Bioleaching of metals from soils or sediments

~

Pergamon

Wal. Sci. Tech. Vol. 37, No.8, pp. 119-127,1998. © 1998 fAWQ. Published by Elsevier Science Ltd

Printed in Great Britain.

PH: S0273-1223(98)OO242-X

0273-1223/98 $19'00 +
BIOLEACHING OF METALS FROM SOILS OR SEDIMENTS R. Tichy*, W. H. RUlkens**, J. T. C. Grotenhuis**, V, Nydl***, C. Cuypers* and J. Fajtlt

* Institute of Landscape Ecology, Czech Academy ofSciences, Na Sddktich 7, 37005 Ceske Budejovice, Czech Republic ** Agricultural University, Department of Environmental Technology, Bomenweg 2. P.O. Box 8129,6700 EV Wageningen, The Netherlands *** University of South Bohemia, Faculty ofAgriculture, Department ofMathematics and Statistics, StudentsM 13, 37005 Ceske Budejovice, Czech Republic t University of South Bohemia. Faculty ofAgriculture, Department of General Plant Nutrition, Studentskti 13, 37005 Ceske Budejovice, Czech Republic

ABSTRACT Bioleaching can be one of few techniques applicable for the removal of toxic metals from polluted soils or sediments. Its principle is a microbial production of sulphuric acid and leaching of metals with it. The use of bioleaching can benefit from the use of low-cost substrates and from a possi ble coupling to other processes of microbial sulphur cycle, like sulphate reduction to treat spent bioleaching liquor, or partial sulphide oxidation to recycle sulphur. For the evaluation of bioleaching, the existence of different leaching strategies is considered, i.e. intensive or extensive extraction. The intensive extraction uses high concentrations of acid at short extraction times, whereas low acid additions and long treatment times are used in extensive processes. On a reference study with wetland sediment receiving mine drainage we demonstrated that the hioleaching is a typical extensive process. The bioleaching experiments involved the use of the different sulphur substrates, i.e. orthorhombic sulphur flower and microbially produced, recycled sulphur from partial sulphide oxidation process. The laller type of sulphur substrate performed considerably beller. © 1998 IAWQ. Puhlished by Elsevier Science Ltd

KEYWORDS Bioleaching; extraction; sediment; sulphur; toxic metals. INTRODUCTION Toxic-metals contaminated soils or sediments have become recently a serious problem in both highly industrialized Western countries, as well as in transition countries of Central and Eastern Europe. The technological solutions for treatment of such sites are usually costly, sometimes even not feasible. This problem is pronounced in moderately contaminated sites, since the treatment efficiency of standard techniques decreases with lowering the overall concentration of metals. This results in efforts to apply innovative technological configurations, like landfarming, heap-leaching, in-situ extraction, or phytoremediation.

119

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R. TICHY et al.

Innovative treatment principles are therefore investigated. Main focus of the research is a possible use of more environmentally friendly treatment conditions, the use of biological processes, and development of proper post-treatment techniques which allow a recuperation of soil properties. In this respect, the microbial sulphur cycle, and its influence on toxic metals in the environment, gets increasing attention worldwide. Sulphur cycle (Figure I) consists of several microbial processes leading to different states of sulphur. These processes comply sulphate reduction, partial sulphide oxidation to elemental sulphur, and completed sulphur oxidation steps.

Contaminated SOil ...

su I§hunc

aCI

Sulphur oxidation

! elemental Tsulphur Sulphide partial oxidation

SUIPhi~

•

Extraction of metals

~Cleaned

soil • Metals , sulphate, acidity

,----'------.,

sulphide (

Sulphate reduction

Metals separation

~

sulp.hate acitlity

Figure I. Schematic representation of a microbial sulphur cycle.

Changes in sulphur oxidation state bring about changes in mobility of cationic toxic metals like cadmium, zinc, copper, cobalt, nickel, mercury. During sulphate reduction, resulting sulphide forms readily metal sulphides, which are poorly soluble in water. On the contrary, the terminal product of sulphur oxidation, i.e. sulphuric acid, mobilizes metals from the solid phase to the liquor. The latter process of sulphur oxidation occurs spontaneously in nature, causing sometimes rather serious environmental problems like acid mine drainage, mine-tailing leachates, or acid sulphate soils. However, it can also serve to mobilize metals under controlled conditions, ego during enrichment of low-grade ores (Bruynesteyn, 1989; Rawlings and Silver, 1995), coal desulphurization (Bos and Kuenen, 1990; Rossi, 1993; Loi et aI., 1994), or removal of toxic metals from contaminated sludge (Couillard and Mercier, 1990; Sreekrishnan and Tyagi, 1995), soil or sediment (van der Steen et ai., 1992; Tichy et aI., 1993). This removal is called bioleaching. This work evaluates a possible use of bioleaching for treatment of a wetland sediment receiving minedrainage waste water. This sediment retains toxic metals under anaerobic conditions due to various processes like sulphate reduction, alkalization, precipitation, and adsorption (Machemer and Wildemann, 1992; Eger, 1994; Gambrell, 1994; Farmer et aI., 1995). An introduction of air into the system can happen e.g. during a period of draughts, as a result of dikes break, or during dredging and further handling of the sediment when the basin fills up with solid material and has to be regenerated (Gambrell, 1994). Once the air is introduced, a rapid decrease of pH is encountered, followed by increased solubilization of toxic metals. This was repeatedly demonstrated for the polluted river sediment (Maass and Miehlich, 1988; Calmano et aI., 1992; Forstner, 1995) and wetland sediment (Gambrell et aI., 1991; Evangelou and Zhang, 1995). The processes involved during sediment oxidation comply oxidation of reduced sulphide-containing compounds (Marnette et aI., 1992), oxidation of ferrous iron and precipitation of ferric hydroxides (Forstner, 1995). Acidification of the environment results in further dissolution of non-sulphidic metal precipitates and in desorption of cationic metals (Forstner 1995). By the abovementioned processes, the polluted sediment can posses hazardous properties and its treatment is required.

Bioleaching of metals

121

The key parameter in bioleaching is the pH of a sediment slurry: the lower are pH values, the higher are metal extraction yields. To achieve sufficiently low pH, proper amount of reduced sulphur or ferrous iron must be present in the treated material. In some cases, the indigenous concentration of reduced sulphur or ferrous-iron containing compounds in the treated material is not high enough to reach sufficiently low pH. Therefore, additional substrate has to be introduced, like ferrous iron (Couillard and Mercier, 1990) or elemental sulphur (Tichy et aI., 1993; Tyagi et aI., 1994; Sreekrishnan and Tyagi, 1995). In our previous works (Tichy et al., 1994; Tichy et at., 1997) we demonstrated a novel bioleaching substrate containing elemental sulphur. This material appears as a waste product during the microbial treatment of sulphidecontaining waste water (Buisman et al., 1990; Janssen et aI., 1995). This process is a part of microbial sulphur cycle (Figure I: sulphide partial oxidation). The microbially produced sulphur provides considerably higher specific surface area, and higher hydrophillicity, compared to the orthorhombic sulphur flower. This leads to much higher oxidation rates of microbially produced elemental sulphur, compared to the sulphur flower (Tichy et aI., 1994; Janssen et at., 1996). To investigate the possible use of bioleaching for treatment of mine-drainage loaded wetland sediment, we carried out the firstly the experiments with adding sulphuric acid. These results were compared to the bioleaching tests. MATERIALS AND METHODS Sediment: The sediment used in our study originates from a wetland close to settlement Lukavice (district Chrudim, 110 km East from Prague, Czech Republic). The system evolved naturally inside of a pool with broken dike closure, and its age is estimated for 20-30 years. The wetland receives water from a granite mine with elevated content of pyrite. We used the sediment from anoxic layer, which was characterized by dark-grey colour. The sediment was taken after the screening of the upper, aerobic (orange) layer and ca. top 10 cm of the anaerobic layer, and in IO-L plastic buckets transported to the laboratory. Prior to the experiments, the sediment was manually homogenized and sieved wet at a I mm mesh. Leaching tests with sulphuric acid: The slurry for the leaching tests was prepared by mixing the fresh anoxic sediment with distilled water at ratio I g fresh weight: 2.5 mL water. The dry solid: liquid ratio was 0.0603: I (g:mL). 400 mL of the suspension was distributed into the 500 mL glass bottles. The bottles were closed with seals penetrated by four 5 mm glass tubes to ensure the exchange with atmosphere and proper oxidative status of the slurry. The suspension was acidified by 1M H2 S0 4 to achieve acid concentrations of 5; 10; 20; 30; 60; 100; 150; 200; 250 mM. The bottles were agitated in a rotatory horizontal shaker (100 rpm). Bioleaching tests: The experimental setup for bioleaching tests was identical to that applied for leaching tests with acid. Instead of sulphuric acid, additions of 0, 0.1, 0.5, 1, and 5 g of elemental sulphur per I litre of sediment slurry were applied. This corresponds to the addition of 0, 3.125, 15.625, 31.25, and 156.25 mM of SO. Orthorhombic sulphur flower (Sulphur praecipitatum) was obtained from Lachema Bmo, Czech Republic. Microbially produced elemental sulphur was produced by a pilot plant treating sulfide-rich wastewater in Eerbeek, The Netherlands. Sulphur suspension was decanted, twice washed with aliquot volumes of distilled water, and dried at 60°C. Its content of elemental sulphur was 97.5% by weight, the rest being microbial biomass debris and other impurities originating from the sulphide-containing wastewater. Experiments were carried out in two independent runs. Analyses: The total Cd, Cu, Fe, Pb, and Zn content in the sediment was determined after HN0 3/H 20 2 wet digestion using the U.S. EPA method 3050 (US EPA, 1987). pH of the sediment slurry was measured with a Hanna Instruments combined pH-electrode HI 1332 B, connected to a Radelkis (Budapest, Hungary) precision digital pH-meter (OP-208/1). The pH-measurements were done directly in the sediment slurry, kept in suspension by slow swirling. To determine the dissolved metals, slurry was centrifugated (15 minutes, 3000 rpm) and filtrated over a Synpor 5 m membrane filter (Pragochema, Prague, Czech Republic). ICP/AES (PU 7450, Leemans Laboratories, U.S.A.) method was used to determine metals in the liquor.

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Numeric data processing: For all data processing, a standard statistical software package Statgrafics 2.6 was applied. Non-linear regression was performed by least-square iteration using Marquart method. The endcriterion for iterations was set as change of least square < I0- 14.

RESULTS AND DISCUSSION Leaching tests with sulphuric acid The pH of the sediment slurry after the addition of sulphuric acid was obviously affected by two major trends: a) the increase of acidity due to oxidative changes, and b) the decrease of acidity due to the pHbuffering, protonation of solid components, proton-adsorption and diffusion into the particles. These two trends worked adversely and resulted in a rather complex behaviour of pH in the sediment slurry, see experimental points in Figure 2. After the addition of H 2S0 4 at concentrations 100 mM H 2S0 4, the pH immediately started to increase. After 5 hours, however, the pH in these treatment begun to decrease. In treatments 150 mM H 2S0 4 , the subsequent pH-decrease was not recorded.

pH

7,..---------------------------, 6'

5

4

3 2

* *

llL-----'------..J----.....l------'------'------.l

o

5 o

20

10 !'.

20 o

80 60 100 Time (hours) 60 100 150 200 250mMH2S04 40

30

'"

.

...

.

*

Figure 2. Changes of pH of a sediment slurry after addition of sulphuric acid.

To predict the pH of a sediment slurry as a function of time and concentration of added acid, we developed a model as follows: (P4'Y + P5·t) pH = Pl + P2'log(Y) + P3'e

P6

+ 1 + P7.t

(I)

Here, Y denotes a concentration of added sulphuric acid (mM), t represents the time of extraction (hours), and PI, P2, P3, P4, P5, P6, P7 are model parameters. The log symbol stands for a natural logarithm and e is a base of natural logarithm. Non-linear fitting of this model yielded the correlation coefficient of r2:0.982 for 121 data points. The curves predicted by the model are plotted in Figure 2. The values of model parameters are summarized in Table I. The concentrations of solubilized metals in the liquor followed changes of pH and time. To involve both parameters, we applied the exposure variable, i.e. time multiplied by protons activity. The exposure was successfully demonstrated as a control variable e.g. in solubilization of pyrite by acid (van der Zee and de Wit, 1993). The concentrations of Cd, Cu, Pb, and Zn followed a clear sigmoidal dependence on exposure (data not shown). Therefore, we applied a regression using modified Gompertz function:

Bioleaching of metals

123

03' ( t· ( w}) 04 Me (mg/L)

= e

(2)

01 + 02'8

Here, Me denotes a concentration of particular metal in the solution, t is time (hours), (H+) is the activity of protons (mM) calculated from the slurry pH, and QI, Q2, Q3, Q4 are model parameters. The formula has operation range at to(H+»O. Resulting values of parameters for the four studied metals are summarized in Table 2. Table I. Fitted parameters in the equation I. The correlation coefficient r 2 for 121 data points yielded 0.982 PI

5.308

± 0.270

P2

-0.3719 ± 0.0195

P3

8.2404 ± 1.0180

P4

-0.00257 ± 0.00047

P5

-0.04063 ± 0.00200

P6

-6.7078 ± 1.3060

P7

0.06363 ± 0.0095

Table 2. Fitted parameters in the equation 2. Cd

Cu

Pb

Zn

Ql

-4.690±0.821

-2.151 ±2.418

-1.301 ± 1.122

-1.226±0.690

Q2

3.405 ±0.847

5.694±2.478

4.749±1.lt9

3.928±0.690

Q3

-0.557±0.189

-1. 976± 1.530

-2.663±1.1oo

-0.545±0.134

Q4

-0.382±0.O75

-0.494±0.095

-0.530±0.069

-0.487±O.049

r'

0.892

0.942

0.912

0.964

N=

ltD

110

106

109

Table 3. Total metals in the sediment, theoretical and actual maximum concentrations of solubilized metals, and maximum extraction yields. Cd

Cu

Pb

Zn

576.1

356.2

Mean (mg/kg)

6.402 485.2

Standard deviation

0.453

8.1

31.9

10.3

Them. maximum in solution (mglLt

0.386

29.31

34.76

21.49

Actual maximum in solution (mg/L?

0.229

19.61

18.85

13.27

66.9

54.2

61.7

Maximum extraction yield (%)C

59.3

A: Total metal content in the sediment multiplied by the solid:liquid ratio. B: Concentration of metal in the solution after 100 hours extraction with 250 mM H,SO, C: Percentage of actual maximum soluble metal from theoretical maximum.

124

R. TICHY et al.

The maximum concentrations of solubilized metals achieved in our experiments approached the theoretical maximum calculated from the total content of metals in the sediment. These values are represented in Table 3. The actual maximum concentrations reach 54.2-66.9% of the total metal content. The 54.2% solubilization of Pb is rather unexpected result, since PbS0 4 is poorly soluble in water (Evans, 1989). However, the low pH-values probably enforced Pb solubilization, so that its concentrations in the liquor were percentically comparable to other metals. Combined the two models, i.e. Equations I and 2, the efficiency of extraction with sulphuric acid can be simulated. In Figure 3, this simulation is presented with H 2S0 4 concentrations equal to molar concentrations of elemental sulphur used later in bioleaching tests, and an extra 250 mM H 2S0 4 level. The resulting extraction curves revealed three different phases: I. the initial, nearly horizontal line, 2. the increasing part, and 3. terminal phase with the curve asymptotically approaching to a horizontal line. For Cd and Zn, the first phase occurred within the first 10 hours of leaching, and the third phase took place at time >80 hours. Similar findings for Cu and Pb were found only with the highest simulated concentration of H 2S0 4 , Lower concentration showed low extraction yields, and the typical three regions were not pronounced. The maximum applied H 2S0 4 concentration, indicated by thick lines in Figure 3, showed nearly horizontal line for Cd and Zn. It shows an increase with time for Cu, Pb, however, the three phases noticed above were also not pronounced.

Extraction (%)

100 Copper 80 60 .------_ . ...... :.::.: ..- - - - 40 .........; .. ---- ...----- - - ~ " " 20 '-"'" " ........ :;-=" " 0" 80 100 40 60 20 80 100 0 60 20 .40 Time (hours) 100 100 Zinc Lead 80 80 60 60 ~";-;-.:-..;-.:-.:-.;:-..:...:-..: 40 40 " " 20 20 ,,, " -..... :.::: :;':.-- ---" -0 0 20 40 60 80 100 80 100 0 40 60 20 0 Concentrations of acid: 3.125 15.625 31.25 156.25 250 mM

100 80 60 40 20 t 00

Cadmium

_-

.~,,;

.........

I

.,'

Figure 3. Simulated extraction yields of Cd, Cu, Pb, and Zn at varying acid additions.

Our results indicate that at short extraction times, i.e. within 20 hours, high concentrations of sulphuric acid must be applied to gain a significant extraction yield. A prolonged extraction with high acid concentration is redundant, since it does not substantially improve the extraction yield. Comparably high extraction yield can be obtained with lower concentrations of acid, however, at prolonged extraction. These findings indicate two different strategies applicable for leaching, i.e. intensive, which uses high concentrations of acid at short extraction, and extensive, which uses long extraction with low acid concentrations. This corresponds to our previous findings with leaching of clay, silt, and sandy soils artificially polluted with zinc (Tichy et ai., 1996), and seems to have a general validity.

Bioleaching of metals

125

Bioteaching tests The pH recorded during bioleaching tests with microbially produced sulphur and orthorhombic sulphur flower is presented in Figure 4. To demonstrate a relevance of the pH-changes for the bioleaching efficiency, the 50% extraction isolines for Cd, Cu, Pb, Zn were calculated from our model (Eq. ) and added to the Figure 4. In all treatments, the microbially produced sulphur acidified faster than the sulphur flower. This phenomenon was described earlier for a sediment-free water suspension (Tichy et ai., 1994), and for conditions of undisturbed soil profile (Tichy et at., 1997). For both types of sulphur, the treatment with I and 5 g SOIL yielded similar results. It means that the concentration of I gIL approached the maximum oxidative capacity of the indigenous microbial community, and higher addition of sulphur was redundant. The microbial sulphur oxidation is controlled by the process of microbial adhesion to the sulphur particles surface (Kelly, 1982). Therefore, the sulphur surface area and the total number of sulphur oxidizing microbes in the suspension are crucial fro the acidification rate. If the surface area far exceeds the total number of microbes in the system, addition of extra elemental sulphur will not affect the intensity of acidification. This effect caused likely the negligible difference between I gIL and 5 gIL elemental sulphur treatments.

pH 7

pH 7

BioI. sulphur

6

.---.....--.

S. flower

6

5

5

4

4 3

/;::;:~~:

II / /

50

Time (hours) Control 0.1 0.5 1 5 9 S per litre •

••.• -

2 100 150 200 250 300 0 . 50

0

0

'"

100 150 200 250 300

Time (hours) Cd Cu Pb

Zn

•

Figure 4. Bioleaching with microbially produced sulphur and sulphur flower with indicated 50% extraction isolines for Cd, Cu, Pb, and Zn.

Generally, the differences among acidification with different concentrations of microbially produced sulphur were less pronounced, compared to the sulphur flower. This also suggests the limitation by maximum oxidizing capacity of the microflora, since the microbially produced sulphur provides considerably higher surface area, compared to the sulphur flower (Tichy et at., 1994). Therefore, much lower additions of microbially produced sulphur reached maximum oxidizing capacity. CONCLUSIONS We demonstrated that two different strategies can be applied for the leaching of toxic metals from the polluted wetland sediment, i.e. intensive and extensive. Similar findings with acidic extraction of different soils suggest that this phenomenon has a general validity. The intensive leaching is performed at high levels of acid (above 150 mM) and extremely low pH «2.5) at short extraction times (max. 20 hours). The extensive leaching is performed at high extraction times (>80 hours), however, at low concentrations of acid and higher pH. Eventual use of bioleaching can be denoted as the extensive strategy. The batch leaching process, i.e. the development of pH and solubilization of Cd, Cu, Pb, Zn, was described using two regression models. The models use two independent variables, i.e. the extraction time and the concentration of added acid. These models offer a future perspective when both independent variables are

126

R. TICHY et al.

assigned their economic value. This will ultimately lead to a simple cost-benefit analysis, which is highly needed for the true evaluation of different leaching strategies, including the bioleaching. The use of microbially produced elemental sulphur as a substrate for bioleaching yielded considerably better results than the orthorhombic sulphur flower. As this effect has been repeatedly observed, we conclude that its use for bioleaching offers an attractive alternative. ACKNOWLEDGEMENT This work was funded by the Grant Agency of the Czech Republic, grant No. 525/96/0671, 'Heavy metals in wetland sediments: risk assessment and possible sanitation'. REFERENCES Bos, P. and Kuenen, J. G. (1990). Microbial treatment of coal. In:Microbial metal recovery. Ehrlich H.L., Brierley c.L. (eds.), McGraw-Hill, New York, pp. 343-377. Bruynesteyn, A. (1989). Mineral biotechnology. J. Biotech. 11, 1-10. Buisman, C. J. N., Geraats, B.G., Ijspert, P. and Lettinga, G. (1990). Optimization of sulfur production in a biotechnological sultide-removing reactor. Biotech. Bioeng. 35, 50-56. Calmano, W., Hong, J. and Forstner. U. (1992). Einfluss von pH-Wert und Redoxpotential auf die Binding und Mobilisierung von Schwermetallen in kontaminierten Sedimenten. Wasser 78, 245-257. Couillard. D. and Mercier. G. (1992). Metallurgical residue for solubilization of metals from sewage sludge. J. Environ. Engn. 118, 808-813. Eger, P. (1994). Wetland treatment for trace metal removal from mine drainage: the importance of aerobic and anaerobic processes. Wat. Sci. Tech. 29/4,249-256. Evangelou. V. P. and Zhang, Y. L. (1995). Pyrite oxidation mechanisms and acid mine drainage prevention, Crit. Rev. Env. Sci. Tech. 25,141-199. Evans, L. J. (1989). Chemistry of metal retention by soils. Environ. Sci. Tech. 23.1046-1056, Farmer, G. H., Updegraff D.M., Radehaus P.M. and Bates E.R. (1995). Metal removal and sulphate reduction in low-sulphate mine drainage. In:Bioremediation of inorganics. Hinchee R.E.• Means J.L. and Burris D.R. (eds.), Battelle Press, Columbus, pp. 17-24. Forstner, U. (1995). Non-linear release of metals from aquatic sediments. In: Biogeodynamics ofpollutants in soils and sediments. Salomons W. and Stigliani W.M. (eds.), Springer Verlag. Berlin, pp. 247-308. Gambrell, R. P., Wiesappe, J. B.• Patrick, Jr. W. H. and Duff, M. C. (1991). The effects of pH. redox, and salinity on metal release from a contaminated sediment. Wat. Air Soil Pollut. 57·58, 359-367. Gambrell, R. P. (1994). Trace and Toxic Metals in Wetlands - A Review. J. Environ. Qual. 23, 883-891. Janssen, A. J. H., Sleyster, R., van der Kaa, c.. Jochemsen, J.• Bontsema, J. and Lettinga, G. (1995). Biological sulphide oxidation in a fed-batch reactor. Biotech. Bioeng. 47. 327-333. Janssen, A. J. H., de Keizer, A .• Van Aelst, A., Fokkink, R .• Yangling. H. and Letlinga. G. (1996). Surface characteristics and aggregation of microbiologically produced sulphur particles. Coil. Surf. B: Biointeifaces 6. 115-129. Kelly. D. P. (1982). Biochemistry of the chemolithotrophic oxidation of inorganic sulphur. Phil. Trans. R. Soc. Lond. B 289. 499-528. Loi, G., Mura, A., Trois. P. and Rossi, G. (1994). The Porto Torres depyritization pilot plant: Light and shade of one year operation. Fuel Pmc. Technol. 40, 261-268. Maass, B. and Miehlich. G. (1988). Die Wirrkung des Redoxpotentials auf die Zusammnesetzung der Porenlosung in Hafenschlicks-fcldern. Mitt. Dtsch. Bodekunde Ges. 56, 289-294. Machemer, S. D. and Wildeman, T. R. (1992). Adsorption compared to sulphide precipitation as metal removal process from acid mine drainage in a constructed wetland. J. Contam. Hydr. 9, 115-131. Marnette, E. C., Hordijk, c.. van Breemen, N. and Cappenberg. T. (1992). Sulfate reduction and S-oxidation in a moorland pool sediment. Biogeochemistry 17, 123-143. Rawlings, D. E. and Silver, S. (1995). Mining with Microbes. Biotechnology 13. 773-778. Rossi. G. (1993). Biodepyritization of coal: Achievements and problems. Fuel 72. 1581-1592. Sreekrishnan, T. R. and Tyagi, R. D. (1995). Sensitivity of Metal-Bioleaching Operation to Process Variables. Process Biochem. 30.69-80. Tichy, R.• Grotenhuis J. T. c., Janssen, A.• van Houten, R., Rulkens, W. H. and Letlinga. G. (1993). Application of the sulfur cyclc for bioremediation of soils polluted with heavy metals. In: Arendt F., Annokee, G. J., Bosman. R. and van der Brink, W. J. (eds.):Proc. Int. Conj. Contaminated Soil '93. May 1993. Berlin. Kluwer Academic Publishers. Dordrecht, 1461-1462. Tichy, R., Janssen, A.. Grotenhuis, 1. T. c., Letlinga. G. and Rulkens, W. H. (1994). Possibilities for using biologically-produced sulfur for cultivation of thiobacilli with respect to bioleaching processes. Bioresour. Technol. 48, 221-227.

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Tichy, R., Grotenhuis, J. T. c., Rulkens, W. H. and Nydl, V. (1996). Strategy for leaching of zinc from artificially contaminated soil. Environ. Techno/. 17, 1181-1192. Tichy, R., Fajtl, J., Kuzel, S., Kohir, L. (1997). Use of elemental sulphur to enhance a cadmium solubilization and its vegetative removal from contaminated soil. Nutrients Cycling in Agroecosystems 46, 249-255. Tyagi, R. D., Blais, 1. F., Deschenes, L., Lafrance, P. and Villeneuve, J.P. (1994). Comparison of Microbial Sulfuric Acid Production in Sewage Sludge from Added Sulfur and Thiosulfate. J. Environ. Qual. 23, 1065-1070. US-EPA (1987). Reference manual for analytical methods for water, soil, and sludge. Method 3050. U.S. Epa, Office of Research and Development, Washington D.C. van der Steen, J. J. D., Doddema, H. J. and de Jong, G. (1992). Extraction of heavy metals from waste streams using thiobacilli. (in Dutch) Ministry of Housing, Physical Planning and Environment (VROM), Directoraat-Generaal Millieubeheer, Report 1992, 10(10), 28 p. van der Zee, S. E. A. T. M. and de Wit, J. C. M. (1993). Theoretical and practical aspects of soil chemical behaviour of contaminants in soil. In:Migration and fate of pollutants in soils and subsoils. Petruzzelli D. and Helfferich F.G. (eds.), NATO ASI Scries, vol. G 32, Springer-Verlag Berlin, pp. 27-45.