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Remediation of heavy metal-contaminated sediments by solid-bed bioleaching C. Löser 7 H. Seidel 7 P. Hoffmann 7 A. Zehnsdorf

Abstract Weisse Elster River sediment from the Leipzig Lowlands region (Saxony, Germany) is anthropogenically polluted by heavy metals. Sediment dredged from a trap to the south of Leipzig was characterized in detail. When freshly dredged sediment contacts air, the material turns acidic because of oxidation processes, the heavy metals become soluble and the sediment poses an environmental risk. We are therefore developing a sediment-treatment process based on heavy metal removal by bioleaching. Leaching experiments were carried out in suspension and in the solid bed. The heavy metals were solubilized to nearly the same extent by H2SO4 dosage (pure chemical extraction) and addition of elemental sulphur (microbial oxidation of S 0 to H2SO4). With increasing dosage of the leaching agent, Zn, Cd, Ni, Cu and Cr were more and more solubilized, whereas Pb was only dissolved in small amounts. The addition of 2% S 0 is considered an optimum dosage. When 5% S 0 was added to the sediment, the pH dropped to 1.76 and large amounts of undesirable compounds such as Ca, Al and Fe were solubilized. The higher the temperature, the faster the metals were solubilized in both suspension and the solid bed. The temperature optimum for activating the indigenous S 0-oxidizing microbes of the sediment lies between 30 and 40 7C. Conditioning of freshly dredged sediment with plants makes it suitable for solid-bed leaching; the kinetics of heavy metal solubilization from sediment conditioned for 6 months with Phragmites australis was the same as from long-term stored sediment. Keywords Aquatic sediment 7 Bioleaching 7 Heavy metals 7 Percolation principle 7 Solid bed

Received: 26 November 1999 7 Accepted: 5 May 2000 C. Löser (Y) 7 H. Seidel 7 P. Hoffmann 7 A. Zehnsdorf UFZ Centre for Environmental Research Leipzig-Halle, Dept. of Remediation Research, Permoserstrasse 15, 04318 Leipzig, Germany

Introduction Heavy metal-contaminated aquatic sediments are a major environmental problem, particularly in densely populated regions and industrialized centres. The waters of Saxony (Fig. 1) alone contain F18 million tonnes of sediments, of which 6 million tonnes have to be urgently removed to maintain the hydraulic function of the rivers. These sediments are in part polluted with heavy metals. For example, a 2.5-km-long stretch of the Weisse Elster River within the city of Leipzig (known as the Elsterbecken) bears 330,000 tonnes of sediment with a metal content of 1296 t Zn, 164 t Cr, 94 t Pb, 81 t Ni, 79 t Cu, 14 t Co, 9.2 t Cd and 2 t Ag (Müller and others 1998). When dredged anoxic sediment contacts air, the material turns acidic because of oxidation processes, the heavy metals become soluble and the sediment poses an environmental risk. Today, dredged sludge is stored for some time in drained basins to reduce its water content and then removed by landfill disposal. However, in view of the limited capacity, deficient sustainability and the prospect of increasing costs of landfill disposal, there is a need for ecologically consistent methods of cleaning up heavy metal-contaminated sediments. Heavy metal-polluted solids can be cleaned up by using physical and chemical processes: – extraction with mineral acids (Müller and Riethmayer 1982; Strasser and others 1995); – extraction with organic acids such as acetic, citric and oxalic acid or with complexing agents (Höll 1995; Thöming and others 1996; Fröhlich and others 1999; Stichnothe and others 1999); – size classification and treatment of the fine fraction by flotation (Venghaus and Werther 1996). A remediation technique that is very promising from an economic point of view is the treatment by bioleaching. Bioleaching is, strictly speaking, a combination of microbial and abiotic, i.e. purely chemical, processes (Bosecker 1997; Ehrlich 1997; Schippers and Sand 1999). The bioleaching of metals from low-grade ores is a wellestablished technology in mining (Krebs and others 1997; Rawlings 1997). During the last decade, the removal of heavy metals from contaminated sewage sludge (Blais and others 1992; Tyagi and others 1993; Benmoussa and others 1994; Coillard and others 1994a; Strasser and others 1995; Shooner and Tyagi 1996;), soil (Gourdon and Funtowicz 1995) and dredged sediment (Calmano and

Environmental Geology 40 (4-5) February 2001 7 Q Springer-Verlag

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Fig. 1 Map of Saxony showing the Weisse Elster River system and the location of the Kleindalzig detritus trap where the examined sediment originates

Fig. 2 Scheme of the sediment treatment concept

Ahlf 1988; Coillard and Chartier 1993; Coillard and others 1994b; Seidel and others 1995a,b, 1996, 1997, 1998) by microbial leaching has been intensively studied. However, the bioleaching of contaminated solids has been mainly investigated under conditions far removed from practicable remediation processes due to: leaching of water-suspended material, low solid content in suspension, the use of expensive and/or ecologically unfavourable additives in high concentrations such as ferrous sulphate, initial acidification, and often inoculation with Thiobacillus cultures. We are now developing a multi-stage sediment treatment process with the intention of applying it on a large scale (Fig. 2). The sediment is first divided into an almost unpolluted coarse fraction and a highly contaminated fine fraction, which is subjected to bioleaching. For economic reasons, only solid-bed leaching according to the percolation principle, similar to the dump leaching of low-grade ores as described, e.g. by Bosecker (1997), Brombacher and others (1997) and Krebs and others (1997), is applicable. Unlike crushed ore, the fine fraction of freshly dredged sediment is practically impermeable to liquids and therefore unsuitable for treatment in a solid bed (Seidel and others 1998). It was observed that dredged sediment that was stored in the open changed its structure 644


and became easily permeable to water when conditioned by suitable plants. Based upon this knowledge, a conditioning step was introduced to prepare freshly dredged sediment for solid-bed leaching. The solubilized metals accumulate in the aqueous phase, from which they are removed by process water treatment. The last step consists in the revitalization of the cleaned-up sediment by addition of lime and compost (the latter prepared from plant material obtained in the conditioning step) to produce a soil-like material that can be used for recultivation of former mining areas. Because the economy of the whole remediation process is mainly determined by the rate and efficiency of the bioleaching step, the objective of this study was to examine the kinetics of heavy metal solubilization under various leaching conditions from Weisse Elster River sediment as a model material.

Cases and solutions

Characterization of heavy metal-polluted Weisse Elster River sediment Our examination of heavy metal-polluted sediment concentrates on material from the Weisse Elster River. This river originates from the Erzgebirge Mountains, runs through the Vogtland Hollow and the Leipzig Lowland region and flows into the Saale River near the city of Halle. The Saale River is a major tributary to the Elbe River (Fig. 1). The Weisse Elster River has a length of 248 km and a drainage area of F5400 km 2 (Müller and others 1998). The detritus of the river comes mainly from eastern Thuringia, the western Erzgebirge Mountains and the Vogtland Region and is in part polluted with heavy metals. The Weisse Elster sediment used in our experiments was freshly dredged from the Kleindalzig detritus trap (see Fig. 1) or taken from a 6-year-old sediment deposit on a neighbouring site. The trap consists of a broad basin to settle grit, sand and suspended matter that are transported by the river. This material consisted of 15% gravel, 75% sand and 10% clay and silt (sieve analysis of sediment that was freshly dredged in 1998). The clay and silt fraction is much more polluted by heavy metals than the other fractions. The fine fraction of Weisse Elster sediment contains large amounts of clay minerals. The crystalline part of the sediment fraction ~20 mm consists of 45% illite, 19% quartz, 11% plagioclase, 9% kaolinite, 7% chlorite, 7% potash feldspar and 1% haematite (Müller and others 1998). The sediment also comprises varying portions of amorphous inorganic compounds and 15 to 20% organic matter (determined as loss on ignition at 550 7C). The organic matter is difficult to degrade by microbes as its portion did not significantly change during long-term storage of dredged sediment under aerobic conditions. The Weisse Elster sediment from the Kleindalzig trap is highly contaminated by heavy metals (Table 1) because the metal content is, in general, one to two magnitudes higher than the shale standard of Turekian and Wedepohl (1961). This standard corresponds to an average elemental composition of natural, i.e. unpolluted, marine clay sediment. The increased heavy metal content is largely caused by anthropogenic causes such as mining, chemical and manufacturing industries as well as municipal wastewater flows. The main contaminants that exceed the maximum permissible limits are zinc and cadmium. The high uranium content can be explained by former uranium mining activities in the upper reaches of the river. An example of a geogenically increased compound is silver, which originates from the Erzgebirge Mountains. The submerged sediment is black, muddy-pasty and anoxic, its sulphur compounds are reduced and the heavy metals are practically immobile (Table 2). When dredged material is stored in the open, it becomes covered with plants that develop spontaneously from the natural seed potential of the sediment. The stored sediment loses wa-

Table 1 Heavy metal content of Weisse Elster River sediment dredged from the Kleindalzig trap in 1993 and then stored in the open for 6 years on a neighbouring site, in comparison with the shale standard and the LAGA Z2 limits (all data related to dry mass) Metals

Zinc Manganese Chromium Copper Lead Nickel Uranium Cobalt Cadmium Silver Mercury

Zn Mn Cr Cu Pb Ni U Co Cd Ag Hg

In dredged sediment a (mg/kg)

Shale standard b (mg/kg)

LAGA Z2 limit c (mg/kg)

3320 957 536 352 324 287 107 55 39 5 ~2

95 850 90 45 20 68 3.7 19 0.3 0.07 0.4

1500 – 600 600 1000 600 – – 10 – 10

a The sediment was sieved to ~2 mm, dried, ground to a particle size of ~60 mm and then analysed by wavelength dispersive X-ray fluorescence spectroscopy b This standard is the average metal content of shale according to Turekian and Wedepohl (1961) c LAGA Z2 limits are the maximum permissible levels for land disposal of soils in Germany (LAGA 1996)

Table 2 Changes in the physicochemical parameters of Weisse Elster River sediment from the Kleindalzig trap during storage for 6 years in the open Parameter

Total solids Redox potential S oxidation degree pH value Mobile Zn content a Mobile Ni content a Mobile Cd content a

(%) (mV) (%) (mg/kg) (mg/kg) (mg/kg)

Freshly dredged sediment

Longterm stored sediment

20 P250 9 7.0 ~0.5 ~0.8 ~0.5

65 c380 65 5.6 230 15 1.5

a

These values were estimated as follows: natural-moist sediment of 20 g dry weight was extracted with 0.2 l water, including the pore water of the sediment, on a laboratory shaker at 12 rpm and 25 7C. Freshly dredged sediment was eluated under an atmosphere of nitrogen to avoid any oxidation. After 24 h of mixing, the metal content of the aqueous phase was analysed by inductively coupled plasma atomic emission spectroscopy. Finally, the mass of each dissolved metal was related to the dry mass of the extracted sediment

ter and is oxidized by atmospheric oxygen (increasing redox potential and the degree of sulphur oxidation; Table 2). As a result of the oxidation processes, the sediment turns acidic and the metals become more soluble. Freshly dredged sediment consists mainly of small particles and is therefore practically impermeable to water


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stream of 5 l/h (water-saturated at the cultivation temperature by using a gas-washing bottle) was continuously passed through the system to ensure aerobic conditions. The course of the leaching process was followed by measuring the pH value and conductivity of the aqueous phase and by analysing liquid samples taken at intervals by means of inductively coupled plasma atomic emission spectroscopy (ICP-AES) with an ELAN 5000 unit (PerkinElmer Corp.) according to the German DIN 38406 E22 method. The solubilized mass of a metal, mMe(t), was calculated from the metal content and the volume of the aqueous phase samples:

3

4

mMe(t)pCMe(t) 7 VL(tp0)PAVS c A[CMe(t)i7VS] Fig. 3 Particle size distribution of Weisse Elster River sediment from the Kleindalzig trap immediately after dredging and after 6 months of storage in the open (particle size measurements by laser diffraction spectrometry)

(Löser and others 1999a). However, sediment stored in the open turns into a greyish-brown, crumbly, soil-like material and becomes easily permeable (decrease in water resistance by way of pore and particle enlargements (Löser and others 1999b)). The plant cover of the sediment is of great importance in structural changes. Sediment from which all spontaneously arising vegetation has been removed by hand, dries and oxidizes but does not significantly change its particle-size distribution (Fig. 3).

Methods of bioleaching The Weisse Elster River sediment used in leaching experiments was either freshly dredged and then conditioned with Phragmites australis for 6 months or stored in the open for 6 years (sediment characteristics as described above). The leaching was observed both in suspension and in the solid bed. In the case of suspension leaching, naturally moist sediment of 20 g dry weight was put into a 500-ml conical flask, and was in part supplemented with 0.1 to 1 g elemental sulphur, mixed with water or sulphuric acid (altogether 200 ml aqueous phase, including the pore water of the sediment) and then shaken at 130 rpm at a constant temperature. Solid-bed leaching was carried out in a laboratory percolator consisting of a column and a storage vessel, both made of glass and having a double wall for thermostating. The column, 0.1 m in diameter and with a sieve insert, was loaded with sediment of 1 kg dry weight (pure or with addition of 20 g sulphur) and the vessel was filled with water (altogether 2 l aqueous phase). During the leaching process, water from the vessel was percolated through the solid bed at a rate of 0.15 l/h and an air 646


i

(1)

i

In this equation, CMe(t) denotes the metal content of the aqueous phase, VL(tp0) is the liquid volume at the beginning and VS symbolizes the volume of the withdrawn aqueous samples. The solubilized metal fraction was calculated from the mMe(t) value and from the mass of the metal of interest in the sediment at the beginning of leaching.

Results and discussion Shake-flask experiments with water-suspended sediment represent a very useful method for examining the kinetics of leaching processes in the laboratory, although the remediation of heavy-metal-polluted material on a large scale is only economically favourable in the case of solidbed leaching. Metals of the Weisse Elster sediment were first leached in suspension with sulphuric acid (Fig. 4). In the case of acid-free suspension, metals that were already mobile in sediment were simply extracted. With increasing H2SO4

Fig. 4 Heavy-metal solubilization after 28 days of suspension leaching of long-term stored Weisse Elster River sediment without elemental sulphur in various sulphuric acid dosages (leaching of water-suspended sediment in shake flasks at 20 7C)

Cases and solutions

dosage, Zn, Cd, Ni and Cu were more and more solubilized, but Cr and Pb were only dissolved after addition of large amounts of H2SO4. Although the heavy metals were to a large extent extracted from suspended matter by sulphuric acid, this extraction method was ineffective in solid-bed leaching. A large acid supply, as is necessary for quick metal solubilization, needs a high sulphuric acid content of the percolated liquid and/or an intensive percolation (Löser and others 1999a). However, highly acidic conditions (i.e. low pH values) lead to an increased mobilization of undesirable non-toxic compounds such as Ca, Al and Fe, which interfere with the process water treatment. Furthermore, solid-bed permeability to water limits the percolation flow. Moreover, the higher the sediment package (an economical need), the more acid has to be supplied to the sediment. Another problem of pure acid leaching is that sulphide-bound metals are only partly dissolved by acids. As an alternative to direct acid supply to the sediment package by the percolation flow, it is also possible to produce sulphuric acid by microbial oxidation of elemental sulphur in the solid bed. The advantage of this method consists in the homogeneous acidification and heavy metal solubilization throughout the whole sediment layer. The amount of sulphur to be added to the sediment is of great practical interest and was at first examined with water-suspended sediment (Fig. 5). The indigenous microorganisms oxidized the sulphur to form sulphuric acid, which in turn reduced the pH within 53 days from 5.62 to values between 4.35 and 1.76, depending on the dosage of elemental sulphur (without S 0 addition no changes in pH occurred). During bioleaching with elemental sulphur, the heavy metals were dissolved to nearly the same degree as in the case of abiotic metal extraction with H2SO4 when compared on the basis of equal additions of mol-related S compounds (see Figs. 4 and 5).

Fig. 5 Heavy metal solubilization after 53 days of suspension leaching of long-term stored Weisse Elster River sediment with various amounts of elemental sulphur (leaching of water-suspended sediment with 0 to 5% of elemental sulphur in shake flasks at 20 7C)

Fig. 6 Heavy metal solubilization (sum of Zn, Ni, Cd, Cu, Cr and Pb) during suspension leaching of long-term stored Weisse Elster River sediment with 2% of elemental sulphur at various temperatures (leaching of water-suspended sediment in shake flasks)

This indicates that the metals are largely fixed in compounds other than sulphides. The more elemental sulphur was added to the sediment, the larger the portion of solubilized toxic metals was. However, high sulphur doses produced strongly acidic conditions, which also led to an extensive mobilization of non-toxic metals, e.g. a 5% S 0 addition resulted in a specific solubilization of 5 g/kg Ca, 10 g/kg Al and 13 g/kg Fe after 53 days of leaching. The addition of 2% sulphur to the sediment is therefore considered an optimum value as it ensures the solubilization of a large portion of the main contaminants Cd and Zn, but limits the mobilization of undesirable metals, especially of Fe. The temperature is considered to be one of the main factor determining the kinetics of bioleaching. For this reason, the bioleaching of water-suspended Weisse Elster sediment supplemented with 2% S 0 was observed at various temperatures (Fig. 6). The higher the temperature, the faster the heavy metals were solubilized. At 40 7C the leaching kinetics was nearly the same as that found at 30 7C. The optimum temperature for the indigenous S 0oxidizing microbes of the sediment was therefore between 30 and 40 7C, provided that no other factor such as CO2 or O2 limited the bacterial growth. Even at temperatures as low as 4 7C metals were leached at a significant rate; in the case of naturally moist sediment supplemented with 2% S 0 and stored at 4 7C, the average leaching rate of Zn, Ni, Cd, Cu, Cr and Pb was F0.3% a day. With the aim of scaling up, the dependence of bioleaching on the temperature was also observed in the solid bed, using sediment with 2% S 0. Changes in process water parameters such as pH value, conductivity, metal content, and sulphate concentration during solid-bed leaching using the laboratory percolator have been previously described (Löser and others 1999a). The higher the tem-


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Fig. 7 Heavy metal solubilization (sum of Zn, Ni, Cd, Cu, Cr and Pb) during solid-bed leaching of long-term stored Weisse Elster River sediment with 2% of elemental sulphur at various temperatures (sediment of 1 kg dry weight leached with altogether 2 l of water using a laboratory percolator)

perature, the faster the desired metals accumulated in the percolated process water and the higher was the mobilized metal content at the end of leaching (Fig. 7). Apart from the transient stagnation of leaching, the kinetics of metal solubilization from the solid bed was comparable with that from suspended matter. This means that the kinetics of metal solubilization during solid-bed leaching of sediment with elemental sulphur was determined by microbial rather than transport processes. Examination of bioleaching at high temperatures is of great practical importance as microbial sulphur oxidation produces large amounts of heat, which, in turn, may substantially increase the temperature of a sediment package. When solid-bed leaching was carried out with 1000 kg sediment and 2% S 0 in a pilot plant, the temperature rose to 42 7C, which was 24 K higher than the surrounding temperature. Heavy metal-polluted material that can be remediated by leaching is predominantly submerged aquatic sediment. This sediment of a muddy–pasty consistency is anoxic, practically impermeable to water and, therefore, unsuited for solid-bed leaching. Consequently, the structure of freshly dredged sediment has to be improved by conditioning with plants (as regards the structural changes and the increase in permeability to water see above) before solid-bed treatment by the percolation principle becomes applicable to this material. To demonstrate the applicability of solid-bed leaching to conditioned sediment, freshly dredged sludge that had been pre-treated with Phragmites australis for 6 months was supplemented with 2% sulphur and then leached at 25 7C using the laboratory percolator. For purposes of comparison, this experiment was also carried out with sediment stored for 6 years in the open. Both sediments were permeable to water and showed nearly the same leaching kinetics (Fig. 8). 648


Fig. 8 Heavy metal solubilization (sum of Zn, Ni, Cd, Cu, Cr and Pb) during solid-bed leaching of Weisse Elster River sediment with 2% elemental sulphur (sediment of 1 kg dry weight leached with altogether 2 l of water at 25 7C, using a laboratory percolator). The sediment was dredged and then either conditioned with Phragmites australis for 6 months or stored in the open for 6 years

Conclusions The leaching of heavy metals from Weisse Elster River sediment, taken from a trap to the south of Leipzig (Saxony, Germany), was observed in suspension and in the solid bed, using shake flasks or laboratory percolators. Acidic conditions were produced by adding either sulphuric acid to the aqueous phase or elemental sulphur to the sediment. After reaching equilibrium, the amount of leached metals was almost independent of the leaching procedure applied (suspension or solid bed) and on the kind of leaching agent used (H2SO4 or S 0). The kinetics of leaching, however, was largely influenced by both the procedure and the leaching agent used. In leaching with S 0, the kinetics of metal solubilization is determined by the microbial oxidation rate of sulphur to sulphuric acid, which takes 20 or more days to reach the stationary phase, regardless of the procedure used. Leaching with H2SO4 proceeds very fast in suspension (about 2 days to reach equilibrium) but needs much longer in the solid bed. In the latter case, the rate of H2SO4 influx into the solid bed is restricted by its permeability to water and by the minimum acceptable pH value of the percolated aqueous phase. Furthermore, unlike all other methods examined, solid-bed leaching with H2SO4 produces a pH gradient and therefore inhomogeneous metal solubilization. Taking all points into account, large-scale remediation of heavy-metal-polluted sediment seems only possible by solid-bed leaching of sediment supplemented with elemental sulphur. The application of optimized process parameters is very important for effective bioleaching. The temperature opti-

Cases and solutions

LAGA (1996) Mitteilungen der Länderarbeitsgemeinschaft Abfall Nr.20 – Anforderungen an die stoffliche Verwertung von mineralischen Reststoffen/Abfällen – Technische Regeln. Erich Schmidt Verlag, Berlin Löser C, Seidel H, Hoffmann P, Zehnsdorf A (1999a) Remediation of heavy-metal polluted river sediments by bioleaching using the percolation principle. In: Schutter G De (ed) Characterisation and treatment of sediment. Proceedings of the CATS4 Conference Antwerpen, 15–17 Sept, Technologisch Instituut, pp 213–222 Löser C, Zehnsdorf A, Hoffmann P, Seidel H (1999b) Conditioning of heavy-metal-polluted river sediments by helophytes. Int J Phytoremediation 1 : 339–359 Müller A, Hanisch C, Zerling L, Lohse M, Walther A (1998) Schwermetalle im Gewässersystem der Weißen Elster. Acknowledgements This work was financially supported by the Abhandlungen der Sächsischen Akademie der Wissenschaften Deutsche Bundesstiftung Umwelt (AZ 12099). We sincerely zu Leipzig, Mathematisch-naturwissenschaftliche Klasse, vol thank Dr. P. Morgenstern and Dr. R. Wennrich (UFZ Leipzig58(6). Akademie Verlag, Berlin Halle) for their heavy metal analyses. Müller G, Riethmayer S (1982) Chemische Entgiftung – das alternative Konzept zur problemlosen endgültigen Entsorgung schwermetallbelasteter Baggerschlämme. Chem-Zeitung 106 : 289–292 Rawlings DE (1997) Biomining: theory, microbes and industrial processes. Springer, Berlin Heidelberg New York Schippers A, Sand W (1999) Bacterial leaching of metal sulBenmoussa H, Tyagi RD, Campbell PGC (1994) Biolixiviation fides proceeds by two indirect mechanisms via thiosulfate or des métaux lourds et stabilisation des boues municipales: efvia polysulfides and sulfur. Appl Environ Microbiol fet de la forme du soufre élémentaire utilisé comme substrat. 65 : 319–321 Rev Sci l’Eau 7 : 235–250 Seidel H, Ondruschka J, Stottmeister U (1995a) Heavy Blais JF, Tyagi RD, Auclair JC, Huang CP (1992) Comparimetal removal from contaminated sediments by bacterial son of acid and microbial leaching for metal removal from leaching: a case study on the field scale. In: Brink WJ van municipal sludge. Water Sci Technol 26(1–2) : 197–206 den, Bosman R, Arents F (eds) Contaminated soil. Kluwer, Bosecker K (1997) Bioleaching: metal solubilization by microDordrecht, pp 1039–1048 organisms. FEMS Microbiol Rev 20 : 591–604 Seidel H, Ondruschka J, Weißbrodt E, Stottmeister U Brombacher C, Bachofen R, Brandl H (1997) Biohydrome(1995b) Reinigung schwermetallbelasteter Sedimente durch tallurgical processing of solids: a patent review. Appl Microbakterielle Laugung – ein Behandlungskonzept. 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mum for the indigenous microbes of the studied sediment lies between 30 and 40 7C. Although a high dosage of S 0 results in an intensive heavy metal solubilization, it also leads to the mobilization of undesirable compounds, which interfere with the process water treatment and increase costs. A dosage of 2% S 0 is therefore considered optimum. Further examination of bioleaching will concentrate on the influence of oxygen and carbon dioxide supply on the leaching rate. Another question is how the indigenous microbes of the sediment react to higher temperatures because the core sediment temperature of a pilot-scale percolator exceeded 40 7C.

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