools filled with rocks that have been colonized by algae [5]. The latter approach has been applied at the Wheal Jane passive treatment plant. This study site was built in 1993 following a catastrophic spill of acid mine drainage from the Wheal Jane tin mine in Cornwall, England. However, the algal ponds have proved highly ineffective at removing soluble manganese (as described below) and an alternative approach, using a fixed-bed column bioreactor, has been initiated as an alternative strategy, which would be equally applicable for treating other mine waters.

2. MANGANESE REMOVAL AT THE WHEAL JANE PASSIVE TREATMENT PLANT A composite, pilot-scale passive treatment plant ("PPTP") was established at the

former Wheal Jane tin mine in Cornwall, England in 1995, as a large-scale experimental facility to examine the efficacy of using constructed "wetlands" to remediate acidic, metal-rich minewaters. A more complete description of the design and operation of the treatment system can be found elsewhere [6]. In brief, a small fraction (<5%) of the acid (pH ~ 3.5) mine drainage (AMD) water flows into the pilot passive treatment plant, where it is fed into three similar composite passive systems (Figure 1). The main variations are the pre-treatment that the AMD is subjected to ahead of entering the "wetlands". These are: (i) addition of lime (CaO) to adjust the AMD to a pre-determined pH value (the "lime-dosed" system); (ii) passage through an anaerobic (manure-based) cell to remove dissolved oxygen and then through an anoxic limestone drain (to passively add alkalinity; the "ALD" system); (iii) no pretreatment (the "lime-free" system). In each system, AMD then flows (by gravity) through a series of five “aerobic cells” (which are considerably larger in the "lime-free" system), then through a single "anaerobic cell" (containing sawdust and cow manure; the "compost bioreactor") and finally through a series of ten smaller "rock filters" (algal ponds). The key roles of the components in the passive systems are:

(a) aerobic cells: oxidation and precipitation of iron, and co-precipitation of arsenic;


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(b) anaerobic cell: sulfidogenesis and alkalinity genesis, causing precipitation of chalcophilic metals as sulfides (e.g. copper, zinc and cadmium) and other metals as hydroxides (e.g. aluminum);

(c) rock filters: (i) oxygenic photosynthesis causing aeration and further increase in pH and precipitation of manganese (as carbonate and/or oxide) and (ii) polishing of processed AMD (e.g. removal of organic carbon compounds draining the compost bioreactors).

Outflow

Aer

obic

Cel

l 1

Inflowing AMD

Aer

obic

Cel

l 3

Aer

obic

Cel

l 5

Com

post

Bio

reac

tor

Roc

k Fi

lters

(10)

MeS removal

Mn removal Fe2+ oxidation and Fe/As removal

Figure 1. Schematic of the pilot passive treatment plant for remediating AMD at the former Wheal Jane tin mine, Cornwall, England The performance of the PPTP has been monitored by a group of geochemists and microbiologists since 1998. In general, the aerobic cells have proved highly effective in fulfilling their roles, but in only one of the three systems (the "lime-free" system) has the performance of the compost bioreactors and algal ponds been satisfactory [7]. Figure 2 shows the concentrations of manganese in waters sampled throughout the three systems in the PPTP in April 2002, and is representative of the general trends that have been observed. Data in Figure 2 show that changes in soluble manganese do occur within the aerobic cells (which are mostly highly acidic, pH <3, due to ferric iron hydrolysis); this is, in part, due to dilution of the AMD by rainwater. However, significant concentrations (~2-5 mg/l) are present in the water draining the compost bioreactors and, as shown in Figure 2, is removed effectively only in the "lime-free" system. Indeed there was evidence of re-mobilization of manganese in the rock filters of the "lime-dosed" system on this sampling occasion.

A major reason for the contrasting trends observed in Mn geochemistry found within the algal ponds is that the necessary elevation in pH resulting from oxygenic photosynthesis only occurred in the "lime free" system, as illustrated in Figure 3. The discrepancy in Mn removal is due, in the main, to the variation in the performance of the compost bioreactors in each of the three systems. Water draining these bioreactors in the "ALD" and "lime-free" systems contains higher concentrations of both ferrous iron and sulfide. Microbiological (and abiotic) oxidation of these species in the algal ponds causes the pH of the water to decline, which is counter-productive to the overall objective [7]. As the treatment plant is built in a temperate climate that receives considerable rainfall, photosynthetic activity appears to be limited such that the rock filters are unable to overcome the additional acid burden from the poorly performing compost bioreactors.


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Figure 2. Manganese concentrations in various components of the pilot passive treatment plant at the former Wheal Jane tin mine; sampling date April 2002

Figure 3. Manganese (bars) and pH (lines) in the compost bioreactor influent (CB in) and effluent (CB out) and in each of the 10 rock filter pools (RF1-RF10). The samples were taken in April, 2002, from the ALD (A) and LF (B) systems. The straight lines at the bottom of each graph represent the lower limit of manganese detection, which was 0.25 mg/l


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3. REMOVAL OF SOLUBLE MANGANESE BY FRESHWATER FE-MN NODULES To overcome the limitations of the rock filters, we have proposed an alternative

treatment strategy based on bioreactors containing immobilized manganese-oxidizing bacteria. Ferromanganese nodules were collected from a trial mine adit located in the Gwydyr forest, north Wales. The irregular, spherical nodules, ca. 1-2 cm in diameter, were found in a pool within the adit (pH 6.9, soluble [Mn] 30 mg/l, conductivity 149 µS/cm). These were transferred to the laboratory and stored at 4°C. Two of the nodules were placed into replicate 250 ml conical flasks containing 100 ml of filter-sterilized "MMS" medium [8], which had been developed to promote the growth of manganese oxidizing/depositing bacteria. MMS medium contains sodium pyruvate (2.4 g/l) and basal salts (pH 7); 50 µl of a 100 mM manganese sulfate solution (filter-sterilized) was added just prior to inoculation with the nodules, and the flasks were incubated, shaken (100 rpm) at 20°C. Samples were withdrawn at regular intervals and concentrations of soluble Mn analyzed colorimetrically by complexing with formaldoxime (Hydrocheck, Cambridge, U.K.). Results, given in Figure 4, show that concentrations of soluble manganese declined steadily during incubation; this corresponded with the appearance of brown-black precipitates (and associated bacteria) within cultures.

Figure 4. Removal of soluble manganese in shake flask cultures containing freshwater ferro-manganese nodules. The error bars (occluded in most cases by symbols) show the range of soluble [Mn] at each sampling occasion

4. COMMISIONING AND PERFORMANCE OF A FIXED-BED BIOREACTOR A fixed-bed bioreactor for removing soluble manganese from mine waters was set up

by mixing freshwater ferromanganese nodules (ca. 10%, v/v) with porous, acid-washed glass beads (8-16 mm) manufactured by Poraver GmbH, Germany [9]. The nodules/beads mixture was packed (to a depth of 14.5 cm) into a 25 by 10 cm perspex column, between 3.5 cm layers of acid-washed pea gravel (Figure 5a). The base and cap of the column contained ports for aeration (applied at ca. 2 l/min), medium circulation and sampling. The column was filled to the top of the upper gravel layer surface with MMS medium, which was then re-circulated through the column using a peristaltic pump until the soluble


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manganese had declined to <10% of the initial concentration, at which point the column was drained and fresh medium introduced. This sequence was repeated several times, and the column was observed to accumulate increasing amounts of brown-black (Mn (IV)) precipitates (Figure 5b). The external plumbing on the bioreactor was then re-configured to allow the system to be tested in continuous plug-flow mode.

Representative data showing removal of soluble manganese from MMS medium with the fixed-bed bioreactor run in recirculating and plug-flow modes are shown in Figure 6. In both cases, >90% of soluble Mn was removed within 4 hours. When run in plug-flow mode, 50% of soluble manganese was removed with a flow rate of 400 ml/h, and 68% when the flow rate was lowered to 195 ml/h.

Figure 5. Fixed-bed bioreactor for removing soluble Mn from solution: (a) after initial construction; (b) several weeks after continuous operation

Figure 6. Performance of the fixed-bed bioreactor in removing soluble manganese from MMS medium when run in recirculating ( ) and plug-flow ( ) modes


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5. DISCUSSION Manganese is the second most abundant heavy metal in the lithosphere, occurring at

between 0.002 and 10% (w/w) in terrestrial ecosystems, ca. 0.2 µg/l in marine waters, and an average of 8 µg/l in fresh waters [10]. As with many other metals, greatly enhanced concentrations of soluble manganese can occur in waters draining coal and metal mines. The major oxidation states of manganese are +2 and +4: Mn (II) can occur as free ions in solution (Mn2+, and Mn(OH)+ in circum-neutral pH waters) or complexed with a variety of organic and inorganic ligands. In contrast, Mn (IV) forms very insoluble oxide phases; MnO2 has a point of zero charge at pH 2.8 which, together with its large specific surface area, gives rise to a marked capacity to adsorb metal cations, including Mn (II).

Thermodynamically, Mn (II) should oxidize spontaneously to Mn (IV) in aerated, neutral pH waters, but the activation energy required is relatively high, and this greatly slows down the process, causing Mn (II) to be far more stable in non-acidic waters than is ferrous iron. In contrast, the adsorption of Mn (II) onto MnO2 is much more rapid. The "activation energy barrier" for Mn oxidation can be overcome biologically. Many microorganisms are known to be able to catalyze the oxidation of Mn (II), and at least some of these are known utilize the energy available from the reaction [10]. Manganese (II) may also form a highly insoluble carbonate (MnCO3; rhodocrosite). Since, however, the concentration of CO3

2- in solution is dictated by pH, formation of this mineral is only significant in neutral-alkaline water bodies.

The existence of marine "manganese nodules" has been known for some time, and these are mined as an ore of the metal. Iron/manganese-rich nodules may also occur in freshwaters, though these tend to be much smaller than their marine equivalents. Stein et al. [11] analyzed prokaryotic populations associated with freshwater ferromanganese micronodules (mm scale) and sediments by extracting nucleic acids, amplification of 16S rRNA genes and phylogenetic analysis of cloned products. Two groups of known Mn (II)-oxidizing genera were identified, Leptothrix and Hyphomicrobium, as well as a group related to metal-reducing bacteria.

To date, the microorganisms associated with the freshwater ferromanganese nodules used in the present study have not been subjected to rigorous analysis, though Mn(IV)-depositing bacteria have been isolated on solid media inoculated with crushed nodules and bioreactor liquor. The increasingly-efficient performance of the bioreactor (in terms of removing soluble Mn(II)) as Mn(IV) deposits accumulate may be due both to the activities of planktonic-phase Mn(II)-oxidizers and to the abiotic uptake of Mn(II) onto MnO2 precipitates, and subsequent oxidation of the metal by sessile bacteria. What is clear is that a simple fixed-bed bioreactor containing Mn(II)-oxidizing/depositing microorganisms can be highly effective in catalyzing the removal of the metal from a synthetic, metal-rich waste water. So far, limited tests of the technology for bioremediating actual Mn-rich mine waters has shown considerable promise (data not shown). Fixed-bed bioreactors for oxidizing and precipitating ferrous iron from AMD have been described by a number of research groups [e.g. 12].

A major advantage offered by metal-removing bioreactors is that they occupy a far smaller "footprint" than, for example, the rock filters/algal ponds at the Wheal Jane passive treatment plant. Land availability frequently restricts the application of passive mine water treatment systems, and low-maintenance biological systems are attractive alternatives in such circumstances.


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ACKNOWLEDGEMENTS We wish to thank DTI (ref. # BTL/20/71), MIRO (U.K.) and the Environment

Agency (U.K.) for financial support. We also acknowledge the contributions made by our colleagues at the University of Reading, Imperial College, London, the Centre for Ecology and Hydrology, the Camborne School of Mines and the Knight Piésold Consulting Group (U.K.) to the "Wheal Jane Project".

REFERENCES 1. D. Banks, P. L. Younger, R.-T. Arnesen, E. R. Iversen and S. B. Banks, Environ. Geol.

32 (1997) 157. 2. B. Nairn and R. S. Hedin, Constructed wetlands for water quality improvement, G. A.

Moshiri (eds.), Lewis Publishers, Boca Raton, Fl, (1993) 187. 3. F. C. Thornton, Ecol. Eng. 4 (1995) 11. 4. J. Bender, J. P. Gould, Y. Vatcharapijarn, J. S. Young and P. Phillips, Water Environ.

Res. 66 (1994) 679. 5. P. Phillips, J. Bender, R. Simms, S. Rodriguezeaton and C. Britt, Water Sci. Technol.

31 (1995) 161. 6. Q. U. I. Hamilton, H. M. Lamb, C. Hallett and J. A. Proctor, J. Chart. Inst. Water.

Environ. Manage. 13 (1999) 93. 7. K. B. Hallberg and D. B. Johnson, Mine Water Treatment:A Decade of Progress, C. A.

Nuttall (eds.), Newcastle upon Tyne, U.K., (2002) 143. 8. Y. M. Nelson, L. W. Lion, W. C. Ghiorse and M. L. Shuler, Appl. Environ. Microbiol.

65 (1999) 175. 9. Å. Kolmert and D. B. Johnson, J. Chem. Technol. Biotechnol. 76 (2001) 836. 10. H. L. Ehrlich, Geomicrobiology, 3rd edition. Marcel Dekker, New York (1996). 11. L. Y. Stein, M. T. La Duc, T. J. Grundl and K. H. Nealson, Environ. Microbiol. 3

(2001) 10. 12. A. Mazuelos, F. Carranza, I. Palencia and R. Romero, Hydrometallurgy 58 (2000) 269.



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Sulfane sulfur of persulfides is the actual substrate of the sulfur-oxidizing enzymes from Acidithiobacillus and

Acidiphilium spp.

T. Rohwerder∗ and W. Sand

Department of Microbiology, Institute for General Botany, University of Hamburg, Ohnhorststr. 18, D-22609 Hamburg, Germany

Tel/Fax: +49 40 428 16 423, e-mail: [email protected]

Abstract The sulfur chemistry of sulfur dioxygenase (EC 1.13.11.18) was analyzed in various

strains of the genera Acidithiobacillus and Acidiphilium. Cell-free extracts oxidized elemental sulfur well, however only in the presence of glutathione (GSH). Sulfur had to be merged with GSH to form the organic persulfide or higher homologues (GSnH, n > 1) prior to the enzymic oxidation to sulfite. The latter reacted non-enzymically further either to sulfate, thiosulfate, or glutathione S-sulfonate (GSSO3

-). In contrast, oxidation of hydrogen sulfide by the dioxygenase system did not require the reduced but the oxidized form of glutathione (GSSG). The latter formed with sulfide non-enzymically GSSH prior to enzymic oxidation. Persulfide species as sulfur donors could not be replaced by other sulfane-sulfur-containing compounds such as thiosulfate, polythionates, aryl thiosulfonates, or bisorganyl-polysulfanes. In conclusion, persulfides and higher monoorganylpolysulfanes are exclusively the substrates for the sulfur dioxygenases of meso-acidophilic sulfur bacteria. On the basis of these findings a biochemical model for both elemental sulfur and sulfide oxidation in Acidithiobacillus and Acidiphilium spp. is proposed.

Keywords: sulfur dioxygenase, sulfur oxidation, sulfide oxidation, Acidithiobacillus, Acidiphilium

1. INTRODUCTION For the oxidation of elemental sulfur a conclusive biochemical pathway has not yet

been identified. Diverse enzymic systems like a reverse sulfite reductase [1,2], a thiosulfate-oxidizing enzyme complex [3,4], and dioxygenases or oxygenase-reductases [5,6,7,8] are considered to be involved. The meso- and moderately thermophilic leaching bacteria seem to have a pathway depending on low-molecular thiols. For in vitro assays glutathione (GSH) is the preferentially used thiol compound (reviewed in [9]). Due to its capability to incorporate molecular oxygen into the sulfur substrate and to produce sulfite as the primary reaction product the system has been named sulfur dioxygenase (EC

∗ Present address: UFZ Center for Environmental Research Leipzig-Halle, Permoserstr. 15, D-04318 Leipzig, Germany


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1.13.11.18; [5]). This GSH-dependent activity has been detected in Acidithiobacillus thiooxidans and Acidithiobacillus ferrooxidans [10,11,12] as well as in Sulfobacillus thermosulfidooxidans [13]. Although the known sulfur dioxygenases are very similar in molecular size and cofactor content, the role of GSH is under dispute. The work of Suzuki [5,10] and Silver & Lundgren [11] suggest a catalytic role for GSH, where organic polysulfanes are formed and sulfur is oxidized at the zero valence state to sulfite (equation 1). In contrast, Sugio and coworkers [12,14] described an enzyme system that probably used dissolved sulfide as substrate. Consequently, GSH would be consumed stoichiometrically in the course of sulfur oxidation due to the preceding non-enzymic reduction of elemental sulfur to sulfide. GSnH + O2 + H2O GSn-1H + SO3

2- + 2 H+ n > 1 (1) So far, the sulfur chemistry and especially the fate of GSH has not been analyzed

leaving the actual activation mechanism for elemental sulfur unresolved. Therefore, we reinvestigated the GSH-dependent sulfur-oxidizing system. Special analytical attention was given to the organic sulfur species that occurred in the in vitro assay. Besides elemental sulfur other sulfane compounds and hydrogen sulfide were tested to elucidate the actual substrate used by the sulfur dioxygenase. For this purpose the enzymic activity of cell-free extracts of At. thiooxidans, At. ferrooxidans, Acidiphilium acidophilum, and Acidiphilium cryptum was investigated.


2.1 Bacterial strains, growth conditions, and cell-free preparations The bacterial strains of this study are listed in Table 1. Growth on elemental sulfur

powder (5 g/L) was performed in a salt solution modified from Mackintosh [15] with 2 mM NH4Cl and an initial pH of 3.0 adjusted with HCl. In addition, a lithotrophic iron medium [15] with 5 g/L iron(II) ions and a glucose-based medium [16] were applied for the growth of iron-oxidizers and facultative or obligate heterotrophs, respectively. All strains were cultivated aerobically at 28°C. Cell disruption was performed at ≤4°C under anaerobic conditions by treatment with glass beads [17]. Suspensions of disrupted cells were centrifuged (20 min, 25,000 x g, twice) to remove intact cells and cell residues. The supernatant, referred to as crude or cell-free extract, usually contained 1 to 2 g/L protein and was used either directly for activity assays or was stored under an anaerobic atmosphere at –25°C (less than 3 months).

2.2 Enzyme assays All assays were performed aerobically at 30°C in phosphate buffer (10 mM, pH 6.5)

with stirring at 300 rpm. If needed, the pH was maintained by titration with 50 mM KOH or 50 mM HCl. In order to determine non-enzymic reactions, assays without protein, with 0.2 g/L BSA, or with heat-inactivated crude extracts (90°C for 30 min) were used. For measuring elemental sulfur oxidation a system of dispersed sulfur in water was developed. To 50 mL of deionized water an equal volume of a saturated acetonic sulfur solution was added and dialyzed against 5 L deionized water to remove the acetone. The whitish dialysis product contained sulfur droplets of 2 to 10 µm in diameter and was used at a concentration of 4 mM. When sulfide, thiosulfate, tetrathionate, and p-toluolthiosulfonate were tested stock solutions of the respective potassium or sodium salts in deionized water were prepared. A mixture of GSSG and higher homologues was obtained by an incubation of 500 mM elemental sulfur (powder) with 100 mM GSH at pH 7.5 (adjusted with KOH) under stirring and anaerobic conditions until the solution got lemon-colored. At this point


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the pH was lowered to 5.0 by addition of HCl. The resulting hydrogen sulfide was removed by evacuation of the gas phase. Finally, the pH was adjusted to 6.5.

2.3 Analyses of sulfur compounds Thiosulfate, polythionates, GSH, GSH-derivatives, and p-toluolthiosulfonate were

analyzed by ion pair chromatography using tetrabutylammonium as counter ion [17]. A HPLC system from Kontron/BIO-TEK Instruments was applied. Chromatograms were recorded at 205, 215, 265, and 300 nm concomitantly with spectra from 190 to 320 nm. Elemental sulfur was analyzed by reversed-phase chromatography followed by UV-detection [18]. Sulfite and sulfate were quantified by ion exchange chromatography and conductivity detection [18] applying a Dionex DX 500 system. Sulfite was fixed with methanal prior to analysis [19]. Dissolved sulfide was determined by the methylene blue method [20]. With the exception of elemental sulfur quantification, all samples had been filtered (nylon filter, 0.2 µm) prior to analysis in order to remove elemental sulfur.


3.1 Oxidation of elemental sulfur via glutathione persulfide The enzymic oxidation of elemental sulfur by cell-free extracts from various meso-

acidophilic sulfur-oxidizing bacteria was performed mainly according to the sulfur dioxygenase assay of Suzuki [5,10]. As has been demonstrated previously for similar enzyme preparations (reviewed in [9]) enzymic oxidation of elemental sulfur could not be observed in the absence of GSH. With GSH, however, elemental sulfur was readily oxidized by crude extracts derived from sulfur oxidizers. In contrast to data of a previous study [21], preparations from the obligate iron(II) oxidizer Leptospirillum ferrooxidans did not show any sulfur oxidation activity (Table 1). The main reaction product of positive assays was thiosulfate (99%). Besides, traces of sulfite, sulfate, and glutathione S-sulfonate were detectable (data not shown). Obviously, sulfite as the first oxidation product rapidly reacted further with the finely dispersed sulfur to thiosulfate (equation 2). Only trace amounts were consecutively oxidized to sulfate or reacted with GSSG to glutathione S-sulfonate (equation 3). SO3

2- + 1/8 S8 S2O32- (2)

GSSG + SO32- + H+ GSSO3

- + GSH (3) GSSG was regularly detected in the assay solutions. It was formed in the course of the

non-enzymic reduction of elemental sulfur with GSH (equation 4 and 5). This reaction occurred in parallel to the enzymic activity and, besides GSSG, GSnG species with n ranging from 3 to 5 and hydrogen sulfide were formed (data not shown). S8 + GSH (GS9H) GSnH + (9-n)/8 S8 (4) GSxH + GSyH GSx+y-1G + H2S x ≥ 2, y ≥ 1 (5)

Comparing sulfur oxidation and reduction activities, it turned out that GSH was not consumed stoichiometrically in the course of elemental sulfur oxidation. In other words, free sulfide was not a substrate for the sulfur-oxidizing enzyme under the experimental conditions. Interestingly, the non-enzymic sulfur reduction by GSH (equation 4 plus 5) and the GSH-dependent enzymic sulfur oxidation were connected in another way: the higher the rates of the enzymic sulfur oxidation were, the lower were the GSH oxidation rates (Fig. 1). Furthermore, for a certain GSH concentration a maximal sulfur dioxygenase activity was obtainable, irrespective of the amount of enzyme, at which no further GSH oxidation occurred. For example, with 100 µM GSH an upper limit of about 350 µM/h


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was achieved (Fig. 1). The explanation for this phenomenon can be deduced from equations 1 and 5. Both reactions compete for the highly unstable monoorganylpolysulfane derivatives (GSnH, n > 1) of GSH. Summarizing, from the analyses of all relevant sulfur compounds in these sulfur dioxygenase assays we can fully confirm Suzuki´s proposal [5] of persulfides and their higher homologues being the actual substrates for the sulfur-oxidizing enzyme system of meso-acidophilic sulfur oxidizers (equation 1). This finding is valid for the genera Acidithiobacillus and Acidiphilium.

Table 1. Specific activities of sulphur dioxygenase in crude extracts of different strains of meso-acidophilic bacteria

strain substrate for cell growth*

activity† µmol.h-1.mg-1

Acidithiobacillus sulfur 3.8 ± 1.4 ferrooxidans R1 iron(II) 3.5 ± 0.6 Acidithiobacillus thiooxidans DSM 504 sulfur 2.4 ± 0.8 Acidithiobacillus thiooxidans K6 sulfur 1.5 ± 0.5 Acidiphilium sulfur 22.4 ± 5.4 acidophilum DSM 700 glucose 7.8 ± 2.5 Acidiphilium cryptum DSM 2389 glucose 0.3 ± 0.1 Leptospirillum ferrooxidans DSM 2705 iron(II) ND‡

*Strains were subcultured at least for 2 years on the indicated substrates. †Specific activity is expressed as the amount of sulfur atoms oxidized to the valence state of sulfite within 1 hour by 1 mg protein. ‡ND not detectable.

Figure 1. Sulfur dioxygenase assay with 100 µM GSH. Relationship between initial GSH consumption rate Vο and enzymic activity


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3.2 Oxidation of hydrogen sulfide Although the free sulfide formed during the incubation of elemental sulfur with GSH

was not oxidized enzymically, hydrogen sulfide was tested separately for enzymic oxidation. As expected, cell-free preparations incubated with or without GSH did not oxidize the sulfide. Surprisingly, low but significant enzymic activity was observed after GSSG addition to the assays (at 1 mM). In this case, the formation of thiosulfate and glutathione S-sulfonate was observed (Fig. 2). The latter was the main oxidation product. Furthermore, GSSG was reduced to GSH. From the stoichiometry of all these compounds (Fig. 2) the following reaction mechanism can be deduced. GSSG reacted with hydrogen sulfide to GSH and GSSH (equation 6). The persulfide was consecutively oxidized by the sulfur dioxygenase to sulfite (equation 1) and then formed with excess GSSG the glutathione S-sulfonate (equation 3). GSSG + H2S GSSH + GSH (6) 2 GSSG + H2S + O2 + H2O 3 GSH + GSSO3

- + H+ (7) Summing up, GSH and glutathione S-sulfonate should be produced in a ratio of 3:1

(equation 7) which agrees quite well with the observed values (Fig. 2). Consequently, sulfide was not oxidized directly, as proposed by Sugio and coworkers [14], but via the formation of persulfide species.

0

5

10

15

20

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entra

tion

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)

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2.5

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3.5

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ratio

GSH

/GSS

O3-

GSH GSSO3-

thiosulfate ratio GSH/GSSO3-

Figure 2. Oxidation of sulfide by cell-free extracts in the presence of 1 mM GSSG

3.3 Oxidation of other sulfane compounds From the results of the oxidation tests performed with elemental sulfur and sulfide, it

can be clearly deduced that the active enzyme preparations contained no enzymic activity for the direct oxidation of these sulfur species. Only the sulfane sulfur of persulfides was oxidizable. In additional experiments the specificity of this enzymic activity was tested with various inorganic and organic sulfane compounds (see 2.2.). Briefly, with none of these compounds an oxidation activity was observed. However, when a mixture of GSSG and higher GSnG species (2<n<6) was incubated with GSH the higher bisorganylpolysulfanes were readily oxidized by active crude extracts (Fig. 3). This indicates that the GSnG species were also degraded via GSnH compounds according to equation 8. The sulfane sulfur of the polysulfanes was oxidized to thiosulfate and glutathione S-sulfonate. The latter compound was formed in parallel with GSH (Fig. 3 A).


1176

Consequently, it was highly likely formed from sulfite and GSSG because this reaction produced equimolar amounts of GSSO3

- and GSH (equation 3). GSnG + GSH GSxH + GSyGn ≥ 3, x ≥ 2, y ≥ 2 (8)

Summarizing, only the sulfane sulfur of persulfide compounds can be oxidized by the sulfur dioxygenase of meso-acidophilic bacteria. Neither polythionates nor bisorganylpolysulfanes are oxidized directly.

0

50

100

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te (µ

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200

300

400

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, GSS

O3- (µ

M)

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GS n

G (

mA

bs m

in)

800

900

1000

1100

GSS

G (µ

M)

B

GS3G GS4G GS5G G SSG

Figure 3 A+B. Oxidation of GSnG species (2 < n < 6) by cell-free extracts in the presence of 200 µM GSH. Amounts of GSnG are expressed as peak areas recorded at 215 nm

3.4 Proposal of a biochemical model for sulfur and sulfide oxidation Compiling all data available, a model for the oxidation of sulfide and elemental sulfur

can now be proposed (Fig. 4). Extracellular elemental sulfur (S8) reacts with thiol groups of special outer-membrane proteins (OMP-SH) and, thus, can be transported as persulfide sulfur (OMP-SSH) into the periplasmic space. Here sulfur dioxygenase (SDO) takes over the sulfane sulfur and oxidizes it to sulfite. Sulfite is oxidized further to sulfate by a sulfite:acceptor oxidoreductase (SOR). This enzyme most probably uses cytochromes (Cyt) as electron acceptors [22,23,24]. Free sulfide is oxidized by a separate sulfide:quinone oxidoreductase (SQR), which is located at the periplasmic site of the cytoplasmic membrane [25,26]. For this reaction polysulfides have recently been proposed to be the oxidation products [27]. However, at least for acidophilic bacteria we believe that membrane-bound persulfides (OMP-SSH) as products of SQR activity are more likely (Fig. 4). In fact, the membrane binding of sulfur at zero valence state was demonstrated in sulfide oxidation experiments with cell-free extracts from strains of At. thiooxidans [25,28]. Candidates for the not yet identified thiol-bearing OMP are the sulfide-binding


1177

protein isolated from At. ferrooxidans AP 19-3 [29] and several outer membrane proteins that have been associated with sulfur oxidation in strains of At. ferrooxidans [30,31,32].

Figure 4. Proposed biochemical pathway for sulfide and elemental sulfur oxidation in leaching bacteria of the Gram-negative genera Acidithiobacillus and Acidiphilium (see text).

ACKNOWLEDGEMENTS This work was supported by grants to W. S. from BMBF via UBA (1490954), DBU

(05333), and the Max-Buchner-Forschungsstiftung (DECHEMA e.V.).

REFERENCES 1. M. Schedel and H. G. Trüper, Biochim. Biophys. Acta, 568 (1979) 454. 2. A. S. Pott and C. Dahl, Microbiology, 144 (1998) 1881. 3. C. G. Friedrich et al., Appl. Environ. Microbiol., 67 (2001) 2873. 4. D. Rother et al., J. Bacteriol., 183 (2001) 4499. 5. I. Suzuki, Biochim. Biophys. Acta, 110 (1965) 97. 6. T. Emmel et al., J. Gen. Microbiol., 132 (1986) 3415. 7. A. Kletzin, J. Bacteriol., 171 (1989) 1638. 8. A. Kletzin, J. Bacteriol., 174 (1992) 5854. 9. I. Suzuki, Methods Enzymol., 243 (1994) 455. 10. I. Suzuki, Biochim. Biophys. Acta, 104 (1965) 359. 11. M. Silver and D. G. Lundgren, Can. J. Biochem., 46 (1968) 457. 12. T. Sugio et al., J. Bacteriol., 169 (1987) 4916. 13. E. N. Krasil`nikova et al., Mikrobiologiya, 67 (1998) 156. 14. T. Sugio et al., Biochim. Biophys. Acta, 973 (1989) 250. 15. M. E. Mackintosh, J. Gen. Microbiol., 105 (1978) 215.


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16. A. P. Harrison Jr., Int. J. Syst. Bacteriol., 31 (1981) 327. 17. T. Rohwerder, Untersuchungen zur enzymatischen Oxidation von Elementarschwefel

bei acidophilen Laugungsbakterien, PhD thesis, University of Hamburg, 2002 (www.sub.uni-hamburg.de/disse/849/dissertation.pdf).

18. A. Schippers and B. B. Jørgensen, Geochim. Cosmochim. Acta, 65 (2001) 915. 19. J. Weiß, Ionenchromatographie, VCH, Weinheim, 1991. 20. Anonymous, Deutsche Einheitsverfahren zur Wasser-, Abwasser- und Schlammunter-

suchung (DEV), VCH, Weinheim, 1984. 21. T. Sugio et al., Appl. Environ. Microbiol., 58 (1992) 431. 22. J. R. Vestal and D. G. Lundgren, Can. J. Biochem., 49 (1971) 1125. 23. K. Nakamura et al., Biosci. Biotech. Biochem., 59 (1995) 11. 24. G. A. H. de Jong et al., J. Mol. Catal. B, 8 (2000) 61. 25. D. J. W. Moriarty and D. J. D. Nicholas, Biochim. Biophys. Acta, 184 (1969) 114. 26. T. Nübel et al., Arch. Microbiol., 173 (2000) 233. 27. C. Griesbeck et al., Biochemistry, 41 (2002) 11552. 28. D. J. W. Moriarty and D. J. D. Nicholas, Biochim. Biophys. Acta, 197 (1970) 143. 29. T. Sugio et al., Agric. Biol. Chem., 55 (1991) 2091. 30. V. Buonfiglio et al., FEMS Microbiol. Rev., 11 (1993) 43. 31. V. Buonfiglio et al., J. Biotechnol., 72 (1999) 85. 32. N. Ohmura et al., J. Bacteriol., 178 (1996) 5776.



1179

Sulfate reduction at low pH by mixed cultures of acidophilic bacteria

Anna M. Sen, Sakurako Kimura, Kevin B. Hallberg and D. Barrie Johnson


Abstract Mixed cultures of acid-tolerant sulfate reducing bacteria (SRB) and other acidophilic

bacteria were shown to form stable microbial communities in sulfidogenic bioreactors. Cultures were set up in pH-controlled (pH 2.5-5.5) bioreactors, and run in batch and continuous mode using glycerol or ethanol as electron donor. Reduction of sulfate (and corresponding formation of hydrogen sulfide), acid-consumption and changes in the indigenous microflora (SRB and other bacteria) were monitored. Sparging of the bioreactors with nitrogen effectively removed the H2S, which was used to precipitate copper as CuS (off-line) or zinc as ZnS (within the bioreactor). Following earlier experiments with undefined mixed populations of anaerobic acidophiles, we studied sulfidogenesis at low pH by mixed cultures of Desulfosporosinus spp. (SRB) and an Acidocella-like acidophilic heterotroph. In these cultures, accumulation of acetic acid (the product of incomplete glycerol oxidation by the SRB) was much less than in the undefined mixed cultures, leading to more efficient metal (zinc) immobilization in. Syntrophy between acid-tolerant SRB and acidophilic heterotrophic bacteria appears to be important in mediating sulfidogenesis in acidic liquors.

Keywords: acid mine drainage; acidophiles; bioremediation; metal recovery; sulfate reducing bacteria

1. INTRODUCTION Studies of the microbiology in extremely acidic liquors, such as acid mine drainage

(AMD) and mineral leachates, have focused primarily on aerobic processes, chiefly because the major reactions responsible for mineral dissolution are oxidative. Some acidophiles are able to grow in anoxic environments [1]. Iron respiration (where ferric iron acts as electron acceptor) seems particularly widespread amongst acidophiles. Iron-oxidizing moderate thermophiles and obligately heterotrophic Acidiphilium spp. can couple the oxidation of organic substrates to the reduction of ferric iron [2-4]. Acidithiobacillus ferrooxidans can also grow anaerobically by iron respiration, using either reduced sulfur or hydrogen as electron donor [5, 6].

Sulfate is found in mineral leaching environments often at concentrations of many grams per liter. Dissimilatory sulfate reduction might therefore be anticipated to be widespread in acidic anaerobic zones (sediments etc.), and there are a number of reports in the literature documenting sulfidogenesis in these situations [e.g. 7, 8, 9]. However, attempts to isolate and cultivate acid-tolerant/acidophilic sulfate reducing bacteria (aSRB)


1180

from mine waters have mostly been unsuccessful. Most mine water isolates appear to be neutrophilic, and are not active below pH ~ 5 [e.g. 7, 10]. There are a number of reasons for this, including the use of inappropriate media (substrates, such as lactate, used routinely to enrich neutrophilic species may be toxic to acidophiles). Sen and Johnson [11] obtained enrichment cultures of aSRB from sediment samples from a copper mine and from geothermal areas on Montserrat. Sulfidogenesis by these mixed cultures occurred at pH <3, but was more rapid between pH 3 and 5.

This paper describes sulfidogenesis by undefined and defined cultures of aSRB, and the immobilization of heavy metals at low pH due to biogenic sulfide. Syntrophy between aSRB and a non-sulfidogenic heterotrophic acidophile is discussed.


2.1 Establishment and maintenance of enrichment cultures Enrichment cultures containing acidophilic/acid-tolerant strains of SRB were obtained

from sediments from the Afon Goch, Parys copper mine (north Wales) and Montserrat, as described elsewhere [11]. These were subcultured in ~ 500 ml of ethanol-containing liquid medium (below) containing ~ 300 ml of porous glass beads (2-4 mm diameter; [12]). The beads were included to act as a surface for bacterial colonization. These cultures were used to inoculate 2 l bench-scale bioreactors (P350; Electrolab, U.K.), which were run in either batch or continuous mode. A schematic diagram of the continuous mode set up is shown in Figure 1. Cultures were maintained at 30°C, and pH control was maintained via addition of either 0.1 M HCl or KOH.

WASTEPUMPED

OUT

MEDIUM PUMPED IN ORP/pH electrodes

Impeller

Acid/ alkali

OFN in

Biofilm of aSRB on Poraver beads

2L bioreactor vessel Gas collection bottles

each containing 500 ml copper acetate (10 mM)

(to samplepoint output and waste vessel)

Sample point (input)

Filter

Figure 1. Bioreactor system set-up for continuous mode experiments. OFN = oxygen-free nitrogen


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2.2 Batch and continuous cultures of mixed (undefined) SRB cultures The growth medium for batch culture experiments contained 10 mM ethanol, 0.002%

yeast extract, 10 mM FeSO4, together with basal salts and trace elements [13]. The medium was heat-sterilized prior to addition of filter-sterilized ethanol and FeSO4. The cultures (working volumes, 1.8 l) were sparged daily for 5-10 minutes with oxygen-free nitrogen (OFN) to remove H2S, which might otherwise inhibit the SRB. Samples were removed at regular intervals and analyzed for sulfate, ferrous iron, ethanol, acetate and protein. Bacteria were isolated and enumerated on overlay solid media, as described elsewhere [11].

The bioreactor was run in continuous mode for the Parys mixed culture only. In this case, a modified growth medium was used, with glycerol (at 5 mM) replacing ethanol as electron donor. In addition, the concentration of FeSO4 was lowered to 1 mM, with sulfate concentrations being maintained as earlier by the inclusion of 9 mM K2SO4. The sterile feed medium was adjusted to pH 3.0 with H2SO4, and was added to the culture vessel at 25 ml/h, while that in the bioreactor was maintained at pH 2.5, 3.0, 3.5 or 4.0. In addition, inflow rates of 50 and 100 ml/h were applied in the pH 4.0-maintained culture. Throughout the working day, the reactor vessel was sparged for about 5 min/h with OFN and the off-gas bubbled though gas bottles containing 10 mM copper acetate, causing precipitation of CuS. Concentrations of glycerol and acetate in the growth vessel were also measured in continuous mode experiments.

2.3 Batch cultures of defined mixed SRB cultures During the course of purifying aSRB on solid media, a novel uncolored colony

morphology was observed on several overlay plates. Physiological and phylogenetic work confirmed that this was an Acidocella-like isolate (unpublished); bacteria of this type had not previously been found in anaerobic bioreactor cultures. A mixed culture of this isolate, PFBC, and Desulfosporosinus isolates from the Parys culture (isolate P1) and from the Montserrat culture (isolate M1) was grown in batch mode in an anaerobic bioreactor. A modified medium was used in this case, containing 5 mM glycerol, 5 mM ZnSO4 5 mM K2SO4 and 1 mM FeSO4. Yeast extract was replaced by the vitamin mixture proposed by Widdel and Pfennig [14] for SRB, added at 1 ml vitamin solution/l medium.

2.4 Analytical techniques Samples were removed from the bioreactor vessels and filtered through 0.2 µm

cellulose nitrate membranes (Whatman, U.K.). Concentrations of ethanol, glycerol and acetate were measured enzymatically (R-Biopharm GmbH, Germany), and protein using the Bradford assay [15]. Sulfate was measured turbidometrically as barium sulfate (Hydrocheck, Cambridge, U.K.), and ferrous iron using the ferrozine reagent [16]. Sulfide production in continuous cultures was estimated by analysis of copper concentrations in the gas bottles by atomic absorption spectrophotometry. Proton consumption by batch and continuous cultures was determined from the amounts of alkali that were required to maintain the cultures at pre-set pH levels.

2.5 Calculations Sulfate reduction rates (SRR) were calculated in mmol SO4

2- reduced l-1·h-1. Hydraulic residence time (HRT), i.e. the average time that the culture medium was retained in the bioreactor, was calculated by dividing the working volume by the flow rate of feed medium.


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3. RESULTS Mixed cultures of aSRB and other anaerobic bacteria were established readily in

enrichment cultures, as described previously [11]. Both the Parys and Montserrat cultures contained aSRB and other bacteria (as evidenced by the formation of black and non-black colonies, respectively, on overlay plates). Since these mixed cultures appeared to be robust (i.e. they were maintained through many sub-cultures) it was decided to carry out initial batch- and continuous-mode experiments with them.

Data from a batch culture run (at pH 4.0), using the Parys culture, is shown in Figure 2. Sulfate reduction in this culture corresponded with proton consumption (alkali generation) that was due, at least in part, to production of H2S from sulfuric acid. Ferrous sulfide was not formed in significant amounts (as would be predicted, given the pH of the culture). Interestingly, ethanol oxidation was only about 3 mM, which was much less than the sulfate reduced (~ 10 mM). Similar data were obtained with the Montserrat culture (not shown). In both cultures, there were general, though sporadic, increases in protein concentrations during culture runs, indicating that the cultures were actively growing under these conditions (data not shown).

Figure 2. Sulfate reduction by the undefined Parys consortium, run in batch mode at pH 4.0 Key: , sulfate; , ethanol; , ferrous iron; , protons consumed

Sulfate reduction was observed with both cultures over the pH range 2.5-5.5. Rates of sulfate reduction were, however, affected by culture pH, and the Parys and Montserrat cultures showed difference responses (Figure 3). The Parys culture displayed a pH optimum for sulfate reduction at 4.0. In contrast, rates of reduction were similar with the Montserrat culture between pH 3.0 and 4.0 (and greater than at pH 2.5 and 4.5) and the greatest rate of sulfate reduction was observed at pH 5.5. The shape of the graph obtained with the Montserrat culture indicated that a "second" maximum in sulfate-reduction rate may occur above this pH (Figure 3).

Data from the Parys culture run in continuous mode (pH 4.0; hydraulic retention time 16.5 h) is shown in Figure 4. Sulfate concentrations in the outflow liquor (which were very similar to those in the bioreactor vessel) were 4-5 mM lower than in the inflowing liquor, as a result of sulfate reduction. Similarly, glycerol concentrations were 3-4 mM


1183

less in the outflow than in the feed liquor, indicating that it was being utilized as the electron donor for sulfate reduction. Acetic acid was present at about 2 mM in the bioreactor vessel and outflow liquor, presumably as a result of incomplete oxidation of glycerol. In cultures maintained at lower pH, glycerol oxidation again appeared to be coupled to sulfate reduction, though concentrations of inflow/outflow glycerol were more similar at pH 3.0 and 2.5 (indicating less active sulfate reduction) than at pH 3.5 and 4.0. Data from the Parys culture run in continuous mode, showing the effects of culture pH and hydraulic retention times, are shown in Table 1.

Figure 3. Sulfate reduction rates (SRR) for batch cultures of ( ) the Parys and ( ) the Montserrat undefined mixed cultures

Figure 4. Sulfidogenesis by the undefined Parys consortium run in continuous mode (pH 4.0; HRT 16.5 h). Key: ( ) sulfate (in); ( ) sulfate (out); ( ) glycerol (in); ( ) glycerol (out); ( ) acetate (out)


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Table 1. Sulfate reduction rates (SRR) in continuous culture experiments with the Parys undefined mixed culture, as affected by culture pH and hydraulic retention times (HRT)

pH HRT (h)

SRR (µmol SO4

2- reduced/h) 2.5 66 40 3.0 66 82 3.5 66 166 4.0 66 134 4.0 33 340 4.0 16.5 460

Data from batch growth of the defined SRB-Acidocella culture are shown in Figure 5. The inclusion of zinc sulfate in this instance led to the formation of zinc sulfide, thereby reducing the amount of acid consumption (to maintain a low pH) by the culture, due to the reaction: Zn2+ + H2S → ZnS + 2H+

Again, glycerol oxidation was coupled to sulfate reduction, which in turn resulted in the removal from solution of stoichiometric amounts of zinc. It was also interesting that, although some acetic acid was produced during days 1-5, this was rapidly consumed during day 6, with concentrations falling from about 2 to <0.1 mM (Figure 5).

Figure 5. Sulfidogenesis coupled to precipitation of zinc in a defined culture of Desulfosporosinus and Acidocella PFBC, grown in batch mode. Key: ( ) sulfate reduced; ( ) soluble zinc; ( ) glycerol; ( ) acetate

4. DISCUSSION Dissimilatory reduction of sulfate to sulfide is of significance in two major respects,

both of which are important for the bioremediation of acidic, metal-rich waters, such as AMD. First is the production of net alkalinity, resulting from the conversion of a strong acid (sulfuric) to a weak acid (hydrogen sulfide), and from the generation of bicarbonate. Secondly, sulfate reduction leads to the precipitation of a number of potentially toxic


1185

heavy metals (cadmium, copper, zinc etc.) as highly insoluble sulfides. Both of these key features were illustrated in the present study, where mixed cultures of aSRB and other acidophiles required addition of protons to bioreactor cultures in order to maintain the pre-set pH levels, and where copper and zinc were both removed from solution (copper off-line by sparging the sulfide from the culture vessel with nitrogen, and zinc in situ).

Sulfidogenesis is used commercially to remediate acidic waters and to recover metals, such as at the Budelco zinc refinery in the Netherlands [17]. However, these systems use neutrophilic strains of SRB, and it is necessary to preclude contact between the SRB cultures and the acidic wastewaters, which is usually achieved by stripping H2S/HS- from the SRB reactors and allowing the sulfide to contact the waste stream in separate tanks. Considerable savings in costs and engineering complexity could be achieved by using single SRB/waste stream tanks, maintained at low pH to allow selective precipitation of metals (e.g. zinc and copper, but not iron). This work has shown that this can be achieved using acid-tolerant/acidophilic strains of SRB, grown in mixed cultures with other acidophiles. Even though ferrous iron was present in all of the culture media used, FeS was not formed in the pH 4.0 reactors, since its solubility product (4.0 x 10-19 M2) is considerably greater than that of other metal sulfides such as zinc sulfide (1.0 x 10-23 M2), which does precipitate at pH 4.0.

Other studies have indicted that, although it is possible to purify aSRB from the consortia, the mixed cultures are inherently more robust than pure cultures (unpublished). One reason for this is that the Desulfosporosinus isolates studied here are incomplete oxidizers, i.e. they produce (and accumulate) acetate from ethanol and glycerol. This is a common trait amongst SRB, but acidophiles have the added problem that organic acids (which occur primarily as undissociated acids in low pH liquors) are toxic at relatively low concentrations [18]. The presence of other bacteria in the anaerobic cultures might therefore serve to lower concentrations of acetic acid to the point at which they are non-toxic. Support for this hypothesis came from the work with the defined population of Desulfosporosinus sp. M1 and Acidocella-like acidophiles. The latter is most closely related to Acidocella sp. WJB3, which is known to be unusually adept (for an acidophile) at metabolizing acetic acid [19]. However, Acidocella spp. are all supposedly obligate aerobes. Attempts to grow pure cultures of strain PFBC anaerobically on acetic acid (or any other substrate) have all failed, yet data from molecular analysis of mixed aSRB/PFBC cultures have shown that the anaerobic aSRB and “aerobic” PFBC grow concurrently in anoxic media (data not shown). Work is currently underway to resolve this apparent conundrum. What was apparent is that acetate concentrations in mixed cultures of Desulfosporosinus M1/Acidocella PFBC were considerably lower than in the "undefined" aSRB consortia (where Bacillus-, Cellulomonas- and Luteococcus-like isolates were detected [13] and Acidocella-like bacteria were not detected).

This work has confirmed that SRB may be metabolically active in acidic liquors, and that their potential for remediation of AMD and related wastewaters is considerable.

ACKNOWLEDGEMENTS A.M.S. is grateful to the BBSRC (U.K.) and British Nuclear Fuels Limited (U.K.) for

provision of a research studentship. K.B.H. and D.B.J. thank the BBSRC (ref. 5/T06496) and the DTI (U.K.; ref. BTL/20/71) for partial support of this work.

REFERENCES 1. D. B. Johnson, FEMS Microbiol. Ecol. 27 (1998) 307.


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2. T. A. M. Bridge and D. B. Johnson, Appl. Environ. Microbiol. 64 (1998) 2181. 3. T. A. M. Bridge and D. B. Johnson, Geomicrobiol. J. 17 (2000) 193. 4. K. Küsel, T. Dorsch, G. Acker and E. Stackebrandt, Appl. Environ. Microbiol. 65

(1999) 3633. 5. J. T. Pronk, J. C. de Bruyn, P. Bos and J. G. Kuenen, Appl. Environ. Microbiol. 58

(1992) 2227. 6. N. Ohmura, K. Sasaki, N. Matsumoto and H. Saiki, J. Bacteriol. 184 (2002) 2081. 7. J. H. Tuttle, P. R. Dugan, C. B. Macmillan and C. I. Randles, J. Bacteriol. 97 (1969)

594. 8. R. A. Gyure, A. Konopka, A. Brooks and W. Doemel, FEMS Microbiol. Ecol. 73

(1990) 193. 9. D. Fortin, M. Roy, J. P. Rioux and P. J. Thibault, FEMS Microbiol. Ecol. 33 (2000)

197. 10. K. A. Küsel, U. Roth, T. Trinkwalter and S. Peiffer, Environ. Exp. Bot. 46 (2001) 213. 11. A. M. Sen and D. B. Johnson, Biohydrometallurgy and the Environment Toward the

Mining of the 21st Century, R. Amils and A. Ballester (eds.), Elsevier, Amsterdam, (1999) 709.

12. Å. Kolmert and D. B. Johnson, J. Chem. Technol. Biotechnol. 76 (2001) 836. 13. A. M. Sen, Ph.D. Thesis, University of Wales, Bangor, (2001). 14. F. Widdel and N. Pfennig, Arch. Microbiol. 129 (1981) 395. 15. M. M. Bradford, Anal. Biochem. 72 (1976) 248. 16. D. R. Lovley and E. J. P. Phillips, Appl. Environ. Microbiol. 53 (1987) 1536. 17. L. J. Barnes, P. J. M. Scheeren and C. J. N. Buisman, Emerging Technology for

Bioremediation of Metals, J. L. Means and R. E. Hinchee (eds.), Lewis Publishers, Boca Raton, Fl., (1994) 38.

18. B. Alexander, S. Leach and W. J. Ingledew, J. Gen. Microbiol. 133 (1987) 1171. 19. K. B. Hallberg, Å. K. Kolmert, D. B. Johnson and P. A. Williams, Biohydrometallurgy

and the Environment Toward the Mining of the 21st Century, R. Amils and A. Ballester (eds.), Elsevier, Amsterdam, (1999) 719.



1187

Sulfur assimilation in Acidithiobacillus ferrooxidans

J. Valdés1, E. Jedlicki2 and D.S. Holmes1,3*

1 Laboratory of Bioinformatics & Genome Biology, University of Santiago, Santiago, Chile,

2 Program of Cellular and Molecular Biology, University of Chile, Santiago, Chile 3 Millenium Institute of Fundamental and Applied Biology, Santiago, Chile

Abstract A bioinformatic analysis of the nearly complete genome sequence of Acidithiobacillus

ferrooxidans, available from The Institute for Genome Research (TIGR) and Integrated Genomics (IG), reveals the presence of a number of genes potentially involved in sulfur uptake and metabolism. The coding potential of these genes is compared to similar genes for known sulfur metabolic pathways in other microorganisms. In addition, the proposed sulfur assimilation genes of A. ferrooxidans are shown to be present in operons in an organization similar to that found in some other organisms. Genes potentially involved in the regulation of sulfur uptake and assimilation have also been detected. Based upon these observations, we present a metabolic model of sulfur uptake and assimilation in A. ferrooxidans that is substantially similar to the genetic organization and metabolic pathway structure that has been experimentally demonstrated to function in other organisms. An interesting exception to this similarity is the observation that A. ferrooxidans appears to have the genetic potential to encode for two different pathways of sulfur activation whereas many organism only have one such pathway. In addition, an interesting gene fusion event seems to have taken place in A. ferrooxidans involving the joining of a cysN gene, encoding a sulfurylase enzyme subunit B, with a cysC-like gene encoding an APS activity.

Keywords: sulfur assimilation, sulfur regulation, cysteine biosynthesis, genome analysis, Acidithiobacillus ferrooxidans

1. INTRODUCTION The assimilation of sulfur is an essential process in all cells being required for the

formation of the amino acids, methionine and cysteine. Sulfur is also used for the construction of FeS centers in proteins involved in electron transfer processes (1) and is found in certain membrane components such as the nod factor of the plant symbiont Rhizobium sp. (2) and in various sulfolipids of Mycobacterium sp. that are thought to bevirulence factors (3).

* Corresponding author: David Holmes: [email protected]. Work supported by Fondecyt No. 1010623 and the Millenium Institute for Fundamental and Applied Biology, Santiago, Chile. We thank the Institute of Genome Research (TIGR) and Integrated Genomics, Inc. (IG) for the use of their partial sequence of the Acidithiobacillus ferrooxidans genome.


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The process of sulfur assimilation, in all organisms begins with the uptake of sulfate ions from the environment through the cellular membrane via an ABC-type membrane pump mechanism. Sulfate ions are generally available in most environments and are abundantly available in mining environments rich in sulfuric acid.

After entry into the cell, the sulfate is made biologically available by activation, a process that involves the reaction between sulfate and ATP, catalyzed by the enzyme ATP sulfurylase to yield adenosine 5´-phosphosulfate (APS) and PPi. A second enzyme, APS kinase, can further phosphorylate APS with another ATP to give 3´-phosphoadenosine 5´-phosphosulfate (PAPS)(1). In plants and phototrophic bacteria the intermediate of assimilative reduction is APS, but in fungi and some chemotrophic bacteria it is PAPS. Depending on the organism, these intermediates are then further reduced and fed into pathways that are shared in all organisms leading to the formation of the amino acids cysteine and methionine. The reason for the evolution of two mechanisms for sulfur assimilation by different organisms is unclear. The enzymes involved in the formation of APS or PAPS show significant sequence similarity and may evolve from a common ancestor.

We have carried out a bioinformatic analysis of the almost complete genome sequence of A. ferrooxidans in an attempt to identify candidate genes potentially involved in sulfur assimilation. Using this information, we suggest possible biochemical pathways for sulfur assimilation and for the regulation of sulfur uptake. It is beyond the scope of this paper to connect the proposed sulfur assimilation pathways with proposed sulfur oxidation genes and pathways in A. ferrooxidans. However, the present paper will aid future studies designed not only to make these connections but ultimately to deduce the interconnected regulatory mechanism that balance sulfur assimilation with sulfur oxidation in a variety of environmental situations.

2. METHODS Known metabolic pathways of sulfur assimilation were obtained from BIOCYC

(http://biocyc.org:1555/META/server.html), KEGG (http://genome.ad.jp/kegg/kegg4. html) and ERGO (http://wit.integratedgenomincs.com/WIT2/CGI). Amino acid sequences derived from genes identified as being involved in sulfate acquisition were used as query sequences to search the partial genome sequence of A. ferrooxidans ATCC 23270 in the TIGR (http:// www.tigr.org/) and ERGO data bases using TBlastN and BlastP respectively. When a prospective candidate gene was identified in TIGR or ERGO its predicted amino acid sequence was then used to formulate a BlastP (http://www.ncbi.nlm.nhi.gov/BlastP/) search of the non-redundant database at NCBI. Only bidirectional best hits were accepted as evidence for putative homologs. Candidate genes and their translated proteins were further characterized employing the following bioinformatic tools available in the web: primary structure similarity relations (http://www.ebi.c.uk/ClustalW/), secondary structure predictions JPred http://www. compbio.dundee.ac.uk/Software/JPred/jpred.html), transmembrane predictions (http://www.ch.embnet.org/software/TMPRED_form.html), motif predictions (http:// www.blocks.fhcrc.org/, http://www.ebi.ac.uk/interpro/, and domain predictions (http:// www.tolouse.inra.fr/prodom.htlm), and prediction of protein localization sites (http://psort.nibb.ac.jp/).


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3. RESULTS AND DISCUSSION Sulfur assimilation, in most organisms, typically starts with sulfate uptake from the

environment by a membrane complex that includes an ABC-type membrane pump (Figure 1). A bioinformatic analysis of the predicted genes of A. ferrooxidans reveals the presence of candidate genes that could encode the proteins for such a complex. One of these genes encodes a possible sulfate-permease protein and lies adjacent to a gene potentially encoding an ABC-type ATP binding membrane pump and a protein of unknown function. The three genes appear to be organized as an operon (data not shown) lending support to the idea that they encode proteins involved in sulfate uptake. On the other hand a positive identification of genes and proteins involved in sulfate uptake is difficult because ABC-type membrane pump proteins belong to a large family of paralogous genes that encode proteins with similar sequences and structures but that differ in the substrate that they transport. Thus bioinformatic identification of an ABC transporter for sulfate uptake by sequence similarity is insufficient to pinpoint its function and experimental validation will be essential.

On the other hand, genes involved in the subsequent activation and utilization of sulfur do not appear to belong to extensive paralogous families. In addition, the genes and pathways of sulfur assimilation are quite conserved among organisms. These considerations make it possible not only to identify candidate orthologous genes in A. ferrooxidans by bioinformatic analysis but also to have a reasonable degree of confidence in such assignations. It also allows us to propose a biochemical model of sulfur assimilation for A. ferrooxidans. It should be stressed that such a model requires experimental validation before it can be accepted. Such experiments are now underway and will be reported in the future. However, it is useful to note that, in the absence of well developed methods of gene transfer in A. ferrooxidans and the concomitant paucity of genetic analyses, metabolic models derived by bioinformatics provide an important way to investigate the potential biochemistry and genetics of the organism. Such studies have already revealed valuable information regarding the amino acid biosynthesis (4), nitrogen metabolism, metal fluxes and other characteristics of A. ferrooxidans. (5).

Table 1. Identification and some properties of the genes predicted to be involved in sulfur assimilation in A. ferrooxidans. (A) Proposed gene name. (B) Putative function assigned to each gene. (C) The best BlastP hit of each candidate gene to orthologs in GenBank and the organism in which the ortholog is found and (D) the percent similarity of the amino acid sequence of the candidate gene with its ortholog from GenBank

A. Gene B. Putative function assigned C. Best Blastp hit and organism

D. % Similarity

CysD1 ATP sulfurylase, catalitic subunit Neisseria meningitidis 71 CysD2 ATP sulfurylase, catalitic subunit Mesorhizobium loti 78 CysN1 ATP sulfurylase, GTP-binding subunit Neisseria meningitidis 69 CysN2 ATP sulfurylase/Adenylylsulfate kinase Mesorhizobium loti 72 CysN3 ATP sulfurylase/Adenylylsulfate kinase Coxiella burnetii 61 CysH APS reductase Allochromatium vinosum 74 CysI Sulfite reductase hemoprotein Bacillus subtilis 67 CysJ Sulfite reductase flavoprotein Bacillus subtilis 59 CysB1 Transcriptional regulator Pseudomonas aeruginosa 69 CysB2 Transcriptional regulator Pseudomonas aeruginosa 52


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A. Gene B. Putative function assigned C. Best Blastp hit and organism

D. % Similarity

CysQ PAPS catabolism Aquifex aeolicus 72 CysA Sulfate /thiosulfate transport Mesorhizobium loti 53 CysW Sulfate/ thiosulfate transport Halobacterium sp. 54 CysM O-acetyl serine (thiol)-lyase-B Pseudomonas putida 70

In contrast to many organisms, where sulfur activation and subsequent reduction occurs either via APS or PAPS, candidate genes have been identified A. ferrooxidans that are predicted to encode enzymes involved in both APS and PAPS formation (Figure 1 and Table 1). In addition, genes have been identified for proteins involved in the subsequent utilization of APS and PAPS and also in the regulation of the uptake and incorporation of sulfur (Figure 1 and Table 1).

Figure 1. Proposed biochemical pathway and genes for the uptake of sulfate and the assimilation of sulfur in A. ferrooxidans based upon bioinformatics analysis of its genome. Potential genes are shown in parentheses. The reason for the dotted line that connects PAPS and sulfite is discussed in the text. Also shown is the feedback inhibition of serine acetyl transferase by cysteine and the regulation of the sulfur uptake and assimilation genes by N-acetyl-L-serine which is discussed in the text

The levels of similarity (Table 1) of the predicted A ferrooxidans amino acid sequences to known proteins involved in sulfur activation are sufficient to allow a reasonable degree of confidence in the assignments made. In addition, many of the candidate sulfur assimilation genes appear to have an operon-like organization similar to those found in other organisms (Figures 2 and 3). Operon-like organization is a strong predictor of similar function for the genes within the operon. Thus, the finding that many


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of the candidate sulfur assimilation genes of A. ferrooxidans are display such organization strengthens the likelihood of the assignments of gene function which we have made using bioinformatics predictions based on comparative protein sequence similarity.

The formation of APS in A. ferrooxidans could be accomplished by the protein products of cysD1 and cysN1, which, in E. coli, encode the heterodimeric enzyme ATP sulfurylase (6). In both A. ferrooxidans (Figure 2) and E. coli cysD1 and cysN1 appear to be organized in an operon. However, the E. coli operon also includes the gene cysC that catalyzes the phosphorylation of APS to give PAPS. This gene appears to be absent in the equivalent A. ferrooxidans operon. However, the A. ferrooxidans genome has an additional two copies of cysN1, which we term cysN2 and cysN3 (Figures 1,2 and Table 1). Interestingly, a detailed examination of the predicted protein sequences encoded by cysN2 and cysN3 reveals the presence of an additional domain in their C terminal regions that exhibit significant sequence similarity to the E. coli cysC (Figure 3). This raises the possibility that the A. ferrooxidans cysN2 and cysN3 encode bi-functional proteins formed by a gene fusion event between cysN1 and cysC and that the resulting fused protein can assume the function of the missing cysC catalyzing the formation of PAPS. A similar bifunctional protein has also been observed in Rhizobium sp. (7). Additional support for this conjecture comes from the observation that a cloned segment of A. ferrooxidans DNA was able to complement an E. coli cysC mutant, demonstrating that the organism has the capacity to encode a cysC-like function (8).

It appears that a transpose-encoding element (Tpa, Figure 2) has inserted between cysN2 and cysQ that may interrupt the coordinated regulation of the cysDNQ operon. The significance of this observation is not known.

Upstream of cysD1 and cysN1 in the A. ferrooxidans genome there are candidate genes for cysH, cysI and cysJ organized in an operon-like structure (Figure 2). Many microorganisms, such as E. coli and S. typhimurium, have a similar gene organization (1). cysI and cysJ encode a heterodimeric sulfite reductase with the form α4β8. This catalyzes the reduction of sulfite to sulfide which s the final reduction state of sulfur for its assimilation and the putative cysI and cysJ in A. ferrooxidans can be assigned with reasonable confidence.

The role of cysH in A. ferrooxidans is more enigmatic. In E. coli, cysH encodes a PAPS reductase without APS reductase activity, but in the plant A. thaliana the cysH product is exclusively an APS reductase. In the bacteria Pseudomonas putida and Mycobacterium spp., cysH exhibits both APS and PAPS reductase activity, but with differing levels of activity (9). The A. ferrooxidans cysH ortholog exhibits a motif characteristic of APS reductase (data not shown) (10), suggesting that it is has at least an APS reductase activity but leaves open, for the present, the possibility that it also has a PAPS reductase activity. For this reason we have drawn in Figure 1 a solid line connecting APS reduction with cysH, signifying a reasonable level of confidence in the assignment, and we have placed a dotted line connecting PAPS reduction with cysH to emphasize that some ambiguity remains regarding cysH function.

The final step in sulfur assimilation is the synthesis of L-cysteine from O-acetyl-L-serine and sulfide. This reaction is catalyzed by one or both of two O-acetyl-L-serine (thiol)-lyase isozymes, designated A and B and encoded by cysK and cysM genes in S. typhimurium and E. coli (1). A. ferrooxidans has a potential cysM ortholog that is predicted to accomplish this final step in sulfur assimilation (Table 1, Figures 1 and 2).


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Figure 2. Examples of two operon-like organizations of the candidate genes of A. ferrooxidans postulated to be involved in sulfate uptake and sulfur assimilation. The proposed function of the various genes designated with prefix "cys" can be found in Figure 1. tpa = probable transposase. hyp = hypothetical gene of unknown function

Figure 3. Comparison between the organization and function of the genes found in the cysDNC operon of E. coli with similar genes of A. ferrooxidans. In two instances in A. ferrooxidans a cysN gene (N2 and N3), potentially encoding a sulfurylase subunit, appears to be fused with a gene with significant similarity to the cysC gene of E. coli encoding an APS activity

In addition to exhibiting potential orthologs of genes involved in sulfur assimilation, A. ferrooxidans also appears to have characteristic mechanisms for the regulation of the relevant genes and enzymes. In plants and enteric bacteria the activity of the sulfur uptake and assimilation genes is regulated by a combination of feedback inhibition of serine transacetylase by cysteine and a gene regulatory system called the cysteine regulon (1). Excess cysteine can bind to acetylserine transferase reducing the activity of this enzyme, which, in turn, limits the production of O-acetyl-L serine. Since this latter compound is a precursor of cysteine its reduction causes a decrease of the synthesis of cysteine. Since A. ferrooxidans appears to have the gene encoding acetylserine transferase it can be assumed that it is subject to a similar feedback inhibition mechanism.

On the other hand, a reduction in the availability of sulfide leads to the accumulation of O-acetylserine, which spontaneously isomerizes to N-acetylserine. This latter compound binds to cysB and the resulting complex serves as an activator of several genes involved in sulfate uptake and sulfur assimilation (11, 12). cysB is a transcriptional regulator that belongs to the lysR family of transcriptional activators. A ferrooxidans has


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two copies of this gene both of which display the characteristic helix-turn-helix motif. of this family. Binding of N-acetylserine to cysB promotes the activation of genes involved in sulfate uptake and sulfur assimilation (1).

A third mechanism of regulation involves the control of levels of PAPS carried out by the cysQ gene product. Reports indicates that accumulation of PAPS is toxic for the cell (13). In A. ferrooxidans we found an ortholog of cysQ associated in an operon-like organization with cysD2 and cysN2 genes (Figure 2). A similar organization can be detected in Brucella abortus by inspection of the Integrated Genomics Inc. database of genome sequences (http://wit.integratedgenomics.com/ WIT2/CGI).

3. CONCLUSIONS Genes and enzymes predicted to be responsible for sulfur uptake and assimilation in

A. ferrooxidans exhibit significant similarity, including conservation of key protein motifs, to those found in a number of other microorganisms. This permits a metabolic model to be constructed that suggests that sulfur assimilation in A. ferrooxidans exhibits substantial similarity to other organisms. The operon-like organization of these genes in A. ferrooxidans is also similar to that observed in other microorganisms. In addition, A. ferrooxidans has a repertoire of genes predicted to be involved in the regulation of sulfur assimilation and this characteristic, together with the operonic organization of the relevant genes, implies that key aspects of the genetic and biochemical regulation of sulfur assimilation are shared between A. ferrooxidans and other organisms.

However, two important differences between A. ferrooxidans and other organisms stand out. First, it appears that A. ferrooxidans may have encoded enzymes that permit sulfur to be activated by both APS and PAPS intermediates, a characteristic that is shared by some but not all organisms. Secondly, compared to most organisms, A. ferrooxidans exhibits a probable gene fusion event in which an ATP sulfurylase-encoding gene has been joined to a gene encoding an APS kinase activity. The recognition and analysis of gene fusion events are becoming increasingly important for building models of gene evolution. In addition, gene fusions, known as Rosetta stone sequences, serve as powerful predictors of gene and protein function in bionformatic analyses (14).

The model of sulfur assimilation shown in Figure 1 was inferred by a range of bioinformatic inferences, including gene similarity comparisons, detection of possible operons and identification of putative genes involved in the regulation of sulfur assimilation. However, the model requires experimental validation, some of which is now underway in our laboratory.

In addition to sulfur assimilation, A. ferrooxidans uses sulfur as an energy and electron source. Several laboratories are actively engaged in deducing the transduction of sulfur to energy in A. ferrooxidans and a comprehensive model of sulfur utilization this organism will require an understanding of the metabolic and regulatory connections between the two pathways.

REFERENCES 1. Kredich, N. M. Biosynthesis of cysteine, Escherichia coli and Salmonella

typhimurium: cellular and molecular biology, In F. C. Neidhardt, J. L. Ingraham, K.B. Low, B. Magasanic, C. M. Schaechter, and H. E. Umbarger (ed.). vol. 1, pgs. 419-428. (1987).

2. Dénarié, J., Debellé, F, and Rosemberg, C. Annu. Rev. Microbiol. 46, pgs. 497-531. (1992).


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3. Williams, S.J., Senaratne, R. H., Mougous, J. D., Riley, L. W., Bertozzi, C. R. J. Bacteriol. 277, pgs.32606-32615.(2002)

4. Selkov, E., Overbeek, R., Kogan, Y., Chu, L., Vonstein, V., Holmes, D., Silver, S., Haselkorn, R., Fonstein, M. PNAS. 7, pgs. 3509-3514. (2000).

5. Barreto, M., R. Quatrini, S. Bueno, C. Arriagada, J. Valdes, S. Silver, E. Jedlicki and D. S. Holmes. (In press, Hydrometallurgy, 2003).

6. Leyh, T. S., Taylor, J. C., Markman, G. D. J. Biol. Chem. 263, pgs. 2409-2416. (1988) 7. Schwedock, J. S., Liu, C., Leyh, T. S., Long, S. R. J. Bacteriol. 176, pgs. 7055-7064.

(1994). 8. Fry,I. J., and E. Garcia. In J. Salley, R. G. L. McCready, and P. L. Wichlacz (ed.),

Biohydrometallurgy. pgs. 171-185. (1989). 9. Bick, J. A., Dennis, J. J., Zylstra, G. J., Nowack J, Leustek T. J. Bacteriol. 182,

pgs.135-142. (2000). 10. Kopriva, S., Buchert, T., Fritz, G., Suter, M., Benda, R., Schunemann, V., Koprivova,

A., Schurmann, P., Trautwein, AX., Kroneck, PM., Brunold, C. J. Biol. Chem. 277. pgs. 21786-21791. (2002).

11. Hryniewicz, M. M., and Kredich, N. M. J. Bacteriol. 173, 5876-5886. (1991). 12. Monroe, R.S., Ostrwski, J., Hryniewicz, M. M., Kredich, N. M. 172, pgs.6919-6929.

(1990). 13. Neuwald, A. F., Krishnan, B. R., Brikum, I., Kulakauskas, S., Suziedelis, K.,

Tomcsanyi, T., Leyh, T. S., Berg, D.E. J. Bacteriol. 174, pgs. 415-425. (1992). 14. Marcotte, E. M., Pellegrini, M., Ng HL., Rice, D. W., Yeates, T. O., Eisenberg, D.

Science.285, pgs.751-3. (1999).



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Survival of acidophilic bacteria under conditions of substrate depletion that occur during culture storage

D. Barrie Johnsona, Debby F. Bruhnb and Francisco F. Robertob a School of Biological Sciences, University of Wales, Bangor, LL57 2UW, U.K.

b Lockheed Idaho Technologies Co., Idaho National Engineering and Environmental Laboratory, P.O.Box 1625, Idaho Falls, Idaho, 83415-22032

Abstract Survival of four mesophilic acidophilic bacteria (two chemolithotrophs,

Acidithiobacillus ferrooxidans and Leptospirillum ferrooxidans, and two heterotrophs, "Ferrimicrobium acidiphilum" and Acidiphilium SJH) was investigated during 500 days of starvation stress. Counts of plateable At. ferrooxidans remained fairly stable for 25 days following completion of ferrous iron oxidation, after which they declined; however, 2 x 104 viable cells/ml stored culture were still present after 500 days of starvation stress. In contrast, numbers of viable L. ferrooxidans went into immediate decline following iron oxidation, and no viable planktonic cells were detected 45 days after iron oxidation. Pure cultures of the two heterotrophic acidophiles displayed similar trends to At. ferrooxidans. When mixed populations of acidophilic mesophiles were subjected to long-term starvation stress, mortality rates of At. ferrooxidans and Acidiphilium SJH were greater, and those of L. ferrooxidans and "F. acidiphilum" less than in corresponding pure cultures. The significance of metabolic diversity, sensitivities to organic compounds and scavenging potential on the survival of acidophilic bacteria is discussed.

Keywords: acidophilic bacteria; mixed cultures; pH stress; starvation stress

1. INTRODUCTION Bacteria that inhabit acidic, metal-rich environments are physiologically diverse, in

their carbon assimilation, temperature-response characteristics and abilities to catalyze the dissimilatory transformations of iron and sulfur [1]. Interest in this group stems primarily from their role in the genesis of acid mine drainage [2] and in their exploitation in mineral processing operations [3]. Besides being characterized by low pH (generally 2-4), environments populated by acidophilic microorganisms generally contain large concentrations of sulfate, iron, other (heavy) metals and metalloids (depending on the geochemistry of the local environment), and small concentrations of dissolved organic carbon. Bacteria such as the iron-oxidizing chemolithotrophs Acidithiobacillus ferrooxidans and Leptospirillum spp. are obligate acidophiles, and are inactive in neutral and mildly acidic environments. A variety of interactions between acidophilic microorganisms have been described [4]. For example, mutualistic associations occur between autotrophic and heterotrophic acidophiles, whereby utilization of organic materials (e.g. leakage products from autotrophic cells) by the latter "detoxify" the


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environment for acidophilic autotrophs, which are often highly sensitive to small molecular weight organic compounds.

The effects of nutrient stress and substrate depletion on the survival of bacteria are of fundamental importance in microbial ecology. Long-term survival in environments that are depleted in nutrients, and which also may be stressful in other ways (e.g. pH, temperature), may result from bacteria having specific survival strategies. Although acidophilic bacteria are important in ore leaching and genesis of acid mine drainage pollution, relatively little has been published on their abilities to survive starvation stress. Zychlinsky and Matin [5] found that survival of the acidophilic mixotroph Acidiphilium acidophilum during "long-term starvation" (up to 22 days) was similar to that reported for neutrophilic bacteria. In this paper, we describe the responses of four mesophilic acidophiles to starvation stress for periods of up to 500 days.


2.1 Bacteria Survival of four mesophilic acidophiles was investigated: (i) At. ferrooxidans (ATCC

23270; the type strain); L. ferrooxidans (strain CF12, isolated from the Noranda Blackbird-cobalt mine, Cobalt, Idaho); (iii) "Ferrimicrobium acidiphilum" (strain T-23 [1]); (iv) Acidiphilium SJH [1]. Two of these bacteria are autotrophs (At. ferrooxidans and L. ferrooxidans) and two are heterotrophs ("F. acidiphilum" and Acidiphilium SJH). Three species (At. ferrooxidans, L. ferrooxidans and "F. acidiphilum") oxidize ferrous iron to ferric, and three (Acidiphilium SJH, F. acidophilum and At. ferrooxidans) can reduce ferric iron to ferrous in oxygen-limited environments.

2.2 Experimental protocols Pure cultures of each acidophile, and a mixed culture containing all four, were grown

in 20 mM ferrous sulfate liquid medium, poised initially at pH 1.8 (to avoid precipitation of ferric compounds in oxidized cultures). Cultures (250 ml in 500 ml conical flasks) were incubated (shaken) for three days at 30°C, and subsequently stored at room temperature (23 ± 2°C) in the dark. Flasks were loosely covered with aluminum foil to reduce evaporation losses.

A second experiment was set up to monitor the effect of pH on the survival of mesophilic acidophiles under conditions of nutrient depletion. Individual cultures of the four bacteria were grown in appropriate liquid media (20 mM ferrous sulfate for At. ferrooxidans and L. ferrooxidans; 20 mM ferrous sulfate/0.02% (wt/vol) yeast extract for "F. acidiphilum"; 0.002% (wt/vol) yeast extract for Acidiphilium SJH). When cell numbers had reached approximately 107/ml, bacteria were harvested aseptically by centrifugation, washed, and re-suspended in acidified (pH 2.5) basal salts. Aliquots (0.5 ml) of cell suspensions were added, separately, to basal salts solutions (200 ml in 500 ml flasks) adjusted at either pH 2.0, 5.0 or 8.0, using dilute H2SO4 or NaOH. Cultures were incubated, unshaken, at room temperature in the dark. All experiments used duplicate cultures.

2.3 Analytical techniques Total numbers of bacteria were assessed by acridine orange direct counts (AODCs)

using a modification of the method of Hobbie et al. [6]. This involved filtering culture aliquots through black polycarbonate membranes (0.2 µm pore size; Nuclepore, U.S.A.),


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rinsing first with acidified (pH 1.0; H2SO4) distilled water and then with alkaline (pH 11.0; NaOH) water, fixing cells with 4% (v/v) glutaraldehyde (10 mins.) followed by a second pH 11.0 water rinse, staining with 0.01% (w/v) acridine orange solution (pH 11.0; 6 mins.) and a final rinse in distilled water. Counts of viable bacteria were made by plating diluted culture aliquots onto selective solid media [7]. These were ferrous iron overlay medium (Feo) for enumerating At. ferrooxidans, L. ferrooxidans and "F. acidiphilum", and yeast extract medium for Acidiphilium SJH. Identification of bacteria was made on the basis of differences in colony morphologies, combined with microscopic examination, as described previously [7]. To prepare samples for scanning electron microscopy (SEM), culture aliquots (containing 106-107 cells/ml) were filtered through Nuclepore membranes (0.2 µm pore size) and rinsed, sequentially, with 0.1 M H2SO4, acidified (pH 2.0) basal salts and "nanopure" water. After fixing with 5% (v/v) glutaraldehyde, cells were dehydrated by flushing with solutions of ethanol (30%, 60%, 90%, 100% (x2); all vol/vol) and critical point dried. Bacteria were viewed in an Amray Model 1830 SEM; cell dimensions were recorded and biovolumes calculated.

Other analyses included determination of ferrous iron concentrations (by titration with 1 mM KMnO4 in dilute H2SO4), dissolved organic carbon (samples pre-filtered through 0.2 µm membranes and DOC estimated using an OI Corporation Model 700 DOC analyzer), and pH using a glass electrode.

3. RESULTS

3.1 Survival trends in pure cultures of acidophiles The acidophilic bacteria used in these experiments varied widely in their abilities to

survive under conditions of substrate depletion. Their survival was affected by the presence of other bacteria (i.e. mixed versus pure cultures) and by environmental factors. Figure 1 shows how total counts of all four acidophiles remained fairly constant during the first 100 days of substrate depletion while, in contrast, counts of cultivatable bacteria differed markedly between isolates. Counts of plateable At. ferrooxidans remained fairly static for ca. 25 days following completion of iron oxidation (which took 3 days) after which time they declined, though significant numbers of cultivatable cells (ca. 2 x 104/ml) were still present some 500 days after the start of the experiment. In contrast, plate counts of the other iron-oxidizing chemolithotroph, L. ferrooxidans, declined dramatically following iron oxidation, and no cultivatable bacteria were recovered on day 45. However, cultures plated after 87 and 100 days did produce colonies (confirmed as L. ferrooxidans), though none was detected beyond these times. Viable counts of pure cultures of the two acidophilic heterotrophic isolates also displayed interesting differences. Plate counts of "F. acidiphilum" showed an initial decline in contrast to Acidiphilium SJH whose numbers increased during this time, though these stabilized from day 28 and, by the end of the experiment, viable counts of "F. acidiphilum" were similar to those of both At. ferrooxidans and Acidiphilium SJH. One difference between "F. acidiphilum" and the other iron-oxidizing acidophiles is the former's more protracted ferrous iron oxidation when cultured in "inorganic" media. Whereas iron oxidation was complete in cultures of At. ferrooxidans and L. ferrooxidans by the time that cultures were put into storage (3 days after inoculation), residual ferrous iron concentrations in cultures of "F. acidophilum" were 7 mM (after 9 days) and 3 mM (after 17 days). Slow ferrous iron oxidation is a feature of iron-oxidizing heterotrophs when grown in the absence of organic carbon (D.B. Johnson, unpublished data).


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Figure 1. Effect of starvation-stress on, (a) total bacterial numbers (AODCs) and (b) culturable bacteria. Symbols: ( ) At. ferrooxidans; ( ) L. ferrooxidans; ( ) F. acidophilum; ( ) Acidiphilium SJH. Error bars represent standard deviations

Table 1. Changes in cell dimensions* of acidophilic bacteria stored under conditions of substrate limitation

Three of the acidophiles displayed net decreases in cell biovolumes after 37 days

compared with those recorded at day 3 (Table 1). The single exception was Acidiphilium SJH, whose cells were longer and thinner at day 37 but which showed no net change in biovolume. Decrease in mean cell biovolume was smallest for At. ferrooxidans among the iron-oxidizing acidophiles. Net biovolume reductions were similar for L. ferrooxidans and "F. acidophilum" during this time; however, while L. ferrooxidans cells had shrunk by similar proportions in both length and width, those of "F. acidophilum" were much shorter than at day 3, though of similar width.


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3.2 Survival trends in mixed cultures of acidophiles Patterns of bacterial survival were subtly different when the acidophiles were grown

in mixed culture (Fig. 2). In particular, numbers of plateable L. ferrooxidans declined at a slower rate than was observed in pure cultures, and plate counts of "F.acidophilum" increased (rather than decreased) during the first 25-30 days of culture storage. In contrast, numbers of viable At. ferrooxidans and Acidiphilium SJH went into more immediate decline than in pure cultures of these acidophiles. As with pure cultures, after 500 days incubation viable cells of all acidophiles (except L. ferrooxidans) were detected in mixed cultures, with "F. acidiphilum" being the most abundant viable acidophile.

Figure 2. Comparison of plate counts of acidophiles, grown in pure and mixed cultures. Data shown are for changes in populations that occurred during the first 45 days of starvation-stress. Error bars represent standard deviations. (a) At. ferrooxidans; (b) L. ferrooxidans; (c) `F. acidiphilum’; (d) Acidiphilium SJH

3.3 Survival trends of acidophiles at different pH The pH of the substrate-depleted media in which cultures of bacteria were maintained

had a major influence on the survival rates of the mesophilic acidophiles. Because of the poor buffering capacity of the basal salts solutions, the addition of the bacterial suspensions caused a shift in pH in all cases, resulting in pH values of 2.15, 4.10 and 6.20 for media adjusted initially to pH 2.0, 5.0 and 8.0, respectively. Subsequent changes in pH in these cultures were minor (<0.1 pH unit). Concentrations of dissolved organic carbon (ca. 4-5 mg/L at the start of the experiment) were about 50% of levels recorded in the first experiment since, in the latter, the culture liquor was spent bacterial medium (which would have accumulated cell leakage products etc.) rather than basal salts solution. Changes in viable bacterial counts recorded in this experiment are shown in Fig. 3. No colonies of L. ferrooxidans were recovered from any culture in this experiment, indicating either that the methodology used in harvesting cells of this acidophile proved lethal, or


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that L. ferrooxidans lost viability very rapidly in the basal salts solutions. At pH 2.15, the overall trends (Fig. 3a) were similar to those found in experiment 1. Viable counts of "F. acidiphilum" exceeded those of other bacteria from day 14 onwards, and plate counts of this heterotrophic iron-oxidizer remained fairly stable throughout the 100 days of culture storage. Viability of At. ferrooxidans fell sharply during the first 40 days of culture storage, but numbers stabilized at about 104 cells/ml by day 100. Counts of viable Acidiphilium SJH, in contrast, were fairly stable for the first 40 days of storage, but then went into severe decline, and no viable cells were detected at day 100. At pH 4.1, patterns of survival of the three acidophiles were quite different, with both At. ferrooxidans and Acidiphilium SJH displaying more prolonged cell viability than at the lower pH; in contrast, plate counts of "F. acidiphilum" declined rapidly and from day 72 onwards no viable cells of this acidophile were detected (Fig. 3b). At the highest pH (6.2), the mortality rate of "F. acidiphilum" was even more pronounced. At. ferrooxidans also died out more rapidly at this pH than at either pH 2.15 or 4.1, but Acidiphilium SJH survived in reasonable numbers (ca. 103-104/ml) throughout the period of culture storage (Fig. 3c).

4. DISCUSSION Microorganisms have developed a variety of strategies for surviving times of

depletion of substrate or of a particular essential nutrient [8]. This may involve the possession and expression of "survival genes", cell proliferation, and reduction in cell size (producing "ultramicrocells" of <0.3 µm diameter). Various methods have been used to differentiate between viable and non-viable cells in a population [9]. The use of plate counts to enumerate viable cells has been criticized, since bacteria may be able to take up and metabolize nutrients without resuming growth and division under the imposed conditions of plating. Oliver [10] described "viable but not culturable" cells as those that can be demonstrated to be metabolically active but which are incapable of sustaining cellular division in media that normally support growth. The solid media and experimental techniques used in the present study have been developed over several years specifically to produce high plating efficiencies of acidophilic bacteria from environmental materials (as well as to facilitate identification of isolates). Numerical estimations of iron-oxidizing acidophiles in laboratory cultures and environmental samples (where bacteria are frequently in a state of substrate limitation) as determined using "overlay" plates have been shown to be similar or greater than the "Most Probable Number" approach [7]; the latter uses liquid media that might be considered to be more conducive to the growth of stressed cells. It is worth noting that the stored cultures that produced zero plate counts of L. ferrooxidans also proved negative (i.e. no iron oxidation) when samples were subcultured in liquid media. Oliver [10] found that "starved" cells responded quickly to the reversal of nutrient limitation whereas the response in viable non-culturable cells was slow; zero response, as described for L. ferrooxidans, indicates that cells were indeed non-viable.

Trends of starvation-survival displayed by the four species of mesophilic acidophiles used in the present work were often very different. The chemolithotrophic iron-oxidizer, L. ferrooxidans displayed much more limited longevity than At. ferrooxidans. This may be due either to an inherently greater capacity of At. ferrooxidans to survive in the absence of extraneous substrates (using storage reserves etc.) or to its greater metabolic versatility. Ferrous iron is the only substrate known to be utilized by L. ferrooxidans; whilst it is possible that some ferrous iron might have been regenerated during culture storage, this is highly unlikely. Even though both Acidiphilium SJH and "F. acidiphilum" can couple the oxidation of organic substrates to the reduction of ferric iron [4] this is generally only


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Figure 3. Effect of pH on long-term survival of mesophilic acidophiles. Symbols: ( ) At. ferrooxidans; ( ) "F. acidiphilum"; ( ) Acidiphilium SJH. Error bars represent standard deviations observed under conditions of limited dissolved oxygen concentrations, for example, in cultures containing extraneous organic materials. Since no reduced sulfur compounds


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were present in the spent media or the basal salts suspension, the only other alternative potential substrates would have been organic. These would have arisen from leakage from active cells, from lysis from dead and dying cells, and as atmospheric inputs. While At. ferrooxidans is often considered to be an obligate autotroph, its ability to incorporate methionine [11] indicates that it has at least some capacity to utilize organic carbon compounds. Further evidence that the survival of At. ferrooxidans was at least in part due to heterotrophic metabolism came from comparing its longevity in pure and mixed cultures. More limited survival in mixed cultures may have been due to reduced availability of organic substrates, resulting from competition with the two acidophilic heterotrophs present. In contrast, the greater sensitivity of L. ferrooxidans to dissolved carbon compounds [12] combined with its inability to utilize organic materials would imply that heterotrophic activity would have a net beneficial effect on survival of this iron-oxidizer, as was observed.

The reason for the "re-appearance" of colonies of L. ferrooxidans at days 87 and 100 is unclear, but may relate to the break-up of cell aggregates. L. ferrooxidans is known to form flocs consisting of cells enmeshed in extracellular polymers [13] in liquid media. These would have not been enumerated by direct or plate counts, under the protocol used. It is possible that these entrapped cells have greater longevity that free-swimming cells; disintegration of flocs during protracted starvation stress may have resulted in the release of viable planktonic cells, which again tended to die out relatively rapidly.

The relative survival of the two heterotrophs, Acidiphilium SJH and "F. acidiphilum", was also affected by competition for the limited organic resources available. Acidiphilium SJH is a particularly adept scavenger in pure cultures, but was out-competed by the iron-oxidizing heterotroph "F. acidiphilum" in highly acidic mixed cultures. Indeed, the greater capacity of "F. acidiphilum" to survive starvation stress than the other mesophilic acidophiles tested suggests that these bacteria may be better adapted to situations where substrate depletion is a common phenomenon, and may also be a reason for the dominance of such bacteria in some environments. However, mortality rates of "F. acidiphilum" in moderately acidic and near-neutral solutions were much greater than in media poised at about pH 2. The resulting reduction in competition for available organic materials was considered to be a major reason for the enhanced survival of both At. ferrooxidans and Acidiphilium SJH in mixed cultures at pH 4.10 than in those poised at pH 2.15. However, in solutions of pH 6.10, the combined stresses of starvation and inhospitable pH resulted in more limited survival of all acidophilic bacteria than in more acidic cultures. It is surprising that near-neutral external pH should have lessened the survival of acidophiles, as the internal pH of these bacteria is close to neutral in active cells, and 2-3 pH units above external pH in starved cells [14]. It is known that the maintenance of ∆pH values in resting acidophiles is passive rather than active.

The trends identified in this study have highlighted the ability of acidophilic bacteria to survive long-term starvation stress, though surviving populations may be very different in composition to those present at the onset of nutrient depletion. Environmental factors appear to have an important influence on survival of these bacteria. Metabolic versatility appears to be one reason for the superior survival of some species of acidophiles. Evidence for survival strategy based on the formation of viable ultramicrocells was limited to the heterotrophic iron-oxidizer "F. acidiphilum". The results show that At. ferrooxidans, Acidiphilium SJH and "F. acidiphilum" can survive in significant numbers for up to at least 500 days in substrate-depleted situations, though L. ferrooxidans has a far lesser capacity to survive under the conditions imposed in our experiments. A limitation of this work was that only one strain of each acidophile was used in experimental work. It is


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conceivable that considerable variation also exists between strains of a particular acidophilic species, and between species in the case of Acidiphilium.

ACKNOWLEDGEMENTS This work was supported by funding provided by the U.S. Department of Energy,

Office of Science, to the Idaho National Engineering and Environmental Laboratory, operated by Bechtel BWXT Idaho, LLC, under contract DE-AC07099-ID13727. In addition, DBJ was supported during the course of this work by a Faculty Fellowship program sponsored by the U.S. Department of Energy, Office of University and Science Education Programs.

REFERENCES 1. K.B. Hallberg and D.B. Johnson. Adv. Appl. Microbiol. 49 (2001) 37. 2. D.B. Johnson. Water, Air & Soil Pollut. Foc. 3 (2003) 47. 3. D.E. Rawlings. Annu. Rev. Microbiol. 56 (2002) 65. 4. D. B. Johnson. FEMS Microbiol. Ecol. 27 (1998) 307. 5. E. Zychlinsky and A. Matin. J Bacteriol. 153 (1983) 371. 6. A.T. Hobbie, J.R. Daley and S. Jaspar. Appl. Environ. Microbiol. 33 (1977) 1225. 7. D.B. Johnson. J. Microbiol. Meth. 23 (1995) 205. 8. S. Kjelleberg. (ed.) Starvation in bacteria. Plenum Press, New York, 1993. 9. D.A. Siegele, M. Almiron and R. Kolter. p.151. In: S. Kjelleberg (ed.), Starvationin

bacteria. Plenum Press, New York, 1993. 10. D.J. Oliver. p.239. In: S. Kjelleberg (ed.), Starvation in bacteria. Plenum Press,

NewYork, 1993. 11. D.J. Oliver and J.K. VanSlyke. Arch. Biochem. Biophys. 263 (1988) 369. 12. D.B. Johnson, M.F. Said, M.A. Ghauri and S. McGinness. p. 403. In: J. Salley, R.G.L.

McCready and P.L. Wichlacz (eds.), Biohydrometallurgy 1989. Canmet, Ottawa, 1990.

13. R. Hallmann, A. Friedrich, H-P. Koops, A. Pommerening-Roser, K. Rohde, C. Zenneck and W. Sand. Geomicrobiol. J. 10 (1992) 193.

14. P.R. Norris and W.J. Ingledew. p. 115. In R.A. Herbert and R.J. Sharp (eds.), Molecular biology and biotechnology of extremophiles. Blackie, Glasgow, 1992.



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Synthesis of nanophase hydroxyapatite by Serratia sp. N14

P. Yonga, R.L. Sammonsb, P.M. Marquisb, H. Luggb and L.E. Macaskiea* a Schools of Biosciences, The University of Birmingham, Birmingham B15 2TT, UK

b School of Dentistry, The University of Birmingham, Birmingham B15 2TT, UK

Abstract A species of Serratia isolated from heavy metal contaminated soil and previously

identified as Citrobacter sp. N14 expresses an atypically high level of acid phosphatase. In the presence of CaCl2 and glycerol-2-phosphate (G2P), biosynthesis of nano-crystalline hydroxyapatite (HA) was achieved by use of immobilised Serratia cells. Calcium ions accumulated as cell-bound calcium phosphates utilizing phosphate released by the enzymatic cleavage of the substrate (G2P). Addition of citrate (to restrict the free Ca2+ concentration) and/or reducing the substrate concentration reduced the size of the HA crystals and promote formation of the nano-phase. Hydroxyapatite crystals were located using Scanning Electron Microscopy (SEM) and identified by X-ray diffraction (XRD). The size of the crystals was calculated to be between 15 and 25 nm, a size more suitable for biomedical applications than that obtained by chemical synthesis.

Keywords: biomineralization, calcium phosphate, hydroxyapatite, Serratia

1. INTRODUCTION Hydroxyapatite (HA) is a well-established bone substitute biomaterial, which has

been successfully used for replacement and augmentation of diseased or traumatised bone [1]. The biocompatibility of this material is related to its close similarity to the composition of the mineral component of natural bone. Hydroxyapatite can also be used to coat metal prostheses, and for orthopaedic and dental applications, using techniques such as plasma spraying. Unfortunately the coating methods currently employed lead to coatings containing a variety of phosphate phases, both amorphous and crystalline, which can be susceptible to dissolution or delamination in clinical use [2,3]. Equivalent problems are also encountered in the fabrication of dense hydroxyapatite components, where difficulties in densification inevitably lead to implants with poor mechanical properties, notably strength and fracture toughness, which cannot be used in load bearing applications. It is well known that as the grain size of materials reduces towards the nanoscale (<200 nm, ideally below 50 nm), properties such as strength are dramatically improved [4]. To achieve this goal requires precursor materials of this nanoscale dimension, which can be subsequently manipulated and fabricated without loss of the fine grain size. Conventional chemical synthesis approaches do not produce particles of the

* To whom correspondence should be addressed (Fax: +44(0)121-414-5925; e-mail: [email protected])


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required dimensions and also invariably lead to clusters, or agglomerates of particles, which lead to inherent weakness in the final implant.

Microbially mediated biomineralization has a high potential for the production of metallic nano-particles. Klaus [5] found that Pseudomonas stutzeri extracted silver from ores and deposited pure metallic silver particles of up to 200 nm diameter inside the bacteria. Palladium (II) can be reduced enzymatically to Pd(0) at the surface of Desulfovibrio desulfuricans [6,7]; the mean crystal size of biologically-manufactured Pd(0) was only half of that obtained by reduction of soluble Pd(II) under H2 without cells, but it is not clear which biological features contributed to this phenomenon [7,8]. In biomineralization, the components of microbial cell walls and their associated structural polymers may be used as organic "scaffolds" for the patterning of extracellular vesicles and associated inorganic minerals [9]. The compartmentalisation of the biological space on and around cells could be utilised to promote spatial localisation for crystal growth, and nucleation could be mediated by organic polymeric substrates in or on the spatial boundary. The inorganic mineral crystals could be in membrane-bound vesicles, in the mucilaginous layers of cell walls or impregnated in biopolymers in the extra-cellular space [9].

A biomineralization method was developed for the accumulation of metal ions from aqueous solution utilizing cells of a Serratia sp., which was originally classified as Citrobacter [10,11]. It was recently reassigned as Serratia on the basis of biochemical and physiological criteria and 16S rRNA homology [12]. In this bioaccumulation process, metal uptake is mediated by the activity of phosphatase enzyme that is localised both periplasmically or within extracellular polymeric materials [13]. This enzyme continues to function in resting cells to liberate phosphate ions via the hydrolytic cleavage of an organic phosphate molecule (e.g. glycerol 2-phosphate (G2P)) and promotes stoichiometric deposition of cell-bound metal phosphate [13]. Early studies showed that under appropriate conditions (high pH) strontium could be desolubilized using this method [14]. The biomineralization of calcium phosphate should occur in similar manner and the purpose of this study was to evaluate the potential of this approach for the microbially mediated synthesis of calcium phosphate. The investigation focused on the biosynthesis of nanocrystalline hydroxyapatite (HA) with the objective of producing a better HA biomineral for biomedical application. The effect of the challenge solution on metal phosphate deposition, the chemical structure and the size of crystals, and their arrangement on the biomass, was investigated.

2. EXPERIMENTAL

2.1 Preparation of biomass Biomass was grown under carbon (lactose)-limited continuous culture at 30°C in an

airlift fermenter (Fig. 1) as described previously [15]. Minimal medium (2.5 l) [15] was added to the fermenter containing up to 200 polyurethane reticulated foam cubes (TM30 supplied by Recticel, Belgium, washed with distilled water three times) threaded on cotton strings, and the culture was started by the addition of 50 ml inoculum which had been pre-grown in similar medium. The culture was allowed to grow batch-wise and switched to continuous mode after 24 h (OD600 ~ 0.6), maintained at steady-state for six days [15], and the foam cubes (heavily loaded with biofilm) were withdrawn and kept in air in a sealed vessel over isotonic saline (8.5 g/l) at 4°C until use. The culture was monitored by measurement of phosphatase activities and cell densities of free cells in the culture outflow [15].


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2.2 Phosphatase activity assay Phosphatase activity was assayed by the release of p-nitrophenol (PNP) from p-

nitrophenyl phosphate (PNPP) [14,15]. Phosphatase activity (units) is defined as nmol of PNP liberated per min per mg bacterial protein, with OD600 (cell optical density at 600 nm) of the cells converted to bacterial protein using a conversion factor obtained by the assay method of Lowry (0.278 mg protein/ml at OD600 =1; path length = 1 cm) [16].

2.3 Bioaccumulation of calcium phosphate by Serratia sp. N14 Ten biofilm-loaded foam cubes (~4.5 mg dry cell mass per cube, washed by dipping

into isotonic saline) were challenged in 50 ml of 50 mM TAPSO/NaOH buffer pH 9.2 with CaCl2 (1 mM or 5 mM) and sodium glycerol 2-phosphate (G2P: 1 mM or 5 mM) with or without citrate (2 mM). The flask was shaken at room temperature. Additional CaCl2, G2P and citrate were dosed into the flask daily (additional final concentrations of 1 mM, 5 mM and 2 mM as appropriate, after allowing for volume reduction after sampling). The dosing regime is summarised in Table 1. Samples were taken daily before dosing for monitoring of residual Ca(II) and phosphate concentrations. Calcium phosphate-loaded cells harvested by centrifugation after eight days of dosing (or as otherwise specified) were treated and analysed by X-ray powder diffraction and electron microscopy as described. Representative data are shown throughout.

2.4 Assay of phosphate and calcium Samples (up to 5 mM Pi) were diluted 30-fold. Sodium molybdate (0.6 ml; 2.5%

(w/v) in 1.67 M H2SO4 was added into 1 ml assay solution and phosphate was visualized by adding 0.4 ml SnCl2. (Fresh solution of SnCl2 was made from 0.25 ml stock solution diluted into 100 ml of 1 M HCl, and the stock was made monthly by dissolving 1.5 g SnCl2 in 2.5 ml concentrated HCl.) The quantity of resultant blue complex was measured at 720 nm (Perkin-Elmer spectrophotometer) and the phosphate concentration estimated by reference to a standard curve, similarly prepared. For assay of Ca(II) 0.1 ml of sample (up to 1 mM Ca(II)) was added to 1.9 ml 20 mM MOPS/NaOH buffer pH 7, and mixed, and Ca(II) was visualized by the addition of 0.1 ml of 0.15% (w/v; aq.) Arsenazo III, with estimation of the pink complex at A631 nm (Perkin-Elmer spectrophotometer) versus a standard solution of CaCl2.

2.5 Localization of cell-bound deposits using scanning electron microscopy Calcium phosphate loaded cells (squeezed from one foam cube) were centrifuged,

fixed with 2.5% glutaradehyde fixative (in 0.1 M Na cacodylate/HCl buffer pH 5.2) at 4°C for one hour, dehydrated at room temperature, treated and examined under a scanning electron microscope as described previously [6].

2.6 Identification of cell bond material and crystal size determination by X-ray powder diffraction (XRD) analysis The Ca-loaded cells (squeezed from 10 foam cubes) were centrifuged after eight days

of challenge with the same daily addition, washed with distilled water and acetone, and then dried at room temperature. The samples were laid on a silicon crystal plate and analysed using a high precision X-ray powder diffractometer (School of Physics, The University of Birmingham) for data collection as described previously [7,11]. The powder diffraction patterns were recorded from 5° to 70° (2θ) with a step length of 0.05° (2θ). The crystal size of cell-bound calcium phosphate/HA was calculated for samples obtained


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from loaded dried biomass or from commercial reference material HA (Capital 60 hydroxylapatite: Plasma Biotal Ltd., Tideswell, Derbyshire, UK) by computing the diffraction pattern using the Langford programme [17].

3. RESULTS AND DISCUSSION When the cells are challenged with calcium and a phosphatase substrate i.e. G2P,

there is a biocatalytic (enzymically-mediated) and an inorganic precipitation combined process, in which the component reactions are: Glycerol-2-phosphate Phosphatase> Glycerol + Phosphate (1) Phosphate + Calcium ____> Calcium phosphate (2)

The synthesis of calcium phosphate would occur where the Ca(II) ions interact with the liberated precipitant, phosphate ions. Calcium phosphate can be synthesised chemically, and many authors have studied the factors of affecting this synthesis [1,2,18]. The type of calcium phosphate and the crystal size are largely dependent on the conditions of synthesis, i.e. the setting time, solution pH, ions concentrations, temperature, etc. Indeed, the bioprecipitation of calcium phosphate by Serratia cells would be affected similarly by these physicochemical factors. The form of phosphate ions released by reaction (1) is pH-dependent. Phosphate is a tri-basic ion (K1 = 7.5 x 10-3, K2 = 6.2 x 10-8, K3 = 5.0 x 10-13) [19]. The major forms of which in aqueous solution would be H3PO4 and H2PO4

- below pH 4.6, H2PO4- and HPO4

2- from pH 4.6 to 9.7 and HPO42- and PO4

3- for pH values above 9.7. In addition, the rate of phosphate release in reaction (1) is dependent on the solution pH since phosphatase activity is pH dependent. Phosphatase activity of the cells was reduced by 70% when the solution pH was changed to 9.0 from 5.5 (Fig. 1a). However, after 24 h the inorganic phosphate concentrations were not significantly different in solutions at pH 7 to 9.2 although the phosphate released was lower in the solution at pH 10.0 (Fig. 1b) attributable to the lower reaction rate. The observation that phosphate release was >95% complete at 24 h (Fig.1b) justified the choice of daily intervals for the dosing regime. The requirement for alkaline pH was shown previously in the case of Sr(II) removal [14] where a compromise was required between the requirement for high pH for precipitation of the alkaline earth metal phosphate and retention of high cellular phosphatase activity.

Many authors have studied the solubility of calcium phosphate salts [1,20,21]. The type of calcium phosphate precipitated would be dependent on the pH value of the challenge solution [20-22]. It has been reported that hydroxyapatite (Ca5(PO4)3(OH)) is the most stable calcium phosphate salt at pH>4, but it precipitates directly only at above pH 8 [20,21]. The current investigation was carried out to define the conditions for the biosynthesis of HA for potential medical applications; the quality of the calcium phosphate crystals is much more important than the overall yield. Calcium bioprecipitation by Serratia was tested in a solution of 50 mM TAPSO/NaOH buffer (pH 9.2) and 1 mM CaCl2, with phosphatase substrate G2P at saturating [23] (5 mM) or limiting (1 mM) concentrations using biofilm-immobilized cells.

The biomineralization process (controlled by enzyme-mediated inorganic phosphate (Pi) concentration), relies on the cellular microenvironment becoming supersaturated with inorganic phosphate released from G2P by the cells. However, nucleation represents an activation energy barrier to the spontaneous formation of a solid phase from a supersaturated solution. Increasing the degree of super-saturation and lowering the interfacial energy can reduce the activation energy for nucleation [24]. Indeed, for 1 mM free Ca2+ in the bulk solution without complexing reagent, it was calculated that


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precipitation should occur for CaHPO4 at 0.1 mM HPO4- (Ksp=1x10-7 for CaHPO4) [25],

and for Ca3(PO4)2 at 1.4x10-7 mM PO4- (Ksp= 2.0x10-29 for Ca3(PO4)2) [25]. However,

biosorption of calcium was observed only after 24 h even in the solution with 5 mM G2P and a released phosphate concentration of higher than 4 mM (Fig.2a.). In the solution with 1 mM substrate the initiation of precipitation was retarded even further (Fig.2b). When the cells were challenged in the presence of citrate, the concentration of free Ca(II) ions was reduced significantly because citrate is a calcium complexing reagent (K = 4.79x104) [25]. By calculation, there would be only 0.02 mM free ions for Ca(II) at a input concentration of 1mM in the solution with 2 mM citrate, which, in turn, would need a phosphate concentration above 5 mM to form calcium phosphate precipitate. It was shown that the initial precipitation of calcium phosphate from the solution with 2 mM citrate was the slowest (Fig. 2c) compared to that from the other two solutions. However once the precipitation started, the rate of calcium uptake was the same in 1mM as in 5 mM G2P solution with or without citrate, although there was a much lower phosphate or free calcium ion concentration in the latter two solutions (Fig.2b, 2c). Therefore, for the formation of calcium phosphate, nucleation was the controlling stage initially. Once the barrier of nucleation was overcome, continued precipitation of calcium phosphate became feasible. Even in a solution with 1 mM CaCl2, 2 mM citrate and 1 mM G2P (added daily), the precipitation was completed within eight days (not shown).

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Figure 1b. The effect of pH on inorganic phosphate (Pi) release by immobilized Serratia cells. Five foam cubes were challenged in 25 ml 50 mM buffer (pH 7-9.2: TAPSO, and pH 10 AMPSO) 5 mM G2P and 2 mM citrate. Samples were taken after 24 hours and centrifuged and the phosphate content of supernatants assayed as described in Experimental


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Figure 2. Phosphate release and calcium accumulation by immobilized Serratia cells Ten biofilm-foam cubes (~4.5 mg dry cell mass per cube; phosphatase activity was 2834 units at harvest) were challenged in 50 ml of 50 mM TAPSO/NaOH buffer (pH 9.2), and: a) 1 mM CaCl2 and 5 mM G2P with same addition daily; b) 1 mM CaCl2 and 1 mM G2P with same addition daily; c) 1 mM CaCl2, 5 mM G2P and 2 mM citrate with same addition daily. , inorganic phosphate concentration in the challenge solution (mM), , calcium removed by Serratia cells (mM)

A scanning electron microscope study (SEM) was performed for samples harvested after eight days using cells challenged with CaCl2 (1 mM), and citrate (0 mM or 2 mM) without (Fig. 3a) or in the presence of 5 mM G2P (Fig. 3b) added daily. Calcium phosphate crystals were visible heavily coating the cell surface but with no apparent differences between the precipitate made from solutions with different concentration of calcium chloride, G2P and citrate. HA crystals were identified by XRD by comparing the pattern with that of the commercial HA and also the XRD standard data file (Fig. 4). The size of HA crystals was calculated by computing the diffraction pattern using the Langford programme [17]; the HA crystals on the cells were only 15-25 nm, substantially below a typical commercial product (Table 1). The largest size (25 nm) of HA crystal was from solution with addition of substrate (G2P) at its saturating concentration (5 mM) in the absence of citrate. The crystal size was considerably reduced when either the concentration of G2P was lowed to 1 mM or citrate (2 mM) was introduced into the solution. When the cells were challenged with 1 mM G2P and 2 mM citrate, Pi release was at the slowest rate and the concentration of free calcium ions was also very low due to citrate complexing. Hence Pi had to remove Ca from the complex to form calcium phosphate precipitate. Indeed, the smallest crystals were from solution with daily addition of 1 mM G2P and 2 mM citrate (15-16 nm). Therefore, the crystals formed at a size governed by the availability of free calcium ions and the rate of phosphate release using the bacterial cell surface as the nucleation template.

4. CONCLUSION The synthesis of nano-particulate calcium phosphate was achieved by challenging

Serratia cells in the solution of 50 mM TAPSO buffer (pH 9.2), 1 mM calcium chloride, 2 mM (or 0 mM) citrate and 5 mM (or 1mM) G2P with the same daily doses. The biosynthesised crystals were located by scanning electron microscopy, and identified by X-ray powder diffraction. The sizes of crystals were 15-25 nm, substantially below a typical commercial product. However, a more detailed study is needed for this biomineralization system to produce the best HA biomineral for commercial exploitation. The mechanical strength and potential biomedical applications of the biosynthesised HA will also need further investigation.


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a. b.

Figure 3. Calcium accumulation by Serratia cells viewed under SEM. Five biofilm-foam cubes were challenged in 25 ml of 50 mM TAPSO/NaOH buffer (pH 9.2) solution with 1 mM CaCl2 only (a.) or with 1 mM CaCl2 and 5 mM G2P with or without 2 mM citrate (b.) with the same daily addition (for eight days). Cells were harvested and treated as described in Experimental.

Figure 4. The X-ray powder diffraction pattern of the cell-bound material and hydroxyapatite. Upper: pattern for the sample from 50 mM TAPSO buffer (pH 9.2) solution with 1 mM calcium chloride, and 5 mM G2P with the same addition daily for eight days. Lower: pattern for the commercial HA crystals. Vertical lines: superimposition of the obtained pattern upon the standard database files for HA [26]


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Table.1. The size of crystals for HA precipitated from different solutions with immobilized cells and the commercialised HA. Samples were centrifuged after eight days, treated and the calcium phosphate crystals were identified by XRD as HA as described in Experimental. The crystal size was calculated by computing the diffraction pattern using the Langford programme [17]

Samples 1 2 3 4 5* 6* HA** CaCl2 1 mM 1 mM 1 mM 1 mM 5 mM 5 mM N/A

G2P 5 mM 5 mM 1 mM 1 mM 1 mM 1 mM N/A

First day addition: 50 mM TAPSO buffer, pH 9.2 Citrate 2 mM 0 mM 2 mM 0 mM 2 mM 0 mM N/A

CaCl2 1 mM 1 mM 1 mM 1 mM 1 mM 1 mM N/A

G2P 5 mM 5 mM 1 mM 1 mM 1 mM 1 mM N/A Daily addition to final concentration for seven days Citrate 2 mM 0 mM 2 mM 0 mM 2 mM 0 mM N/A HA size (nm) 21 25 15 20 16 20 167 *Note that following the initial dose of 5 mM Ca2+ 1 mM Ca2+ was dosed subsequently in samples 5 and 6. **Commercial HA: Capital 60 hydroxylapatite, Plasma Biotal Ltd., Tideswell, Derbyshire, UK.

ACKNOWLEDGEMENTS This project was supported by the BBSRC (Grant No. 6/E11940). The authors wish to

thank Dr J. I. Langford (School of Physics & Space Research, University of Birmingham) for assistance with the XRD technique and Mrs L. Tomkins and Mr P. Whittle for help with electron microscopy. The authors also thank Recticel (Belgium) for the gift of reticulated foam for the growth of biofilm.

REFERENCES 1. M. Jarcho, Clin. Orthopaed. Rel. Res., 207(1981)290. 2. S. D. Cook, J. F. Kay, K. A. Thomas and M. Jarcho, Int. J. Oral Maxillofacial Impl.,

2(1987)1. 3. A. N. Cranin, M. Mehrali and M. Baraoidan, J. Implantol., 20(1994)235. 4. R. W. Siegel, Mechanical properties of nanophase materials. In Synthesis and

Properties of Mechanically Alloyed and Nanocrystalline Materials, Parts 1 & 2, Materials Science Forum, 235 (1997) 851.

5. T. Klaus, R. Joerger, and C. Granqvist. Proc. Nat. Acad. Sci., 96 (1999) 13611. 6. P. Yong, N. A. Rowson, J. P. G. Farr, I. R. Harris and L.E. Macaskie. J. Chem. Tech.

Biotechnol. 77 (2002) 593. 7. P. Yong, N. A. Rowson, J. P. G. Farr, I. R. Harris and L. E. Macaskie. Biotechnol.

Bioeng., 80(2002)369. 8. P. Yong, N. A. Rowson, J. P. G. Farr, I. R. Harris and L. E. Macaskie. Environ.

Technol., In press (2003). 9. S. Mann, J. Chem. Soc Dalton Trans., 21 (1997) 3953. 10. L. E. Macaskie, R. M. Empson, A. K. Cheetham, C. P. Grey, and A. J. Skarnulis,

Science, 257 (1992) 782. 11. P. Yong and L.E. Macaskie J. Chem. Tech. Biotechnol., 63(1995)101. 12. P. Pattanapipitpaisal, A. N. Mabbett, J. A. Finlay, A. J. Beswick, M. Paterson-Beedle,

A. Essa, J. Wright, M. R. Tolley, U. Badar, N.A. Ahmed, J. L. Hobman, N. L. Brown and L. E. Macaskie, Environ. Technol., 23 (2002) 731.


1213

13. L. E. Macaskie, K. M. Bonthrone, P. Yong, and D. T. Goddard, Microbiology, 146 (2000) 1855.

14. L.E. Macaskie and A.C.R. Dean. Biotechnol. Lett., 7 (1985) 627. 15. J. A. Finlay, V. J. M. Allan, A. Conner, M. E. Callow, G. Basnakova and L. E.

Macaskie, Biotechnol. Bioeng., 63(1999)8. 16. B. C. Jeong and L. E. Macaskie, Enz. Mic. Technol., 24 (1999) 218. 17. J. I. Langford. National Institute of Standard and Technology Special Publication 846.

Proceedings of the International Conference. Accuracy in Powder Diffraction II, NIST, Gaitherburg, MD, May 26-29, (1992) 110.

18. C. Durucan and P. Brown. J. Mat. Sci.: Materials in Medicine, 11 (2000) 365. 19. G. H. Jeffery, J. Bassett, J. Mendham, R. C. Denney (eds). Vogel’s Textbook of

Quantitative Chemical Analysis, Fifth Edition. Longman Scientific Technical, John Wiley And Sons Inc., New York, USA (1989).

20. P. Boudeville, S. Serraj, J.M. Leloup, J. Margerit, B. Pauvert, and A. Terol. J. Mat. Sci.: Materials in Medicine, 10 (1999) 99.

21. S. Matsuya, S. Takagi, and L.C. Chow. J. Mat. Sci.: Materials in Medicine, 11 (2000) 305.

22. E. Fernandez, F.J. Gil, M.P. Ginebra, F.C.M. Driessens, J.A. Planell and S.M. Best. J. Mat. Sci.: Materials in Medicine, 10 (1999) 223.

23. P. Yong and L.E. Macaskie. Biotechnol. Bioeng., 55(1997)821. 24. S. Mann. Nature, 332 (1988) 119. 25. J.A. Dean (ed). Lange's Handbook of Chemistry, McGraw-Hill, Inc. New York, 1992. 26. Anon. Powder Diffrection File. Card No. 09-0432, JCPDS, Swarthmore, PA, USA

(1992).



1215

The effect of maintenance on the ferrous-iron oxidation kinetics of Leptospirillum ferrooxidans

C.J.N. Dempers1, A.W. Breed2 and G.S. Hansford

Gold Fields Minerals Bioprocessing Laboratory, Department of Chemical Engineering, University of Cape Town, Rondebosch, 7701, South Africa

Abstract The results of batch experiments carried out using a predominantly Leptospirillum

ferrooxidans culture indicated that the maintenance requirement of energy sufficient cultures is a combination of the variable and constant maintenance energy requirements. For this reason, modelling the bioenergetics of energy sufficient cultures assuming a constant maintenance energy requirement results in a significant overestimation of the maximum bacterial yield. However, the variable maintenance equation proposed by Pirt (Pirt, 1982) may be used to determine the maximum bacterial yields and the maintenance coefficients from both substrate lean and substrate rich experiments. This is because this model is able to account for variations in the growth rate as a result of changes in the maintenance requirement due to the microorganisms being limited by more than one factor.

Keywords: batch processing, bioreactors, Chemoautotrophs, kinetic parameters, maintenance requirements, modelling

1. INTRODUCTION Recent work has provided strong evidence that the bioleaching of sulfide minerals

occurs via a multiple sub-process mechanism [1-5]. According to Sand et al. [2], the disulfides pyrite (FeS2), molybdenite (MoS2), and tungstenite (WS2) are chemically oxidised by the ferric-iron present in the bioleaching medium via the main intermediate thiosulfate:

++−+ ++→++ H6Fe7OSOH3Fe6FeS 22322

32 (1)

The thiosulfate is subsequently, degraded in a cyclic process to sulfate, with elemental sulfur being a side product [2]:

++−+− ++→++ H10Fe8SO2OH5Fe8OS 2242

3232 (2)

The above thus explains why ferrous-iron oxidizing bacteria alone, e.g. Leptospirillum ferrooxidans, are able to oxidise these metal sulfides.3 1 MRT Africa, Park Place East, 104 North Rand Road, Hughes, Boksburg, 1460, South Africa 2 Department of Chemical Engineering, University of Sydney, NSW 2006, Australia 3 The genus Leptospirillum has recently been shown to have two species, viz. Leptospirillum ferrooxidans and Leptospirillum ferriphilum [6]. This differentiation is not however considered in this work.


1216

In contrast to the above, metal sulfides such as galena (PbS), sphalerite (ZnS), chalcopyrite (CuFeS2), hauerite (MnS2), orpiment (As2S3), and realgar (As4S4) are oxidised by both the ferric-iron and protons present in the bioleaching solution; the main intermediates of these reactions are polysulfides and elemental sulfur [2]:

++++ ++→++ 2n2

23 FeSH5.0MHFeMS (n > 2) (3) +++ ++→+ HFeS125.0FeSH5.0 2

83

n2 (4) +− +→+++ H2SOOHO5.1S125.0 2

4228 (5)

These metal sulfides are thus degradable by bacteria that are able to oxidize sulfur compounds, e.g. Acidithiobacillus thiooxidans.4

If ferrous-iron oxidising bacteria are present within the system, then the ferrous-iron produced by Reactions 1 to 4 is subsequently oxidised to the ferric form by these bacteria:

OH2Fe4H4OFe4 23

22 +→++ +++ (6)

A multiple sub-process mechanism, such as the one described above, thus suggests that the overall process can be expressed as a number of interconnected chemical and bacterial sub-processes, the kinetics of which may be studied separately, and the results obtained used to predict the performance of bioleach reactors for a variety of different minerals, micro-organisms and operating conditions.

To date a number of kinetic models for bacterial ferrous-iron oxidation have been proposed [8]. These models can be broadly classified as either empirical or Michaelis-Menten/Monod based. In the proposed models, the bacterial specific growth rate, µ, is usually related to the bacterial specific ferrous-iron utilisation rate, +2Fe

q , via the

maximum bacterial yield on ferrous-iron, maxXFe2Y + , and a constant maintenance coefficient

on ferrous-iron, +2Fem ; viz. using the Pirt Equation [9]:

+

+

+ +µ

= 2

2

2 FemaxXFe

Fe mY

q (7)

The bacterial specific ferrous-iron utilisation rate, +2Feq , is defined as follows:

X

FeFe c

rq

2

2

+

+

−= (8)

where:

+2Fer is the ferrous-iron production rate, and

Xc is the bacterial concentration expressed as mol C.l-1.

From Equation 7 it is thus apparent that the Pirt Equation [9] is based on the assumption that the biomass specific maintenance requirement is constant and independent of the growth rate; the maintenance energy requirement is the minimum amount of substrate, per unit of biomass, which is required to maintain the vital functions of the microorganisms.

Although the Pirt Equation [9] has been widely used to model the ferrous-iron oxidation kinetics of the bacteria encountered in bioleaching, a dependence of the experimentally determined yield and maintenance coefficients on the growth conditions 4 Acidithiobacillus thiooxidans was previously named Thiobacillus thiooxidans [7].


1217

has been reported [5]. Boon [5] reported that it was necessary to fit the Pirt Equation [9] to the initial and late batch data separately, i.e. the calculated values of max

XFe2Y + and +2Fem

depended on whether the growth was energy sufficient or energy limited. Similar dependencies on the growth conditions have been observed during experiments performed using other microorganisms [10-13].

Neijssel and Tempest [14] suggested that these deviations could be attributed to the growth of the microorganisms being limited by factors other than the energy source. This led these workers to propose that the maintenance requirement was not constant, but varied with changes in the bacterial growth rate. Pirt [15] modified the relationship developed by Neijssel and Tempest [14] by assuming that the energy required for maintenance included a term that decreased with an increase in the specific growth rate. Although the variable maintenance approach has been successfully used to describe the kinetics of non-bioleaching microorganisms, it has yet to be used in the modelling of the ferrous-iron oxidation kinetics of the bacteria used in bioleaching operations.

The primary objective of the work presented below was to determine whether or not the variable maintenance equation proposed by Pirt [15] could be used to describe the batch growth data of a predominantly L. ferrooxidans culture during both energy sufficient and energy-limited growth. A further objective of the work was to determine whether or not the kinetic parameters determined in this way were similar to those determined previously during continuous ferrous-iron oxidation experiments performed using the same bacterial culture and similar growth conditions.5

2. THEORETICAL ASPECTS If bacteria are grown in continuous culture and under conditions in which the carbon

source is limited to carbon dioxide, then the carbon dioxide production rate, 2COr , can be

used to estimate the biomass production rate growth rate, Xr , [5]:

2COX rr −= (9)

However, if the bacteria are grown in batch culture and under conditions in which the carbon source is limited to carbon dioxide, then the bacterial concentration at time, t, can be estimated from the carbon dioxide production rate,

2COr , and the initial bacterial concentration,

0Xc :

∫=

=

−+=tt

0tCOXX dtrcc

20 (10)

Furthermore if the stoichiometric formula of bacteria is assumed to be CH1.8O0.5N0.2 [18, 19] and energy for bacterial growth and maintenance is obtained from the oxidation of ferrous-iron, performing mass and charge balances and solving in terms of the carbon dioxide and oxygen production rates,

2COr and 2Or , respectively, yields the degree-of-

reduction balance:

222 COOFer2.4r4r −−=− + (11)

5 The continuous experiments are reported elsewhere [16, 17] whereas the batch experiments were performed during the course of the current study.


1218

If the maintenance energy requirement is assumed to be constant and independent of the growth rate and the theoretical maximum growth yield, max

XFe2Y + , occurs when the maintenance energy, +2Fe

m , is zero, then the relationship between the amount of substrate consumed by the biomass for bacterial growth and maintenance can be described using Equation 7, i.e. the Pirt Equation [9]. However, if the maintenance requirement is assumed to vary with changes in the bacterial growth rate and if the maintenance energy term is assumed to include a portion that decreases with an increase in the specific rate of growth, then the relationship between the bacterial specific substrate utilisation rate, the maximum substrate specific yield and the and the growth rate must be described using [15]:

)µk1(mmY

µq vFeFemaxFe 22

X2Fe

2 −++= ++

+

+ (12)

In Equation 12, +2Fem is the constant maintenance coefficient on ferrous iron, whereas

vFe2m + and k are growth dependent maintenance coefficients. Pirt [15] also postulated that

the growth rate dependent maintenance energy requirement decreases to zero as the specific growth rate, µ, approaches the maximum bacterial specific growth rate, maxµ . This in turn implies that:

maxµ1k = (13)

Combining Equations 12 and 13 and rearranging yields:

vFeFemax

vFe

maxXFe

Fe 22

2

2

2 mmµµm

Y1q ++

+

+

+ ++⎟⎟⎠

⎞⎜⎜⎝

⎛−= (14)

A graphical representation of Equation 14 is shown in Figure 1; the two lines, (a) and (b), represent experimental data obtained under different energetic conditions.

From Figure 1 and the definition of the constant maintenance requirement, it is apparent that, under conditions in which the growth rate is limited by the concentration of the energy source, i.e. during energy limited growth, the maintenance energy requirement of the microorganisms is constant and lower than the maintenance energy requirement when the growth is limited by another factor, e.g. the concentration of a specific trace metal. Under these conditions the constant maintenance coefficient can thus be obtained from the intercept of the solid line and the y-axis, i.e. from line (a). Furthermore, because the variable maintenance requirement decreases to zero under energy-limited conditions, the maximum growth yield can also be determined from the slope of line (a).

In contrast to the above, the maintenance energy requirement of energy sufficient cultures is a combination of the variable and constant maintenance energy requirements (see line (b) in Figure 1). The variable maintenance energy requirement can therefore be determined from the difference between the values of the y-intercepts determined during energy sufficient and energy limited growth experiments; the value of the variable maintenance coefficient will thus depend on the limiting factor. These factors include the concentrations of phosphate, sulfate and ammonium ions, the concentration of heavy metals and the bacterial growth rate.


1219

µ (h-1)

0

q Fe2+ (m

ol F

e2+.(m

ol C

)-1.h

-1)

0

mFe2+

mFe2+ + mFe2+v

µmax

ab

qFe2+max

1/YFe2+max

Figure 1. Relationship between the bacterial specific growth rate, µ, and the bacterial specific ferrous-iron utilisation rate, +2Fe

q , assuming a) constant maintenance and b) variable maintenance

In addition to the above, it can be seen that the maximum bacterial specific ferrous-iron utilisation rate, max

Fe2q + , and the maximum bacterial specific growth rate are defined by the intersection of the solid and the broken lines.

Equation 14 can also be written in terms of the bacterial specific oxygen utilisation rate,

2Oq , hence the maximum bacterial yield on oxygen, maxXO2

Y , the constant maintenance coefficient on oxygen,

2Om and the growth dependent maintenance coefficient on oxygen, vO2

m , may be calculated in the same way as the ferrous-iron based parameters. The validity of the yield and maintenance coefficients calculated determined in this manner can in turn be can be checked using the degree of reduction balance, Equation 10; i.e. using:

maxXFe

maxXFemax

XO2

2

2 Y2.41Y4

Y+

+

−= (15)

4m

m2

2

FeO

+

= (16)

It is thus also apparent from Figure 1 and the preceding discussion that the values of the kinetic parameters determined from experimental data can be highly dependent on the conditions used.

3. MATERIALS AND METHODS The batch ferrous-iron oxidation experiments performed during the course of this

study were carried out at pH 1.70 and 30, 35 and 40°C and at 40°C and pH 1.10, 1.30, 1.50 and 1.70, using a predominantly L. ferrooxidans culture. The bacterial culture used was initially obtained from a vat-type two-stage (2×20 l) continuous bioleaching mini-plant treating an arsenopyrite/pyrite concentrate from Fairview Gold Mine in Barberton, South Africa [20]. The inoculum for each of the batch experiments was obtained from steady state continuous cultures of microorganisms from the bioleaching mini-plant grown


1220

on ferrous-iron medium at a dilution rate of 0.04 h-1 and the temperature and pH to be used in the batch experiment.

The batch experiments themselves were carried out in 2l air-sparged, agitated bioreactors. The bioreactors had a H/D of 1.32 and a working volume of 1l. Circulating water from constant temperature baths through the bioreactor jackets controlled the temperature in the bioreactors. The pH of the solution in the bioreactors was monitored continuously and maintained at the required pH by the addition of concentrated sulphuric acid (98%). The redox potential of the bioleaching solution in the bioreactors was measured continuously using Metrohm redox electrodes (Pt-Ag/AgCl) and logged by computer.

The ferrous iron media used consisted of 12 g.l-1 FeSO4.7H2O, 1.11 g.l-1 K2SO4, 0.53 g.l-1 (NH4)2HPO4, 1.83 g.l-1 (NH4)2SO4 and 10 ml.l-1 trace element solution [21] adjusted to between pH 0.95 and pH 1.30 using H2SO4

conc. Air, at a flow rate of 1 l.min-1, was supplied to the bioreactors using Brooks mass

flow controllers. The off-gas from the bioreactors was dried prior to passing through CO2 and O2 gas analysers. This enabled the oxygen utilisation rate, -rO2, carbon dioxide utilisation rate, -rCO2, and biomass concentration, cX, to be determined [4]. The experimental equipment used is described in greater detail elsewhere [3].

The total iron concentration in solution was determined by both atomic adsorption spectroscopy (AAS) and by titration with potassium dichromate [22]. This enabled the ferric/ferrous-iron ratio and the ferrous- and ferric-iron concentrations to be determined using a calibration curve for the specific electrode and the Nernst equation [5]. Changes in the ferrous- and ferric-iron and total iron concentrations with time allowed the ferrous-iron production rate, +2Fe

r , to be calculated.

4. RESULTS AND DISCUSSION In the first instance an attempt was made to determine the kinetic parameters, max

XFe2Y + , and +2Fe

m using the Pirt Equation [9] i.e. by plotting µ vs. +2Feq . A typical plot of µ vs.

+2Feq for the experiment performed at 40°C and pH 1.10 is shown in Figure 2.

From Figure 2 it is evident that there are three distinct linear regions, hence it was not possible to determine the values of max

XFe2Y + and +2Fem in this way. However, the similarity

between Figs. 1 and 2 suggests that the micro-organisms used were in fact limited by different factors during the period over which the experiment was performed and should therefore be modelled using the variable maintenance equation, viz. Equation 14.

The values of the maximum bacterial yield on ferrous-iron, maxXFe2Y + , and the constant

maintenance coefficient on ferrous-iron, +2Fem , were therefore calculated by linear

regression of the data obtained during the latter stages of the batch experiments (region cd); it was assumed that there was no variable maintenance requirement in this region. This allowed the variable maintenance coefficient on ferrous-iron, v

Fe2m + , to be estimated from a regression line drawn through the data obtained during the initial stages of the batch experiment (region ab). Finally, the maximum bacterial specific growth rate, maxµ , and the maximum bacterial specific ferrous-iron utilisation rate, max

Fe2q + , were determined


1221

from the intercept of the regression lines drawn through the data obtained during the initial and latter stages of the batch experiments, as demonstrated in Figure 1.

The average ferrous-iron and oxygen-based parameters calculated in this manner are listed in Table 1 and Table 2. The validity of the yield and maintenance coefficients listed in Table 1 and Table 2 was checked by comparing the experimental (calculated) values with the values predicted by the "degree of reduction" balance i.e. Equations 15 and 16. These comparisons are displayed in Figure 3, from which it is apparent that the correlation is good (average correlation coefficient R2 = 0.9596).

µ (h-1)

0.00 0.02 0.04 0.06 0.08

q Fe2+ (m

ol F

e2+.(m

ol C

)-1.h

-1)

0

4

8

12

16

a

d

b

c

Figure 2. Variation in the experimentally observed bacterial specific ferrous-iron utilisation rate, +2Fe

q , with changes in the bacterial specific growth rate, µ

Table 1. Average ferrous-iron based bioenergetic parameters

Experimental conditions pH 1.70 40°C 30°C 35°C 40°C pH 1.50 pH 1.30 pH 1.10

maxFe2q + (mol Fe2+.(mol C)-1.h-1) 14.65 22.31 10.65 14.32 18.30 16.36 max

XFe2Y + (mol Fe2+.(mol C)-1.h-1) 0.0056 0.0055 0.0036 0.0058 0.0054 0.0058

+2Fem (mol C.(mol Fe2+)-1.h-1) 0.540 0.855 0.002 0.95 2.26 1.00

vFe2m + (mol Fe2+.(mol C)-1.h-1) 7.85 6.50 11.32 6.06 10.08 5.95

maxµ (h-1) 0.079 0.119 0.038 0.077 0.087 0.089


1222

Table 2. Average oxygen based bioenergetic parameters

Experimental conditions pH 1.70 40°C 30°C 35°C 40°C pH 1.50 pH 1.30 pH 1.10

maxO2

q (mol Fe2+.(mol C)-1.h-1) 3.37 5.63 2.61 3.50 4.49 4.00 max

XO2Y (mol Fe2+.(mol C)-1.h-1) 0.0228 0.0227 0.0158 0.0236 0.0221 0.0239

2Om (mol C.(mol Fe2+)-1.h-1) 0.131 0.203 0.0917 0.2312 0.5638 0.2423 vO2

m (mol Fe2+.(mol C)-1.h-1) 2.32 1.59 2.74 1.52 2.51 1.49 maxµ (h-1) 0.074 0.123 0.040 0.077 0.087 0.090

YO2 (mol C.(mol O2)

-1)

0.000 0.008 0.016 0.024 0.032

4YFe

2+/(1

-4.2

YFe

2+)

0.000

0.008

0.016

0.024

0.032Figure 3(a)

mO2 (mol O2.(mol C)-1.h-1)

0.00 0.15 0.30 0.45 0.60

(mFe

2+)/4

(mol

Fe2+

.(mol

C)-1

.h-1

)

0.00

0.15

0.30

0.45

0.60Figure 3(b)

mO2

v (mol O2.(mol C)-1.h-1)

0.0 0.8 1.6 2.4 3.2

(mFe

2+v )/4

(mol

Fe2+

.(mol

C)-1

.h-1

)

0.0

0.8

1.6

2.4

3.2Figure 3(c)

Figure 3. Comparison of the predicted and experimental relationships for (a) the maximum bacterial yield, (b) the constant maintenance coefficient and (c) the variable maintenance coefficient


1223

In addition to the above, it is also apparent from the results listed in Table 1 and Table 2 that the maximum bacterial yields on ferrous-iron and oxygen and their respective constant and variable maintenance coefficients, did not vary significantly with either temperature or pH. For this reason, average maximum bacterial yield and constant maintenance coefficients on ferrous-iron and oxygen were calculated by linear regression using the data from the latter stages of all the batch experiments performed during this investigation. These values are listed in Table 3, together with previously reported values of these parameters. From the results presented in Table 3 it is apparent that the average maximum bacterial yields and constant maintenance coefficients determined during the course of this investigation are similar to previously published values.

Table 3. Average values of the maximum bacterial yield and constant maintenance coefficient on ferrous-iron and the maximum bacterial yield and maintenance coefficient on oxygen together with the values reported previously

Breed [3] Van Scherpenzeel et al. [23] Current work

maxXFe2Y + (mol C.(mol Fe2+)-1) 0.0075 0.011 0.0060

+2Fem (mol Fe2+.(mol C)-1.h-1) 1.196 0.444 1.261

maxXO2

Y (mol C.(mol O2)-1) 0.0279 0.046 0.0247

2Om (mol O2.(mol C)-1.h-1) 0.2376 0.0425 0.3096

The maximum bacterial specific ferrous-iron utilisation rates, maxFe2q + , and the maximum

bacterial specific growth rates, maxµ , calculated using the values listed in Table 3 and the variable maintenance equation, Equation 14, are listed in Table 4, together with previously published values of these parameters.

Table 4. Comparison of the maximum bacterial specific growth rates and ferrous-iron utilisation rates calculated using the variable maintenance equation and those published by Breed [3]

Breed [3] Current work Experimental

conditions maxµ

(h-1)

maxFe2q +

(mol Fe2+.(mol C)-1.h-1)

maxµ (h-1)

maxFe2q +

(mol O2.(mol C)-1.h-1) 30°C; pH 1.70 0.0397 8.65 0.074 14.65 35°C; pH 1.70 0.0638 11.01 0.123 22.31 40°C; pH 1.70 0.0862 13.62 0.040 10.65 40°C; pH 1.50 0.1238 19.02 0.077 14.32 40°C; pH 1.30 0.1077 15.57 0.087 18.30 40°C; pH 1.10 0.1027 15.26 0.090 16.36

From these results it is apparent that the values of maxFe2q + and maxµ calculated using the

variable maintenance model are also similar to previously published values of these parameters. The results presented thus suggest that during the initial stages of a batch experiment (region ab in Figure 2) the microorganisms are not substrate-(energy-)limited, whereas in the latter stages of the experiment (region cd in Figure 2) they are substrate (energy) limited. The points, (b) and (c) in Figure 2 represent points at which a change in the limiting substrate occurs. In fact, within the transition region (region bc in Figure 2) the limiting substrate is continually changing, hence a line with a negative slope is


1224

obtained. In other words, during batch experiments the growth of the microorganisms are limited by more than one factor separately, hence the data cannot be modelled using the constant maintenance equation. However, it can be modelled using the variable maintenance equation.

5. CONCLUSIONS The results of the batch experiments performed during the course of this study

indicate that the maintenance requirement of microorganisms is actually a combination of the variable and constant maintenance energy requirements. The variable maintenance requirement depends on the limiting substrate, whereas the constant maintenance requirement is dependent on the energy source. Where the bacterial growth rate is limited solely by the energy source, the variable maintenance requirement decreases to zero. Modelling the bioenergetics of energy sufficient cultures assuming a constant maintenance energy requirement thus results in a significant overestimation of the maximum bacterial yield. However, the variable maintenance equation proposed by Pirt [15] may be used to determine the maximum bacterial yields and the maintenance coefficients from both substrate lean and substrate rich experiments. This is because this model is able to account for variations in the growth rate as a result of changes in the maintenance requirement due to the micro-organisms being limited by more than one factor, e.g. ferrous-iron, ammonia, phosphate, sulfate, etc.

It is thus suggested that the variable maintenance equation be used instead of the constant maintenance equation when quantifying the bioenergetics of microorganisms, including those encountered during the bioleaching, unless the energy source is known to be the limiting substrate. The above is especially important for the case of bioleaching using ferrous-iron oxidising microorganisms at low redox potentials (high ferrous-iron concentrations), e.g. during mesophilic chalcopyrite bioleaching and during thermophilic bioleaching.

REFERENCES 1. G.S. Hansford and T. Vargas, Hydrometallurgy, 59 (2001) 135. 2. W. Sand, T. Gehrke, P.-G. Jozsa and A. Schippers, Hydrometallurgy, 59 (2001) 159. 3. A.W. Breed, Studies on the mechanism and kinetics of bioleaching with special

reference to the bioleaching of refractory gold-bearing arsenopyrite/pyrite concentrates, PhD Thesis, University of Cape Town, Cape Town, South Africa, 2000.

4. M. Boon, G.S. Hansford and J.J. Heijnen, Biohydrometallurgical Processing I, T. Vargas, C.A. Jerez, J.V. Wiertz and H. Toledo (eds.), University of Chile, Santiago, Chile, 1995.

5. M. Boon, Theoretical and experimental methods in the modelling of biooxidation Kinetics of sulphide minerals, PhD Thesis, Technische Universiteit Delft, The Netherlands, 1996.

6. N.J. Coram and D.E. Rawlings, Appl. Environ. Microbiology, 68 (2002) 838. 7. D.P. Kelly and A.P. Wood, Int. Jour. Systematic and Evolutionary Microbiol., 50

(2000) 511. 8. M. Nemati, S.T.L Harrison, C. Webb and G.S. Hansford, Biochem. Eng. J., 1 (1998)

171. 9. S.J. Pirt, Proceedings of the Royal Society of London. Series B: Biological Sciences,

163 (1965) 224.


1225

10. J.G. Kuenen, Growth yields and “maintenance energy requirement” in Thiobacillus species under energy limitation, Arch. Microbiol., 122 (1979) 183.

11. A.J. Downs and C.W. Jones, Arch. Microbiol., 105 (1975) 159. 12. O.M. Neijssel and D.W. Tempest, Arch. Microbiol., 106 (1975) 251. 13. A.H. Stouthamer and C.W. Bettenhausen, Arch. Microbiol., 102 (1975) 187. 14. O.M. Neijssel and D.W. Tempest, Arch. Microbiol., 107 (1976) 215. 15. S.J. Pirt, Arch. Microbiol., 133 (1982) 300. 16. A.W. Breed, C.J.N. Dempers, G.E. Searby, M.N. Gardner, D.E. Rawlings and G.S.

Hansford, Biotechnol. Bioeng., 65 (1999) 44. 17. A.W. Breed and G.S. Hansford, Biochem. Eng. J., 3 (1999) 193. 18. C.A. Jones and D.P. Kelly, J. Chem. Technol. Biotechnol., 33B (1983) 241. 19. J.A. Roels, Energetics and kinetics in biotechnology, Elsevier Biomedical Press,

Amsterdam, 1983. 20. A.W. Breed, S.T.L. Harrison and G.S. Hansford, IBS-BIOMINE 97, Australian

Mineral Foundation, Glenside, Australia, 1997. 21. W. Vishniac and M. Santer, Bacteriol. Revs., 21 (1957) 195. 22. G.H. Jeffrey, J. Bassett, J. Mendham and R.C. Denney (eds.), Vogel’s Textbook of

Quantitative Chemical Analysis, 5th edition, Longman Scientific & Technical, New York, 1989.

23. D.A. van Scherpenzeel, M. Boon, C. Ras, G.S. Hansford and J.J. Heijnen, Biotechnol. Prog., 14 (1998) 425.



1227

The kinetics of thermophilic ferrous-iron oxidation

G.E. Searby and G.S. Hansford

Gold Fields Mineral Bioprocessing Laboratory, Department of Chemical Engineering, University of Cape Town, Private Bag, Rondebosch, 7701, South Africa

Fax: +27-21-689 7579. E-mail: [email protected]

Abstract The kinetics of ferrous iron oxidation by a thermophilic archaeal culture was

investigated in continuous culture, at temperatures ranging from 65 to 75ºC, and a pH of 1.5.

The reaction kinetics were followed by on-line monitoring of the solution redox potential and of the concentrations of oxygen and carbon dioxide in the off gas stream. The specific rate of ferrous iron utilisation was calculated from the rates of oxygen and carbon dioxide utilisation via a degree of reduction balance, and examined as a function of the ferric/ferrous iron ratio, determined from the redox potential and checked by measuring dissolved iron concentrations.

The results were modelled using a kinetic model developed for mesophilic oxidation based on Michaelis-Menten enzyme kinetics and proportional to the ferric/ferrous iron ratio. The effect of temperature was accounted for by the incorporation of an Arrhenius term.

The thermophiles achieved specific iron utilisation rates similar to those reported for L. ferrooxidans, but was active at significantly lower redox potentials, and less active at high redox, more comparable to At. ferrooxidans.

Keywords: thermophiles, bioleaching, kinetics, iron oxidation, continuous culture

1. INTRODUCTION The recalcitrance of chalcopyrite to bioleaching by mesophilic microorganisms has

led to interest in high temperature leaching. Thermophilic archaea have been shown to be capable of effective leaching. (Le Roux, 1988, and Dew et al, 1999).

Like mesophilic systems, thermophilic mixed cultures often contain species that are predominantly iron oxidisers and others that are predominantly sulfur oxidisers. This indicates the possibility that the leaching mechanism is similar to that generally accepted for conventional bioleaching microbes, i.e. a chemical ferric leach followed by microbially mediated ferrous iron oxidation reforming the primary ferric leach reagent, and microbial oxidation of sulphur compounds released.

An understanding of the ferrous iron oxidation kinetics of these archaea at their operating temperature is therefore desirable so that their bioleaching capabilities may be fully exploited.


1228

Mesophilic ferrous iron oxidation (mostly focussed on Acidithiobacillus ferrooxidans) has been studied extensively. A number of kinetic models have been proposed, starting with empirical systems such as the logistic equation and simple Monod models, followed by more developed models. These were often based on Michaelis-Menten enzyme kinetics and adapted to show substrate and/or product inhibition and effects of various parameters such as dissolved oxygen concentration, pH, and temperature.

Ingledew (1982) proposed a chemiosmotic theory for ferrous iron oxidation in At. ferrooxidans, involving the generation of a transmembrane potential and a proton-motive force by splitting the two half reactions in the overall chemical reaction of ferrous iron oxidation across the cell membrane. This model highlighted the influence of the solution redox potential on the microbial growth and oxidation kinetics and has led to the generation of models based on redox potential (Huberts, 1994; Boon, 1995; and Meruane et al, 2002).

It is proposed that the oxidation of ferrous iron by thermophiles may be controlled by a similar mechanism, and hence that the kinetics can be described using a similar model.

2. MATERIALS AND METHODS Experiments were performed in three 1.7L, air-sparged glass vessels, each with an

H/D ratio of 1.32, using a working volume of 1L. Temperature in the reactors was maintained at 65, 70 and 75ºC by a heated water jacket. pH was maintained at 1.5 by manipulation of the feed pH and by addition of concentrated sulphuric acid.

The reactors were kept airtight, allowing the measurement of the change in oxygen and carbon concentration in the entering and exiting gas streams. Off-gas analysis and the measurement of the change in the solution potential have been shown to be an effective system for the investigation of ferrous-iron oxidation (Boon, 1996).

Compressed air was passed through a gas-chiller and a series of filters to obtain bone-dry medical quality air and fed into the reactor at a constant rate via a Brooks 5850S mass flow control valve, and sparged below the impeller. Agitation and gas mixing was achieved by a Lightnin 315 impeller running at 400 rpm. Air leaving the reactor passed through a reflux condenser, which served the dual purpose of drying the gas before it enters the gas analysis system, and avoiding evaporative water loss in the reactor. The offgas then passed through a cloth filter and a Hartmann and Braun CGEK sample gas conditioner before entering the analysers. Oxygen concentration was measured by a Hartmann and Braun Magnos 6G paramagnetic oxygen analyser and carbon dioxide using a Hartmann and Braun Uras 4 infrared photometer.

The culture used was a mixed culture of extreme Sulfolobus-like thermophilic archaea grown at pH 1.5 and at a temperature of 70ºC. The culture was taken initially from a semi-continuous system grown on a chalcopyrite concentrate and has been maintained in continuous culture on ferrous iron for more than 2 years.

The culture was grown in a basal salt medium containing 0.5 g/L MgSO4.7H2O, 0.4

g/L (NH4)2SO4 0.02 g/L K2HPO4.3H2O and 0.1 g/L KCl (Clark and Norris, 1996). The

energy source was 12g/L FeSO4. Reduced sulphur was added as 0.225g/L K2S4O6 as the thermophiles are unable to assimilate sulphate.

Experiments were performed in continuous culture at dilution rates ranging from 0.015 to 0.9 hr-1. Continuous flow was obtained by feeding the fresh nutrient medium via a variable speed peristaltic pump, and by removing the resultant reactor liquor at a fixed liquid level by a fixed speed peristaltic pump, maintaining a constant volume. The reactors


1229

were maintained at each dilution rate for 5 residence times before steady state was assumed. Steady state was verified by the generation of steady data for a further residence time. Wall growth was eliminated by daily scrubbing of all reactor surfaces.

The reaction kinetics was followed by off-gas analysis. In an autotrophic system, the only carbon source is atmospheric carbon dioxide, and thus the carbon dioxide utilisation rate can be used to follow the growth kinetics.

2COX rr −= (1)

This provides a non-invasive on-line measurement of the cell concentration, in terms of moles of carbon fixed.

Dr

c 2COx

−= (2)

Autotrophic microbes oxidise iron as their energy source,

OH2Fe4OH4Fe4 23microbes

22 +⎯⎯⎯ →⎯++ +++ (3)

and utilise the energy produced for cell synthesis (typical cell stoichiometry CH1.8O0.5N0.2)

CO2, H2O, 0.20.51.84 NOCHNH ⎯→⎯+ (4)

Simultaneous solution of the species and charge balances for the species in Reactions 3 and 4, or the degree of reduction balance over the same set of species (Roels, 1983)

0ri

ii =γ∑ (5)

yields the rate of microbial ferrous iron oxidation in terms of the measured off-gas parameters.

222 OCOFe r4r2.4r −−=− + (6)

The solution redox potential was measured using a Pt Ag/AgCl2 redox electrode. This was related to the ferric/ferrous-iron ratio in the reactor via the Nernst equation, approximating activities as concentrations.

⎟⎟⎠

⎞⎜⎜⎝

⎛+= +

+

]Fe[]Fe[ln

zFRTEE 2

3'0 (7)

The ferrous-iron and total iron concentrations were determined by titration with K2Cr2O7 (Jeffrey et al, 1989). Ferric iron concentrations can then be calculated by the difference between the ferrous and the total iron concentrations. This provided a check for the ferric/ferrous-iron ratio as determined by redox measurements.

The specific iron utilisation rate was modelled using a Michaelis-Menten-form model dependent on the ferric/ferrous ratio, previously used to model L. ferrooxidans (van Scherpenzeel, 1998, Breed et al, 1999) and At. ferrooxidans (Boon, 1996)

]Fe[]Fe[K1

qq

2

3

Fe

maxFe

Fe

2

2

2

+

+

+

+

+

+= (8)

Breed et al (1999) described the effect of temperature on Leptospirillum ferrooxidans as an Arrhenius function of the qmax constant


1230

RTE

0maxFe

a

2 ekq−

=+ (9)

and a linear function of the KFe2+ 0453.0T0002.0K 2Fe

−=+ (10)

Nemati and Webb described the effect of temperature on At. ferrooxidans simply as an Arrhenius function of qFe2+

max.

3. RESULTS AND DISCUSSION Continuous reactors were run at temperatures of 65, 70, and 75ºC, pH 1.5, 12gFe2+.L-1

and steady state data was obtained at dilution rates of 0.02, 0.033, 0.04, 0.05, 0.059, 0.071 and 0.08 h-1. Washout occurred at dilution rates between 0.09 and 0.11 hr-1.

All three systems are characterised by a maximum cell concentration at intermediate dilution rates decreasing at both high and low dilution rates (Figure 1). At low dilution rates the system is energy-limited increasing the cell maintenance and reducing the cell concentration, whilst at high dilution rates, the high growth rates stress the cells, which also leads to increased maintenance requirements.

The rates of ferrous iron oxidation determined from the rates of oxygen and carbon dioxide utilisation via Equation 6 were compared in Figure 2 to those determined from the difference between feed and reactor contents iron concentrations. The results confirm the applicability of the degree of reduction balance in thermophilic systems.

0

0.5

1

1.5

2

0 0.05 0.1 0.15Dilution Rate (hr-1)

Cel

l con

cent

ratio

n (m

mol

C.l-1)

0

5

10

15

0 5 10 15-4rO2 - 4.2 rCO2

-rFe

2+

Figure 1. Steady state cell concentrations for each dilution rate at 65(∆), 70 () and 75ºC ()

Figure 2. Confirmation of degree of reduction balance at 65(∆), 70 () and 75ºC ()

The dependence of the system on the redox potential via the ferric/ferrous iron ratio, as predicted by Equation 8, was demonstrated by plotting the specific iron utilisation rate obtained at each steady state against the ferric/ferrous iron ratio measured (Figure 3).


1231

0

5

10

15

20

0.1 1 10 100 1000

[Fe3+]/[Fe2+]

Spec

ific

iron

utili

satio

n ra

te

(mol

Fe.m

olC

.hr

)

Figure 3. Specific iron utilisation rates obtained at 65(∆), 70 () and 75ºC () as a function of the ferric/ferrous iron ratio and modelled by Equation 8

The specific oxidation rates increase with an increase in temperature, this would indicate that the experiments were run within the microbes' preferred operating range. Above this range, thermal deactivation is expected, and a consequent drop in rate.

The results show a similar dependence on the ferric/ferrous iron ratio as was observed in mesophilic systems. The specific utilisation rate reaches a maximum at the lowest ferric/ferrous iron ratio corresponding to the fastest dilution rate before washout. The rate decreases with increasing redox potential until the ferrous iron concentration reaches levels below the minimum required for further growth. This has been described as the threshold substrate concentration (Braddock et al, 1984).

The data was modelled using Equation 8, obtaining the kinetic constants 2+maxFe

q and

2FeK + by linear regression. The results show good agreement between data and prediction over the range of potentials measured, suggesting that thermophilic ferrous iron oxidation is proportional to the ferric/ferrous iron ratio and in turn the redox potential.

Table 2. Kinetic constants 2+maxFe

q and 2+FeK derived from the data and compared to

literature values

Temperature 2+maxFe

q 2FeK +

65 11.08 0.155 70 12.32 0.098 Present work 75 19.36 0.227

L. ferrooxidans (Breed et al, 1999) 40 13.58 0.0033

At. ferrooxidans (Boon, 1995) 30 8.8 0.05

The 2FeK + values obtained show more commonality with those of Acidithiobacillus

ferrooxidans, than with those of Leptospirillum ferrooxidans. This has been described as the affinity of the microbe for ferrous iron, with L. ferrooxidans having the capacity to oxidise much lower concentrations of iron and hence operate at much higher redox potentials. Higher 2Fe

K + values indicate that the ferric/ferrous iron ratio or redox potential


1232

at which the oxidation and growth rates begin to drop away is lower. This may be of interest in the investigation of why thermophiles are more able to bioleach chalcopyrite, where better leaching may be achieved at lower solution redox potentials.

0

5

10

15

20

25

0.1 1 10 100 1000 10000

[Fe3+]/[Fe2+]

Spec

ific

iron

utili

satio

n ra

te

(mol

Fe.m

olC

.h)

Figure 4. A comparison of the relationship between 2+Fe

q and the ferric/ferrous iron ratio for thermophiles (70ºC (—), 75ºC (—)), and mesophiles At. ferrooxidans (----), and L. ferrooxidans(----)

The specific growth rate; this was compared to the specific iron utilisation rate to determine the yield and maintenance coefficients from the Pirt equation (Pirt, 1965)

+

+

+ +µ

= 2

2

2 FemaxXFe

Fe mY

q (11)

Straight lines fitting the constant maintenance Pirt equation (Pirt, 1965) are only expected for substrate limited growth. Deviations from linearity at high redox potential and high growth stress may be attributed to increased maintenance requirements and may be better described using the variable maintenance equation (Pirt, 1982).

Table 3. Bioenergetic parameters Yield ( 2+maxFe X

Y ) and the maintenance coefficient (mFe2+)

Temperature (ºC)

2+maxFe X

Y (mol C.(mol Fe2+)-1)

2Fem +

(mol Fe2+.mol C-1.h-1) 65 0.0082 0.394 70 0.0082 1.265 75 0.0071 1.308

The maximum yield is a very weak function of temperature, showing no significant change within the temperature range measured, and not differing greatly from values measured at mesophilic temperatures, lying between two values obtained for Leptospirillum-like systems — 0.0059 and 0.011 molC.(molFe2+)-1 obtained by Breed et al (1999) and van Scherpenzeel (1998) respectively. The maintenance coefficient is a much stronger function of temperature, indicating increased stress with increased temperature.

The effect of temperature on 2+maxFe

q (Figure 5) can be modelled by the Arrhenius equation. The data can thus be modelled as


1233

]Fe[]Fe[K1

e1070.2q

2

3

Fe

RT437.54

9

Fe

2

2

+

+

−

+

+

+

×= (12)

0

5

10

15

20

0 0.05 0.1Specific Growth Rate µ (h-1)

q Fe2

+ (m

olFe

2+.m

olC-1

.h-1

)

y = -6547.6x + 21.718R2 = 0.8814

0

1

2

3

4

0.00285 0.0029 0.00295 0.0031/T (degC-1)

ln(q

max

)

Figure 5. Determination of the energetic parameters, yield ( 2+

maxFe X

Y ) and the maintenance coefficient ( 2+Fem ) at 65(∆), 70 () and 75ºC ()

Figure 6. Effect of temperature on 2+

maxFe

q , showing fit of Arrhenius equation

0.00

5.00

10.00

15.00

20.00

0.1 1 10 100 1000

[Fe3+]/[Fe2+]

Spec

ific

iron

utili

satio

n ra

te

(mol

Fe2+

.mol

C-1.h

-1)

Figure 7. Specific iron utilisation rates obtained at 65(∆), 70 () and 75ºC () were compared to values predicted using Equation 8 (—) and Equation 12 (---)

4. CONCLUSIONS The kinetics of ferrous iron oxidation using thermophilic archaea was investigated in

continuous culture. The rate of iron consumption calculated from measured iron concentrations correlated well with that calculated from off gas analysis via the degree of reduction balance. This indicates that the overall reaction stoichiometry for microbial ferrous iron oxidation remains the same as at mesophilic temperatures.


1234

The results were successfully modelled using a Michaelis-Menten form kinetic model, developed by Boon (1996) for mesophilic bioleaching, in which the specific rate of iron oxidation is proportional to the ferric/ferrous iron ratio. The effect of temperature was accounted for by the incorporation of an Arrhenius term.

The maximum specific iron utilisation rates determined for the thermophilic systems were not greatly different to those achievable in mesophilic systems, so simple rate of iron oxidation cannot explain the advantage of bioleaching at elevated temperatures. However, the lower tolerance of high redox potential can lead to a bioleach system that self-regulates its redox to a lower level where the rate of ferric leaching is enhanced.

NOMENCLATURE cX concentration of bacteria mmol C.L-1 D dilution rate h-1 E redox potential of the solution (Pt-Ag/AgCl) mV

'0E equilibrium redox potential for Ag/AgCl electrode mV

Ea activation energy kJ.mol-1 F Faraday constant coulombs.mol-1

[Fe2+] concentration of ferrous-iron mmol Fe2+.L-1 [Fe3+] concentration of ferric-iron mmol Fe3+.L-1 Ko Arrhenius constant mol Fe2+.(mol C)-1.h-1

KFe2+ ferrous-iron based kinetic constant in bacterial ferrous-iron oxidation dimensionless

mFe2+ maintenance coefficient on ferrous-iron mol Fe2+.(mol C)-1.h-1

qFe2+ bacterial specific ferrous-iron utilisation rate mol Fe2+.(mol C)-1.h-1

maxFe2q + maximum bacterial specific ferrous-iron utilisation

rate mol Fe2+.(mol C)-1.h-1

R Universal gas constant kJ.K-1.mol-1

-rFe2+ ferrous-iron utilisation rate mmol Fe2+.L-1.h-1 -rCO2 carbon dioxide utilisation rate mmol CO2.L-1.h-1 -rO2 oxygen utilisation rate mmol O2.L-1.h-1 rX biomass production rate mmol C.L-1.h-1

maxXFe2Y + maximum bacterial yield on ferrous-iron mol C.( mol Fe2+)-1

z number of electrons involved in a reaction dimensionless γi Degree of reduction of species i dimensionless µ bacterial specific growth rate h-1

REFERENCES 1. Boon M., 1996, "Theoretical and Experimental Methods in the Modelling of

Biooxidation kinetics of Sulphide Minerals", PhD Thesis, Technische Universiteit Delft, The Netherlands.

2. Boon M., Hansford G.S. and Heijnen J.J., 1995, "The Role of Bacterial Ferrous Iron Oxidation in the Bio-Oxidation of Pyrite", Vargas T., Jerez C.A., Wiertz J.V. and


1235

Toledo H. (eds.), Biohydrometallurgical Processing, 1, Santiago: University of Chile, 153-163.

3. Braddock J.F., Luong H.V. and Brown E.J., 1984, "Growth kinetics of Thiobacillus ferrooxidans isolated from arsenic mine drainage", Applied and Environmental Microbiology 48, (1), 48-55.

4. Breed A.W., Dempers C.J.N., Searby G.E., Gardner M.N., Rawlings D.E. and Hansford G.S., 1999, "The Effect of Temperature on the Continuous Ferrous-iron Oxidation Kinetics of a Predominantly Leptospirillum ferrooxidans Culture", Biotechnology and Bioengineering, 65, 44-53.

5. Clark, D.A., Norris, P.R., 1996, “Oxidation of Mineral Sulphides by Thermophilic Micro-organisms” Minerals Engineering vol. 9 no. 11, pp 1119-1125

6. Crundwell, F. K., 1997, “The kinetics of the chemiosmotic proton circuit of the iron-oxidising bacterium Thiobacillus ferrooxidans”, Bioelectrochemistry and Bioenergetics, 43, 115-122

7. Dew, D.W., van Buuren, C., McEwan, K., Bowker, C., 1999, “Bioleaching of Base Metal Sulphide Concentrates: A Comparison of Mesophile and Thermophile Bacterial Cultures”, In: Proceedings of International Biohydrometallurgy Symposium IBS ‘99, Ed: by R. Amils, A. Ballester, Madrid, Elsevier, pp. 229-238

8. Huberts R., 1994, "Modelling of ferrous sulfate oxidation by iron oxidising bacteria - a chemiosmotic and electrochemical approach", PhD Thesis, University of the Witwatersrand, Johannesburg, South Africa.

9. Ingledew W.J., 1982, “Thiobacillus ferrooxidans: The bioenergetics of an acidophilic chemolithotroph”, Biochimica et Biophyisca Acta, 683, 89-117

10. Jeffery, G.H., Basset, J., Mendham, J. and Denney, R.C., 1989, “Vogel’s textbook of quantitative chemical analysis”, 5th edition, Longman Scientific and Technical, New York.

11. Le Roux, N.W., Wakerley, D.S., 1988, “Leaching of chalcopyrite (CuFeS2) at 70 ºC using Sulfolobus”, In: Biohydrometallurgy: Proceedings of the International Symposium Warwick 1987, Ed. by P.R. Norris and D.P Kelly, University of Warwick, 305-319.

12. Meruane, G., Salhe, C., Wiertz, J., Vargas, T., 2002, A novel electro-chemical-enzymatic model which quantifies the effect of the solution Eh on the kinetics of ferrous iron oxidation of Acidithiobacillus ferrooxidans.”, 80, (3), 280-288

13. Nemati M. and Webb C., 1997, "A kinetic model for biological oxidation of ferrous-iron by Thiobacillus ferrooxidans", Biotechnology and Bioengineering, 53, (5). 478-486.

14. Pirt S.J., 1965, "The maintenance energy of bacteria in growing cultures", Proceedings of the Royal Society B, 163, 224-231.

15. Pirt S.J., 1982, "Maintenance energy: a general model for energy-limited and energy-sufficient growth", Archives of Microbiology, 133, 300-302.

16. Roels J.A., 1983, “Energetics and kinetics in biotechnology”, Elsevier Biomedical Press, Amsterdam, 20-31

17. van Scherpenzeel D.A., Boon M., Ras C., Hansford G.S. and Heijnen J.J., 1998, "Kinetics of ferrous-iron oxidation by Leptospirillum bacteria in continuous cultures", Biotechnology Progress, 14, 425-433.



1237

The role of microorganisms in dispersion of thallium compounds in the environment

A. Sklodowska, M. Golan and R. Matlakowska

Warsaw University, Faculty of Biology, Laboratory of Environmental Pollution Analysis, CEMERA – Centre of Excellence, Miecznikowa 1, 02-096 Warsaw, Poland

E-mail: [email protected]

Abstract Thallium is a highly toxic element and very rarely studied in the context of

environmental hazards connected with zinc and non-ferrous metal industry. Microorganisms naturally existing in post-flotation and smelt wastes can participate in thallium release from waste deposits and can contribute to its dispersion in the environment. Twenty-one isolates were obtained from wastes of a non-ferrous smelter in Southern Poland characterised by high heavy metal contamination. Ten isolates showed high activity in thallium leaching from wastes (post-flotation and smelt wastes) as well as from pure thallous sulphide. Additionally, cadmium and lead were bioleached from wastes. The isolated bacteria indicated thallium resistance at a concentration up to 100 mg/l and some of them were able to survive in good condition at a concentration of up to 4 g/l. The same bacteria were isolated from rivers and wastewater in this region. A preliminary characterisation of isolates was performed. It was shown that some petroleum products i.e. asphalt-base crude occasionally used for waste immobilisation at the edge of pond or flotation surfactants partially stopped the activity of sulphide oxidising bacteria.

1. INTRODUCTION Microorganisms naturally existing in post-flotation and smelt wastes can participate in

thallium release from waste deposits and can contribute to its dispersion in the environment.

The role of microbes in the dispersion of inorganic metal salts (especially sulfides) has been known for years. Oxidation of these compounds is the way of gaining energy, needed in many biochemical processes such as CO2 fixation etc. Metal ions are unused products of reactions, which can penetrate into the environment. Since the 70’s it has been known, that thiobacilli (today genus Thiobacillus is divided into a few genera, belonging to another subclasses of Proteobacteria [6]) can divide thallous sulfide, to obtain sulfide ions – the energy source [5]. Free thallous ions can penetrate into the environment, and take part in many biogeochemical cycles.

Microbes can methylate thallium ions, producing dimethylthallium – Me2Tl+. A method of estimating the concentration of this compound in environmental samples was worked out by Schedlbauer and Heumann [8]. The mechanism of this process is still unknown. Probably methylated cobalamine is needed, as in the mercury methylation [4].


1238

The role of this process in the environment is unknown. Me2Tl+ is very stable [8]. This can suggest, that it is the way of detoxification of the microbe’s environment, similarly to the methyllation of mercury [4].

For the majority of microbes thallium is highly toxic. By disturbing metabolic processes, it stops cell division. For Anacystis nidulans 15 ppm of thallium in medium completely inhibit all metabolic processes, for Chlamydomanas reinhardtii only 5 ppm [7].

Thallium is a highly toxic element and very rarely studied in the context of environmental hazards connected with zinc and non-ferrous metal industry. The research carried out in Southern Poland enabled the identification of several regions, which are seriously threatened by thallium as well as to indicate direct sources of pollution. Polluted regions included mainly the surroundings of the zinc smelter and post-flotation waste ponds.

2 MATERIALS AND METHODS

2.1 Collecting and storing of samples Samples of wastes of a non-ferrous smelter in Southern Poland were collected in

October 2001. The material was stored in single-use, sterile plastic tubes. Microbiological analysis was carried out within 24 hours. The remaining material was stored at -20°C.

2.2 Bacterial strains Ten strains were isolated from wastes. Non-modified strains of Halothiobacillus

neapolitanus and Paracoccus versutus received from The Department of Bacterial Genetics, Warsaw University were used as a reference due to their potential similarity to isolates from wastes.

2.3 Isolation of strains 5 g of fresh wastes were added to 50 ml of sterile Beijerinck’s medium. The mixture

was incubated on rotary shaker, at room temperature. After 24 hours, the culture was transferred to solid medium and cultivated at 28°C for 4 days. Single colonies were inoculated on fresh solid medium every 5 days.

2.4 Media 1. Beijerinck’s medium [2] with 10 g of Na2S2O3

.5H2O, as the only energy source and 2 ml of Tuovinen’s salts [10]. Solid medium contained 25 ml of 3% phenol red’s solution, pH = 7,5.

2. Modified LB medium [9] with 20 g of NaCl, pH = 7,5 3. Modified LB medium with thiosulphate [9] with addition of 20 g Na2S2O3

.5H2O, pH = 7,5.

4. Davis medium [3] with glucose, as carbon source. 5. Modified Beijerinck’s medium I with 10g of Tl2S instead of thiosulphate. 6. Modified Beijerinck’s medium II: 60 g of wastes, dried to dry weight in 105°C

instead thiosulphate, and without Tuovinen’s salts.


1239

2.5 Estimation of bacterial cell number in wastes 5 g of fresh wastes were added to 50 ml of sterile NaCl solution (0.9%) and stored on

laboratory rotary shaker for ½ hour. Solutions of this culture were then inoculated on solid media: Beijerinck, modified LB with NaCl and Davis. Plates were stored at 28°C for 7 days. After the incubation the colonies on every plate were counted.

2.6 Characterisation of isolates Isolated strains were inoculated on solid media to test their ability to grow in different

conditions. Beijerinck’s, modified LB, modified LB with Na2S2O3 and Davis media were used. All isolates were stained using Gram method.

The resistance of strains to thallous ions was tested. Bacterial strains were inoculated on solid modified LB medium with Na2S2O3, containing Tl+ (as TlNO3) in concentrations: 25, 50, 75, 100 ppm, or solid Beijerinck’s medium with analogous ratios of thallim. Paracoccus versutus and Halothiobacillus neapolitanus were inoculated on the same media.

2.7 Preparation of a mixture of strains Fresh cultures of all isolated strains on Beijerinck’s medium were prepared. After 5

days of incubation at room temperature on a laboratory shaker 1 ml of each culture was added to sterile medium without thiosulphate. This mixture was used as an inoculum in next experiments.

2.8 Experiments

2.8.1 Thallium bioleaching from thallous sulphide 100 ml of Beijerinck’s medium with 10 g of Tl2S was inoculated with mixture of

bacterial strains. Experiment was conducted in Erlenmayers’ flasks (300 ml) on a laboratory shaker at room temperature. Non-inoculated medium, cultivated under the same conditions, as cultures was control for this experiment. Two series of culture (designated 1st culture and 2nd culture) were prepared. Every day pH, Tl+ concentration and the number of bacterial cell were measured.

2.8.2 Thallium bioleaching from wastes 200 ml of Beijerinck’s medium with 60 g of wastes (dried at 105°C) was inoculated

with mixture of strains. Incubation was conducted in Erlenmayer flasks (500 ml), on a shaker at room temperature. Non-inoculated medium served as a control for this experiment. Two series of culture (designated 1st culture and 2nd culture) were prepared. The concentration of thallium, cadmium and lead was estimated. Additionally, the pH and the number of bacterial cell were assessed.

The number of bacterial cells was estimated after staining with DAPI and counting on filters under the epifluorescence microscope.

2.9 Chemical analysis The concentration of metals in acidified supernatant and mineralised wastes was

measured using Flame Atomic Absorption Spectrometer SOLAAR M6. Before the analysis wastes were dried at 105°C and mineralised in Millestone

Laboratory Microwave System with 65% HNO3 and 36% H2O2 (9:1).


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The analysis of thallium in soils was carried out with diisopropylether extraction, according to Asami et al. [1]. Mineralised wastes were filtered and distilled water was added, giving the total volume of 100 ml in the flask 50 ml were moved to Erlenmayer flask, 6 ml of HBr (36%) and 1 ml of 0,1% solution of CeSO4

.4H2O was added. After 15 minutes the total volume was placed into a separator (200 ml), 5 ml of ether was added and then shaken for 5 minutes. Organic phase was collected and evaporated. The sample was then dissolved in 5 ml of 3% HNO3

3. RESULTS AND DISCUSSION Concentration of heavy metals and pH of wastes of a non-ferrous smelter were

measured and presented in Table 1.

Table 1. Heavy metals concentration and pH of wastes

Concentration of heavy metals [mg/kg d.w.] pH Tl Cd Pb Zn

7.00 – 7.10 40 – 50 120 – 130 18000 – 21000 4500 – 5000

The bacterial cell number isolated from wastes able to grow on different media was estimated. For all media (Beijerink’s, modified LB and Davis) similar results were observed: 104-105 cells per mg of wastes (wet weight).

Twenty-one bacterial strains were isolated from wastes on Beijerinck’s medium. From all isolates, 10 were chosen on the basis of their growth ability. All isolates were Gram-negative, or Gram-variable (young cells were negative, and after 10 days of incubation, positive). Eight of strains were rods, and two of them were too small to identify their morphology under a light microscope.

Table 2. Characteristics of isolates

Growth on medium: Strain Gram Morphology

Beijerinck Modified LB Davis Modified LB with Na2S2O3

1. negative rod + - - + 2. negative rod + - - - 3. negative rod + - - - 4. variable rod + + + + 5. variable rod + - - + 6. variable rod + - - + 7. variable n/e + - - - 8. negative rod + - + + 9. negative rod + - - + 10. negative n/e + + + + n/e – not estimated

The isolated bacteria indicated thallium resistance at a concentration up to 100 mg/l

and some of them were able to survive in good condition at a concentration of up to 4 g/l (data not shown).


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Table 3. Resistance of freshly isolated bacterial strains to different concentration of thallium ions

Growth on medium with thallium in concentration: Strain 25 ppm 50 ppm 75 ppm 100 ppm

1. + + + + 2. + + + + 3. + + + + 4. + + + + 5. + + + - 6. + + + + 7. + + + + 8. + + + + 9. + + + - 10. + + + + H.neapolitanus + + - - P. versutus + - - -

The ability of isolates to thallium leaching from pure Tl2S was tested in the first experiment. Sulphide ion was the only energy source for microorganisms. The concentration of Tl+ in a supernatant indicated the rate of the leaching process. The highest concentration of Tl+ (1624 ppm) on the 9th day of cultivation was obtained (Fig. 1). In the 2nd culture the highest concentration was observed at the beginning of the experiment – 250 ppm. From the 5th day it droped to 110 ppm on the last day of cultivation. Throughout the experiment the biggest concentration was noticed in the 1st culture. For the first five days the concentration was stable 210-320 ppm, then it started to increase, reaching 1624 ppm on the 9th day in the 1st culture. This phenomenon may be explained by the irregular structure of the crystal, or some impurities in crystal net.

In both cultures brown, stable sediment, localised on flask walls, under the liquid line was observed.

Figure 1. Bioleaching of thallium from Tl2S – concentration of Tl ions in leaching solution

Throughout the experiment bacterial cell number systematically increased in both cultures. For both series of cultures similar results were obtained – from about 1 x 104/ml to about 7 x 105/ml (Fig. 2).


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0

20

40

60

80

1 2 3 4 5 6 7 8 9 10

time (days of experiment)

cel

l num

ber [

104 /m

l]

first culture second culture

Figure 2. Bacterial cell number per ml of culture during bioleaching of Tl2S

Figure 3 Bacterial cells stained with DAPI adhered to particle of wastes during bioleaching process

Figure 3 presents mixture of isolates cultivated in mineral medium containing wastes. The bacterial cells stained with DAPI attached to particle of wastes are visible.

In both cultures pH decrease of about 1 unit was observed. In the first culture the decrease from 7.40 to 6.50 and in the 2nd one from 7.00 to 6.20 was observed while in the control from 7.29 to 7.00. In both cultures pH was stable for the first 5 days, then it decreased (Fig.4.).

To check the ability of isolates to leach heavy metals (esp. thallium) from wastes, the experiment using modified Beijerinck’s medium II was carried out. Thallium concentration in wastes (mg/kg d.w.) was measured and its decrease showed the process efficiency. Additional parameters of process were: bacterial cell number, pH changes. The rate of cadmium and lead bioleaching were also measured.

Apart from the fouling of culture medium, no other differences between cultures and control images were obsewved during the experiment.

In the first culture 40% of thallium was leached. This was more than in the second culture, where it was only 25%. In the control less than 10% of thallium was leached. At


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the beginning of the experiment the concentration of thallium was different for both cultures and the control. Probably because the wastes are an unhomogenic blend with unspecified ratio of different compounds (Fig. 5).

Cadmium was bioleached in 100% in the 1st culture and more than in 70% in the 2nd one. In the control flask 17% decrease of the concentration was noticed (Fig. 6).

Figure 4. Changes of pH of supernatant during bioleaching of Tl2S

01020304050

1 4 7 10 13


Tl [

ppm

]

first culture second culture control

Figure 5. Thalium concetration in wastes during bioleaching experiment

0

50

100

150

1 4 7 10 13


Cd

[ppm

]


Figure 6. Cadmium concentration in wastes during bioleaching process For lead 70% decrease of the concentration for the 1st culture and more than 50% for

the 2nd one was obtained at the end of the experiment, while in the control about 25%. For


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lead, differences in the concentration at the beginning of the experiment, were smaller than for thallium and cadmium (Fig. 7).

05000

10000150002000025000

1 4 7 10 13


Pb [p

pm]


Figure 7. Lead concentration in wastes during bioleaching process Bacterial cell number during the experiment was similar for both cultures: from

3.20x106 to 8.05x107 per ml in the 1st one and 5.40x106 – 5.00x107 for the 2nd one. Lag-phase was longer in the 2nd culture, than in the 1st one (Fig. 8).

020406080

100

1 2 3 4 5 6 7 8 9 10 13


cell

num

ber

[106 /m

l]

first culture second culture

Figure 8. Bacteria cell number during bioleaching of wastes Throughout the experiment the pH was stable due to probable buffer capacity of

wastes that includes a significant amount of carbonate. The maximal ratio of changes was 0,8 unit in the 1st culture, 1 unit in the 2nd one and only 0,1 unit in the control (Fig. 9).

Isolated strains are able to leach thallium from both: pure salts and ores. Isolates are very interesting due to their ability to grow in high concentrations of thallium and other heavy metals, which are thought to be highly toxic. It was shown that microbes' activity might be one of the reasons of thallium contamination near ores treatment plants and near postflotation wastes deposits. This problem raises many new questions concerning the physiology of their resistance to metal ions. Additionally, isolated microbes might be used in the biohydrometallurgical methods of this metal extraction.

This research will be continued to understand the problem of the taxonomy and physiology of these isolates, and find possibilities of using them into biotechnological processes.


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Another issue which arises is how to inhibit microbes activity in the deposits, or how to retain dissolved metal ions within the deposit.

66,46,87,27,6

1 2 3 4 5 6 7 8 9 10


pH

first culture second culturecontrol

Figure 9. Changes of pH of supernatant during bioleaching of wastes

REFERENCES 1. T. Asami, C. Mizui, T.Shimada, M. Kubota, Fresenius J. Anal. Chem. 356 (1996) 348. 2. M.W. Beijerinck, Zentralbl. Bakteriol. Parasitenkd. Infenktionskr. Hyg. Abt. II , 11

(1904) 593. 3. D.H. Davis, M. Doudoroff, R.Y. Stanier, M. Mandel, Int. J. Syst.Bacteriol., 19 (1969)

375. 4. G.M. Gadd, Encyclopedia of Microbiology, Vol. 2, Academic Press, Inc., 1992. 5. M. Galizzi, E. Ferrari, Appl. Environ. Microbiol., 32 (1976) 433. 6. D.P. Kelly, A.P. Wood, Int. J. Syst. Evolution. Microbiol., 50 (2000) 511. 7. B. Lustigman, L.H. Lee, J. Morate, F. Khan, Bull. Environ. Contam. Toxicol., 64

(2000) 565. 8. O.F. Schedlbauer, K.G. Heumann, Appl. Organometal. Chem., 14 (2000) 330. 9. A. Sklodowska, R. Matlakowska, W. Ludwig, Acta Microbiol. Polon., Vol. 45 No 2

(1996) 131. 10. O.H. Tuovinen, D.P. Kelly, Arch. Microbiol., 111 (1973) 257.

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