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Effects of bacterial activities on the release of heavy metals from contaminated dredged sediments
Chemosphere 56 (2004) 619–630 www.elsevier.com/locate/chemosphere
Effects of bacterial activities on the release of heavy metals from contaminated dredged sediments C. Lors *, C. Tiffreau 1, A. Laboudigue CNRSSP, BP 537, 930 Boulevard Lahure, 59505 Douai Cedex, France Received 29 April 2003; received in revised form 27 February 2004; accepted 20 April 2004
Abstract The potential impact of indigenous bacterial processes on the release of heavy metals from dredged sediment deposits was investigated. Batch re-suspension experiments were conducted in order to investigate the release of Zn, Cd, Cu and Pb from a polluted anoxic sediment submitted to oxidative perturbations. The concentrations of heavy metals, sulphate and dissolved organic carbon (DOC) were periodically recorded, and cell counts were performed to follow the evolution of several bacterial species. The specific effects of microbial processes were quantified by performing resuspension assays on sterilised samples. Moreover, the effect of an initial acidification of the system was studied. The results showed that metal release was mainly due to oxidation of sulphide minerals contained in the sediment. Sulphur-oxidising bacteria such as Acidithiobacillus thiooxidans were identified to play a major role in the process, by enhancing the oxidation kinetic. However, the acid production resulting from these reactions was almost totally buffered by the dissolution of the calcite present in the sediment. Copper was released to a lesser extent, and a strong association with organic matter was observed. Lead was not observed in solution, because of its low solubility at neutral conditions and of its re-adsorption on the solid phase. The initial acidification of the system resulted in an faster growth of the acidophilic A. thiooxidans. A subsequent pH drop originating from microbial processes was then observed during the first stages of the experiment. As a consequence, drastic increases in metal (Zn, Cd) release were observed. 2004 Elsevier Ltd. All rights reserved. Keywords: Dredged sediments; Heavy metal release; Acidithiobacillus; Thiobacillus; Re-suspension; Bacterial activity
1. Introduction Due to erosion, urban and industrial run-offs, waterways are subject to a constant silting up. As a consequence, it is necessary to undertake regular dredging operations, in order to keep the waterways in good conditions. These operations generate about
*
Corresponding author. Tel.: +33-3-2771-2681; fax: +33-32771-0707. E-mail address: [email protected] (C. Lors). 1 Present address: CEA Cadarache, DED/SAMRA/LTCR, 13108 St. Paul lez Durance.
250 000 m3 of sediments each year for the French NordPas de Calais region (Bogusz, 1997). However, because of their high sorption capacity, sediments act as a sink for toxic substances (heavy metals, organic pollutants) (F€ orstner and Kersten, 1989; Calmano et al., 1996). Consequently, dredged materials are frequently contaminated above safety limits and can represent a secondary source of pollution. As a matter of fact, sediments are generally deposited along streams or in deposit sites because of the cost of other treatment options. However, disposal of dredged material can modify the initial physico-chemical and microbiological conditions prevailing inside the sediments, simply because initially anoxic sediments become progressively
0045-6535/$ - see front matter 2004 Elsevier Ltd. All rights reserved. doi:10.1016/j.chemosphere.2004.04.009
Quartz (SiO2 ), Calcite (CaCO3 ), Albite (NaAlSi3 O8 ), Microcline (KAlSi3 O8 ), Kaolinite, Illite Mineralogical major phases (XRD)
273 5
Mn
2. Materials and methods
50 000 4000
P
3600 200
Fe
18 000 1000
TOC
6.5 0.5
%
53 6
Humidity
pH
oxygenated once they are deposited. This induces major changes in the oxydo-reduction potential, the pH and the speciation of major and trace elements (Tack et al., 1996, 1999). This also promotes the growth of aerobic bacteria playing a role in metal mobilisation, particularly those of genus Acidithiobacillus (Bosecker, 1997). These bacteria are responsible for the oxidation of insoluble metal sulphides to soluble metal sulphates with a high acidification of sediments (Blais et al., 1993; Tichy et al., 1998; Chen and Lin, 2001). As a consequence, this can lead to modifications of biosorption and bioleaching equilibria and to enhanced leaching of the metallic pollutants from the deposit sites (Petersen et al., 1997). The aim of this study is to evaluate the influence of the bacterial activity on the solubilisation of heavy metals contained in dredged sediments. More precisely, the objectives are to investigate the occurrence of indigenous sulphur-oxidising bacteria in sediments and to assess the potential of this microflora to leach toxic metals from dredged materials. The impact of an acidification of the medium (acidic rainfall, acidic run-off, etc.) on the metal solubilisation was also studied.
7.2 0.1
C. Lors et al. / Chemosphere 56 (2004) 619–630
Concentration
620
150 10 190 10 770 15
Cu Cd
6600 300
Pb Zn Ca Unit
Table 1 Heavy metals concentrations and main characteristics of the dredged sediment
The sediment has been sampled using a mechanical shovel (between 0 and 50 cm depth) from the Scarpe channel near the city of Raches (North of France). Once dredged, samples were immediately placed in plastic buckets, and stored at 4 C. The samples were bottled with their initial moisture level (i.e. no drying or draining was performed) and kept saturated in their porewater in order to maintain anoxic conditions during storage. Prior to experiments, the sediments were manually homogenised and wet sieved through a 2 mm mesh. Heavy metal concentrations and the main characteristics of the sediment are reported in Table 1. The Ca, Zn, Pb, Cd, Cu, Mn, P and Fe contents were determined according to the PR NF ISO 14869-1 standard: 0.5 g of dried sediment was heated to 450 C during 4 h and then allowed to cool down. The calcinated solid was then moistened with 4 ml of deionized water and transferred to a Teflon PTFE beaker. 10 ml of hydrofluoric (HF) and 3 ml of perchloric (HClO4 ) acids were added to the sediment, and the mixture was let to settle for 12 h. The beaker was then heated at 150 C during 2 h and evaporated to near dryness. Digests were diluted after filtration (0.45 lm) to 100 ml with a 2% HCl solution. Ca, Zn, Pb, Cd, Cu, Mn, P and Fe in the solution were then determined using Inductively coupled Plasma-Atomic Emission Spectrometry (ICP-AES) (Jobin Yvon, France, JY 138 Ultrace). The total organic carbon (TOC) was determined with a TOC
mg/kg dry sediment
2.1. Sediment sampling and characterisation
C. Lors et al. / Chemosphere 56 (2004) 619–630
analyser (Shimatzu, TOC 5000A with SSM 5000A module) according to the NF ISO 10694 standard. The TOC was obtained by subtracting inorganic carbon (IC) from total carbon (TC). The moisture content was determined by gravimetry, according to the NF ISO 11465 standard. The initial pH of the sediment was obtained following the NF ISO 10390 protocol. However, special care was taken in order to avoid oxidation during the measurement: a 1:5 solid/liquid suspension was prepared with fresh sediment and deaerated deionized water inside a N2 saturated glove box. The pH was measured inside the glove box after a 6 h stirring period, using a combined electrode (Ingold). 2.2. Batch experiments Leaching experiments were conducted both on fresh anoxic sediments, and on sterilised sediments in order to determine the part of abiotic processes in metal mobilisation. For sterilisation, some weighted amounts of sediments were placed in closed bottles and sterilised by gamma-ray irradiation (42 kGy). The slurry for the leaching tests were prepared by mixing fresh anoxic or sterilised sediment with a sterile leaching solution ([CaCl2 ] ¼ 0.1 mM; [NaNO3 ] ¼ 0.2 mM) at a ratio of 2 g wet sediment to 100 ml leaching solution. This solution was used in order to (i) avoid the potential impact of distillated water on bacterial cells due to osmotic factors, (ii) simulate the background concentration of the local rain water. For both fresh and sterilised sediments, two leaching solutions were used: one with neutral pH conditions and the other one with an initial pH previously adjusted to 4 with 0.1 N nitric acid. For sterilised sediments, this preparation was carried out under a sterile laminar hood. Batch experiments were carried out in 250 ml glass bottles and continuously agitated on a horizontal shaker at 20 C. The bottles were capped with porous foam rubbers in order to ensure gas exchange between the bottle contents and the atmosphere and to prevent microbiological contamination. For each of the four conditions studied (sterilised, non-sterilised, neutral solution, acidified solution), six samples were prepared, each sample being prepared in triplicate (i.e. 4 · 6 · 3 ¼ 72 total samples). The release of metals in solution was monitored as a function of time (respectively at day 1, 7, 14, 27, 62, 98 for biotic conditions and at day 1, 7, 16, 23, 28, 76 for abiotic conditions). At each date, the triplicate microcosms of each of the four conditions were sacrificed for
621
analysis. For each microcosm, a 1 ml suspension sample was withdrawn under agitation for counting of total, sulphur-oxidising and sulphate-reducing bacterial populations. The rest of the microcosm was centrifuged at 13 000 rpm for 30 min and the supernatant was sampled for physico-chemical analyses (pH, dissolved organic carbon (DOC), SO2 4 , metals). The pH of the leachate was measured using a pHmeter (Consort C833) equipped with a temperature correction sensor and a combined Ingold electrode. The total Zn, Cd, Cu, Fe, Pb, Mn, Al, Ca, DOC and sulphate contents in the supernatant were determined after a 0.45 lm additional filtration. Metal concentrations were measured by ICP-AES (Jobin Yvon 138 Ultrace), the DOC was determined using a Shimatzu TOC-5000A using catalytic oxidation with an infrared CO2 quantifier in accordance to the NF T 90-102 standard. Sulphate ions were quantified by a colorimetric method according to the NF T 90-040 reference protocol. 2.3. Microbiological analyses Enumeration of total bacterial and sulphur-oxidising bacterial populations were carried out on microplates of 96 wells from Nunc (Nunclon Delta, Germany), on which each well contained 250 ll of specific medium. The Nutrient broth (NB, Difco, USA) medium (diluted to 1/10) was used for total bacterial count. The medium of Razzel and Trussel (1963) in addition with a solution of FeSO4 Æ 7H2 O (c ¼ 20 g/l) was used for Acidithiobacillus ferrooxidans enumeration. For acidophilic sulphur-oxidising bacteria, the modified medium 317 (Chen and Lin, 2001) was combined with a color indicator (bromocresol green, c ¼ 0:02 g/l). The complete compositions of these two media are described in Table 2. The pH of acidophilic sulphur-oxidising bacteria and A. ferrooxidans medium was adjusted respectively to 4.5 and 1.8 with H2 SO4 (1 N). Each well was inoculated with 25 ll of suspension sample diluted to 100 –108 with Ringer solution (Ramsay, 1984). The inoculated microplates were incubated at 30 C during 15 days and then evaluated for colored products. Enumeration of bacteria was carried out using the Most Probable Numbers (MPN) method (De Man, 1977). Sulphate-reducing bacteria were counted in ready-touse culture tubes (Fisher Scientific, ref. A80 29406) containing a solid medium specific to sulphate-reducing bacteria (American Petroleum Institute, 1965). The
Table 2 Composition of medium (g/l of deionized water) for acidophilic sulphur-oxidising (ASO) bacteria and A. ferrooxidans growth ASO bacteria A. ferrooxidans
KH2 PO4
K2 HPO4
(NH4 )2 SO4
MgSO4 Æ 7H2 O
FeSO4 Æ 7H2 O
CaCl2
Na2 S2 O3
– 0.4
3.5 –
0.3 0.8
0.5 0.2
0.018 20
0.25 –
10 –
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tubes were previously heated to 100 C to liquefy the medium and then cooled down to 55 C for inoculation. Each of them was inoculated with 1 ml of sample diluted to 100 –102 . Three repetitions were performed. After incubation at 30 C during 15 days, the number of sulphate-reducing bacteria of the tested sample was determined from the black coloration of the medium using the MPN method (De Man, 1977).
considered that metals are associated with sulphidic and organic compounds. As a consequence, oxidation of dredged material is generally considered to cause dramatic modifications of the heavy metal mobility (Calmano et al., 1993; F€ orstner, 1995; Tack et al., 1996; Stephens et al., 2001). More recently, the use of spectroscopic methods has directly evidenced the occurrence of heavy metal sulphides in anoxic river sediments (Dood et al., 2000; Large et al., 2001; Isaure et al., 2002). More specifically, FeS2 (pyrite), ZnS (sphalerite) and PbS (galena) have been clearly identified in sediments originating from the same sampling spot as the one chosen for this study (Isaure et al., 2002). As shown in Fig. 1, aerating the sediment in biotic conditions leads to a marked release of sulphate in solution, which is probably due to the oxidation of sulphide solid phases. Although the oxidation of sulphidic minerals is generally accompanied by a large proton release (F€ orstner, 1995), the pH decrease observed in our case is quite low (from about 8 to 7). As a matter of fact, the acidification is buffered by the dis-
3. Results and discussion 3.1. Batch experiments in neutral medium 3.1.1. Metal mobilisation in leaching experiments Different studies based on chemical extraction schemes have shown that heavy metals present in anoxic sediments are generally associated with phases depicted as ‘‘residual’’ or ‘‘oxidable’’ (Kersten and F€ orstner, 1986; Giani et al., 1994; Cauwenberg and Maes, 1997; Yu et al., 2001). From these observations, it is generally
300
10
Biotic condition
Biotic condition Abiotic condition
Abiotic condition
250
[SO4 ] (mg/l)
8
200 150
2-
pH
9
100
7 50 0
6 0
20
40
60
80
0
100
20
Time (days) 50
60
80
100
80
100
200 Biotic condition
Biotic condition
Abiotic condition
40
Abiotic condition
150
[Ca] (mg/l)
[DOC] (mg/l)
40
Time (days)
30
20
100
50 10
0
0 0
20
40
60
Time (days)
80
100
0
20
40
60
Time (days)
Fig. 1. Evolution of pH and sulphate, organic carbon, calcium concentrations in leachates for experiments in neutral biotic and abiotic conditions.
C. Lors et al. / Chemosphere 56 (2004) 619–630
solution of calcite, which is present in large amounts in our sediment. Consequently to this dissolution, calcium is released in solution with a kinetic comparable to that of sulphate ions. Beside, it has been observed the monitoring of an experimental deposit, that the release of sulphate and calcium in solution can lead to saturation with respect to gypsum (CaSO4 ), and thus, to a limitation of the Ca and SO2 concentrations in solution 4 (Tiffreau et al., 1999). However, the concentrations observed in the present study are far from the saturation equilibrium, which is typically observed for concentration values of 1.5 g/l SO2 and 0.5 g/l Ca. In biotic 4 conditions, the oxidation of the metal-bearing sulphidic phases results in a release of high amounts of zinc in solution during the first 62 days, reaching more than 4 mg/l (Fig. 2). Cadmium follows the same evolution as zinc in proportional amount. In contrast, the proportion of released copper appears to be much lesser than for Zn and Cd. This is consistent with the observations made by Stephens et al. (2001). This lower mobility is attributed to the higher affinity of Cu for organic matter and for
623
ferric (hydr)oxides (haematite, goethite, ferrihydrite) that can form during the oxidation of the system. In abiotic conditions, the release of sulphate and calcium is much lower than in the presence of microorganisms (Fig. 1). This confirms that in the absence of microbial activity, chemical oxidation of sulphide minerals proceeds to a much lesser extent than in biotic conditions (Petersen et al., 1997). Moreover, the concentrations of calcium and sulphate do not show any increase as a function of time. This seems to indicate that only a small fraction of the sulphide pool seems to be readily available for chemical oxidation, compared to the biochemical one. As a result, the release of Zn and Cd in solution is also much lower than in biotic conditions (Fig. 2), illustrating the major role of bacterial activity in the leaching of these metals. Concerning copper, a completely different behaviour can be observed in Fig. 2. In abiotic conditions, a higher release of Cu is observed, which exhibits a linear trend as a function of time. The same discrepancy between biotic and abiotic conditions can be noticed for the DOC
6
0.20 Biotic condition
Biotic condition
5
Abiotic condition
Abiotic condition
4
[Cd] (mg/l)
[Zn] (mg/l)
0.15
3 2
0.10
0.05 1 0.00
0 0
20
40
60
80
0
100
20
Time (days)
60
80
100
80
100
0.10
0.10 Biotic condition
Biotic condition
Abiotic condition
Abiotic condition
0.08
[Pb] (mg/l)
0.08
[Cu] (mg/l)
40
Time (days)
0.06
0.04
0.06
0.04
0.02
0.02
0.00
0.00 0
20
40
60
Time (days)
80
100
0
20
40
60
Time (days)
Fig. 2. Evolution of zinc, cadmium, copper and lead concentrations in leachates for experiments in neutral biotic and abiotic conditions.
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adsorption is supposed to proceed via pure adsorption or coprecipitation with (hydr)oxides of Fe, Al or Mn which form at neutral pH conditions. In our case, this is supported by the fact that neither Fe, nor Mn, nor Al were detected in solution, indicating that the solution was saturated with respect to the (hydr)oxides of these elements.
evolution (Fig. 1). Moreover, when the Cu concentration in solution is plotted as a function of the DOC concentration, both parameters appear to be strongly linearly correlated (R2 ¼ 0:97). This suggests that Cu is associated with some organic compounds, which are released in solution in sterile conditions. Although the association of copper with organic matter present in dredged sediment is frequently reported (Kersten and F€ orstner, 1986; Calmano et al., 1993), the reason for the DOC release in solution remains obscure. Luo et al. (2001) have reported a similar phenomenon (i.e. release of DOC and copper in solution) for a 10 kGy gammaray sterilisation performed on a sandy loam. They suggested that the DOC release was due to the decomposition of killed organisms by surviving species, or that the sterilisation could render some parts of the nonmicrobial soil organic matter more decomposable. In our case, it is to note that such a release has not been observed in another series of batch experiments conducted on sediments from the Scarpe River sterilised with a lower dose (35 kGy) (Caille, 2002). Moreover, since no bacterial activity was detected in sterile samples during the whole experiments (see next paragraph), it is more likely that gamma-ray sterilisation performed for this study might have caused a radio-chemical degradation of the organic matter contained in the sediment. Compared to Zn, Cd and Cu, Pb was not detected during the experiment in both sterile and non-sterile conditions (Fig. 2). These results support the observations made by F€ orstner (1995) and Caille et al. (2003), who have reported that lead is released only during the very early hours of the re-suspension experiments and is then quickly re-adsorbed on the solid phases. This re-
1.E+12
3.2. Evolution of bacterial populations during the bioleaching experiment The results of total bacterial microflora and specific microflora countings (sulphur-oxidising and sulphatereducing bacteria) are reported in Fig. 3. The results concerning sterile microcosms are not shown, since they indicated that no microbial activity was detected during the whole experiments. This allows to ensure that sterile conditions were maintained during the re-suspension experiment. In non-sterile experiments, the total bacterial microflora is relatively abundant (7 · 108 bacteria/g of dry sediment) and stable during the experiment. Obviously, this indicates that the sediment possesses nutritive qualities for optimal bacterial activity. During the re-suspension experiment, the sulphatereducing bacterial population decreases progressively over time, down from 104 to 101 bacteria/g dry sediment. This can be attributed to the establishment of aerobic conditions, which are unfavourable to the growth of these strict anaerobic bacteria. For sulphur-oxidising bacteria, A. ferrooxidans is not detected during the experiment whereas an increase in acidophilic sulphur-oxidising bacteria is observed during
Total bacterial microflora A. thiooxidans Sulfate-reducing microflora
Bacteria / g dry sediment
1.E+10
1.E+08
1.E+06
1.E+04
1.E+02
1.E+00 0
20
40
60
80
100
Time (days) Fig. 3. Evolution of total bacterial and specific (sulphur-oxidising and sulphate-reducing) populations during the neutral leaching experiment.
C. Lors et al. / Chemosphere 56 (2004) 619–630
time, after a latency time of 7 days. Based on the study of 21 different sewage sludges, Blais et al. (1993) have shown that among all the acidophilic sulphur-oxidising bacteria, A. thiooxidans is the only species involved in the leaching of heavy metals. As a consequence, it can be considered that the growth observed in our study is due to the development of A. thiooxidans. The latency period seems to be an adaptation period of these microorganisms to the initial medium conditions (pH > 7), which are not favourable to optimal growth of these acidophilic bacteria. Moreover, the stabilisation of the growth of A. thiooxidans observed after 20 days may also be the consequence of the medium neutrality. Indeed, optimal growth of A. thiooxidans has been reported to occur between pH 2.5 and 4.0 (Blais et al., 1992). However, it can be noticed in Fig. 1 that the release of sulphate and calcium starts immediately after the beginning of the resuspension. Hence, it is probable that another sulphuroxidising bacterial population developing preferentially in neutral conditions is promoted, or that chemical oxidation occurs during this period of time. However, when comparing the curves obtained in sterile and nonsterile conditions (Fig. 1), it appears that the chemical oxidation alone cannot explain the oxidation rate observed during the 7 first days. This rather implies that other microbial species are responsible for the early oxidation of the sediment. As a matter of fact, Lascourreges-Berde€ u (1996) has shown that the bioleaching of metals contained in lake sediments seems to result from the effect of different sulphur-oxidising bacterial groups, depending upon optimal pH. Moreover, Blais et al. (1993) and Chen and Lin (2001) have observed that Thiobacillus thioparus is able to actually oxidise sulphide minerals in conjunction with A. thiooxidans. These authors have shown that T. thioparus is mainly responsible for the oxidation at neutral pH conditions and that its activity drops when the medium turns acidic. In contrast, the activity of A. thiooxidans is reduced in neutral conditions and increases when the pH of the medium drops. As a consequence, it is likely that in our experiments, the bacterial oxidation of the sediment is due to the concomitant activity of T. thioparus and A. thiooxidans, T. thioparus being responsible for the oxidation occurring during the first days of the resuspension assay.
4. Batch experiments in acidic medium 4.1. Metal mobilisation in leaching experiments In biotic conditions, the use of an acidic leaching solutions results in an initial decrease of the solution pH (Fig. 4). However, this acidification is rapidly buffered by calcite dissolution, since the pH reaches a stable value
625
of 6 after only 14 days. As a consequence of the enhanced dissolution of CaCO3 , the release of calcium in solution is shifted to higher values, compared to the leaching experiments conducted with a neutral solution (Fig. 1). Besides, the acidification also causes a dissolution of a higher amount of iron (Fig. 5). As a matter of fact, Fe reaches 30 mg/l after one day of re-suspension and decreases rapidly to be undetectable after 14 days. This can be attributed to the dissolution of iron-rich solid phases such as iron (oxi)hydroxides (ferrihydrite, haematite, goethite) in the early stages of the experiment, and to the re-precipitation of these phases when the pH reaches neutral values again. Concerning sulphate, one has to note that the release of SO2 4 is slightly less than the release observed in neutral conditions (Fig. 1). However, due to the uncertainties associated to the measured values, it is difficult to determine if this discrepancy is really significant. As a consequence, it seems that the initial acidification has little if no impact on the oxidation patterns of the sulphide mineral present in the sediment. However, compared to the experiments conducted in neutral conditions, Fig. 5 shows that the initial acidification causes a much higher release of Zn and Cd in solution. As a matter of fact, the release of both elements is 10-fold increased. Since the sulphate release curves have shown that the oxidation of the sediment does not seem to be enhanced by the initial acidification, the metal release is mainly due to the dissolution of other bearing phases like iron (oxi)hydroxides, as stated above. As a consequence, a small quantity of lead is also observed in solution during the first days of the experiment (Fig. 4). Due to its high affinity toward iron solid phases, the dissolved Pb vanishes when iron phases start to re-precipitate, due to adsorption and/or coprecipitation reactions. These results are strengthened by works of Lee and Touray (1997), who showed that a decrease in pH from 7 to 4 induced 5–10 times higher solubilisation rates for Zn and Cd than for Pb. Concerning the release of copper, no significant influence is observed after the initial acidification of the system (Fig. 5). This can be linked to the fact that copper was observed to be associated with the organic matter contained in the sediment. In abiotic conditions, one can note that the initial pH decrease is much lower than in biotic conditions, and is rather similar to the one observed in neutral and abiotic conditions (Fig. 1). In the same way, the release of Zn, Cd, Cu, Ca2þ and SO2 4 are nearly identical to the one observed in neutral abiotic conditions (Figs. 1 and 2). This induces that, in the absence of microorganisms, the initial acidification of the medium has no noticeable influence on the intrinsic chemistry of the system. As a consequence, this demonstrates that the initial pH drop, and the resulting higher metal release observed in biotic conditions are mainly due to the microbial activity in the
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250
Biotic condition
Biotic condition
Abiotic condition
8
Abiotic condition
[SO4 ] (mg/l)
200
2-
pH
6
4 2
150
100
50 0
0 0
20
40
60
80
0
100
20
Time (days) 50
60
80
100
80
100
600 Biotic condition
Biotic condition
Abiotic condition
40
500
30
[Ca] (mg/l)
[DOC] (mg/l)
40
Time (days)
20
10
Abiotic condition
400 300 200 100
0
0 0
20
40
60
80
100
Time (days)
0
20
40
60
Time (days)
Fig. 4. Evolution of pH and sulphate, organic carbon, calcium concentrations in leachates for experiments in acidic biotic and abiotic conditions.
sediment. Since slightly acidic conditions are observed in the early stages of the experiment, it is unlikely that neutrophilic species such as T. thioparus may play a role in these processes, since their range of pH growth is located between 6 and 8 (Lascourreges-Berde€ u, 1996). Besides, since the sulphate production does not seem to proceed to a greater extent either in biotic neutral or biotic acidic conditions, it appears that the microbial acidification is due to other biochemical pathways than the sole sulphide oxidation. Indeed, the oxidation of ferrous to ferric iron is frequently reported to cause a noticeable acidification of the medium, and ironoxidating microorganisms such as A. ferrooxidans or L. ferrooxidans are frequently used for bioleaching assays of metals contained in sludge or soils (Tyagi et al., 1993; Zagury et al., 1994; Solisio et al., 2002). Another possibility is that the initial pH drop observed in biotic and acidic conditions is merely due to the conjunction of the initial acid input and of the additional acid production originating from the microbial oxidation of sulphide and ferrous compounds.
As observed in neutral medium, copper is largely released in solution and very well correlated to the DOC evolution (r2 ¼ 0:98). Moreover, the release of copper and organic carbon are quantitatively identical to the one observed in neutral medium (Figs. 1 and 2). The fact that the release of DOC appears to be independent of the pH supports the hypothesis that this release is probably due to a physical degradation of the organic matter caused by the gamma-ray sterilisation (radiolysis of organic compounds), rather than a chemical desorption. 4.2. Evolution of bacterial population during the bioleaching experiment Microbiological analyses show that during the first days of the experiment, the total bacterial microflora is lower than in neutral conditions (around 7 · 105 bacteria/g dry sediment in acidic conditions, compared to 7 · 108 bacteria/g in neutral conditions) (Figs. 3 and 6). This lower population can be explained by the acidifi-
C. Lors et al. / Chemosphere 56 (2004) 619–630 50
1.0 Biotic condition
Abiotic condition
0.8
[Cd] (mg/l)
[Zn] (mg/l)
Biotic condition
Abiotic condition
40
30
20
10
0.6
0.4
0.2
0
0.0 0
20
40
60
80
100
0
20
Time (days)
40
60
80
100
80
100
Time (days) 0.10
0.10 Biotic condition
Biotic condition
Abiotic condition
Abiotic condition
0.08
[Cu] (mg/l)
0.08
[Pb] (mg/l)
627
0.06
0.04
0.02
0.06
0.04
0.02
0.00
0.00 0
20
40
60
80
100
Time (days)
0
20
40
60
Time (days)
40 Biotic condition Abiotic condition
[Fe] (mg/l)
30
20
10
0 0
20
40
60
80
100
Time (days) Fig. 5. Evolution of zinc, cadmium, lead, copper and iron concentrations in leachates for experiments in acidic biotic and abiotic conditions.
cation of the medium, which exerts a selection pressure on non-specific microorganisms. As a matter of fact, when pH rises to about 6, the global bacterial population increases and reaches the stable value 5 · 109 bac-
teria/g dry sediment, as observed for the experiments conducted at neutral conditions (Fig. 3). As observed when no initial acidification was performed, the sulphate-reducing bacteria population
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C. Lors et al. / Chemosphere 56 (2004) 619–630 1.E+12
Total bacterial microflora A. thiooxidans Sulfate-reducing microflora
Bacteria / g dry sediment
1.E+10
1.E+08
1.E+06
1.E+04
1.E+02
1.E+00 0
20
40
60
80
100
Time (days) Fig. 6. Evolution of total bacterial and specific (sulphur-oxidising and sulphate-reducing) populations during the acidic leaching experiment.
continuously decreases, from 104 down to 5 · 101 bacteria/g of dry sediment, due to the establishment of aerobic conditions, which are unfavourable for these strict anaerobic bacteria. For sulphur-oxidising bacteria, A. ferrooxidans are not detected during the experiment. As a consequence, the role of these bacteria in the initial pH drop of the system (see above) cannot be clearly highlighted. As suggested above, the pH drop observed in biotic and acidic conditions seems to be rather due to the conjunction of the initial acid input and of the additional acid produced by the sulphur-oxidising bacteria. In contrast, the increase of A. thiooxidans appears to be much faster than in neutral conditions, and no latency phase is observed (Fig. 6). Hence, the initial acidification of the medium allows an optimal growth of these acidophilic bacteria in the early stages of the experiment. However, one can note that as the system is again progressively buffered, the same plateau value is nearly attained for the A. thiooxidans population compared to the experiment performed without acidification. As in the same time, the sulphate are steadily released in solution, the stabilisation of the A. thiooxidans population is apparently not due to a substrate limitation. This is rather a consequence of the neutral pH of the system which is not favourable for the growth of these bacteria.
5. Conclusion The experiments carried out in this study have evidenced the major role played by indigenous microorganisms in the mobilisation of metals from aerated
dredged sediments. For experiment conducted in neutral medium, the results showed that the release of Zn and Cd was mainly due to the bacterial activation of sulphidic bearing phases. The implication of indigenous acidophilic sulphur-oxidising bacteria such as A. thiooxidans in this process was clearly highlighted. However, it likely that other species such as T. thioparus might have also played a role, at least in the early days stages of the oxidation. The initial acidification to pH ¼ 4 of a sterilised sediment suspension has no significant influence on the release of Zn, Pb and Cd, due to the buffering effect of the calcite contained in the sediment. For non-sterilised samples, this acidification resulted in an initial pH drop of the solution, which was rapidly buffered by the calcite dissolution. It appeared that this initial pH drop was due to the indigenous microbial activity, although no increase of the oxidation kinetic of sulphide was observed. Nevertheless, this initial acidification was sufficient to increase the release kinetics of Zn and Cd by a factor 10. This implies that care has to be taken to the conditions used for re-suspension experiments and their subsequent use in risk assessment, even if the sediment studied exhibits a high buffering capacity. The release kinetic of copper was greatly influenced by the sterilisation process, since the gamma-ray treatment resulted in a large release of DOC and copper, these two concentrations being linearly correlated. Although some uncertainties remain about the origin of this DOC release (likely due to radiolysis of organic compounds), this highlighted the strong association of copper with the organic matter present in dredged sediments.
C. Lors et al. / Chemosphere 56 (2004) 619–630
References American Petroleum Institute, 1965. Recommended Practice for Biological Analysis of Subsurface Injection Waters, second ed., vol. 38. Blais, J.F., Auclair, J.C., Tyagi, R.D., 1992. Cooperation between two Thiobacillus strains for heavy metal removal from municipal sludge. Can. J. Microbiol. 38, 181–187. Blais, J.F., Tyagi, R.D., Auclair, J.C., 1993. Bioleaching of metals from sewage sludge: microorganisms and growth kinetics. Water Res. 27 (1), 101–110. Bogusz, D., 1997. Traitement des sediments et des boues toxiques: etat des etudes et des travaux menes en France, oral communication. Environnement Nord-Pas de Calais, Quebec, Seminar, Lille. Bosecker, K., 1997. Bioleaching: metal solubilization by microorganisms. FEMS Microbiol. Rev. 20, 591–604. Caille, N., 2002. Mobilite et phytodisponibilite du mercure dans des dep^ ots de sediments de curage. These de doctorat specialite ‘‘Sciences agronomiques’’, Institut National Polytechnique de Lorraine. Caille, N., Tiffreau, C., Leyval, C., Morel, J.L., 2003. Solubility of metals in an anoxic sediment during prolonged aeration. Sci. Total Environ. 301, 239–250. Calmano, W., Ahlf, W., F€ orstner, U., 1996. Sediment Quality Assessment: Chemical and Biological Approaches. In: Calmano, W., F€ orstner, U. (Eds.), Sediments and Toxic Substances: Environmental Effects and Ecotoxicology. Springer-Verlag, Berlin, pp. 2–35. Calmano, W., Hong, J., F€ orstner, U., 1993. Binding and mobilisation of heavy metals in contaminated sediments by pH and redox potential. Water Sci. Technol. 28, 223–235. Cauwenberg, P., Maes, A., 1997. Influence of oxidation on sequential chemical extraction of dredged river sludge. Int. J. Environ. Anal. Chem. 68, 47–57. Chen, S.-Y., Lin, J.-G., 2001. Bioleaching of heavy metals from sediment: significance of pH. Chemosphere 44, 1093–1102. De Man, J.C., 1977. MPN tables for more than one test. Eur. J. Appl. Microbiol. 4, 307–316. Dood, J., Large, D.J., Fortey, N.J., Milodowski, A.E., Kemp, S.A., 2000. A petrographic investigation of two sequential extraction techniques applied to anaerobic canal bed mud. Environ. Geochem. Health 22, 281–296. F€ orstner, U., 1995. Non linear Release of Metals from Aquatic Sediments. In: Salomons, W., Stigliani, A. (Eds.), Biogeodynamics of Pollutants in Soils and Sediments. SpringerVerlag, Berlin, pp. 247–308. F€ orstner, U., Kersten, M., 1989. Assessment of Metal Mobility in Dredged Material and Mine Waste by Pore Water Chemistry and Solid Speciation. In: Salomons, W., F€ orstner, U. (Eds.), Chemistry and Biology of Solid Waste: Dredged Material and Mine Tailings. Springer-Verlag, Berlin, pp. 171–214. Giani, M., Gabellini, M., Pellegrini, D., Costantini, S., Beccaloni, E., Giordano, R., 1994. Concentration and partition of Hg, Cr and Pb in sediments of dredge and disposal sites of the northern Adriatic Sea. Sci. Total Environ. 158, 97–112. Isaure, M.P., Laboudigue, A., Manceau, A., Sarret, G., Tiffreau, C., Trocellier, P., Lamble, G., Hazemann, J.L., Chateigner, D., 2002. Quantitative Zn speciation in a contaminated dredged sediment by l-PIXE, l-SXRF,
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Tack, F.M., Singh, S.P., Verloo, M.G., 1999. Leaching behaviour of Cd, Cu, Pb and Zn in surface soils derived from dredged sediments. Environ. Pollut. 106, 107–114. Tichy, R., Rulkens, W.H., Grotenhuis, J.T.C., Nydl, V., Cuypers, C., Fajtl, J., 1998. Bioleaching of metals from soils or sediments. Water Sci. Technol. 37 (8), 119–127. Tiffreau, C., Isaure, M.P., Laboudigue, A., Lors, C., Marseille, F., 1999. Problematique des sediments toxiques: impact sur un sol non pollue du dep^ ot de sediments contamines. CNRSSP report 99/35.
Tyagi, R.D., Blais, J.F., Auclair, J.C., 1993. Bacterial leaching of metals from sewage sludge by indigenous iron-oxidizing bacteria. Environ. Pollut. 82, 9–12. Yu, K.C., Tsai, L.J., Chen, S.H., Ho, S.T., 2001. Chemical binding of heavy metals in anoxic river sediments. Water Res. 35, 4086–4094. Zagury, G.J., Narasiah, K.S., Tyagi, R.D., 1994. Adaptation of indigenous iron-oxidizing bacteria for bioleaching of heavy metals in contaminated soils. Environ. Technol. 15, 517– 530.
Effects of bacterial activities on the release of heavy metals from contaminated dredged sediments C. Lors *, C. Tiffreau 1, A. Laboudigue CNRSSP, BP 537, 930 Boulevard Lahure, 59505 Douai Cedex, France Received 29 April 2003; received in revised form 27 February 2004; accepted 20 April 2004
Abstract The potential impact of indigenous bacterial processes on the release of heavy metals from dredged sediment deposits was investigated. Batch re-suspension experiments were conducted in order to investigate the release of Zn, Cd, Cu and Pb from a polluted anoxic sediment submitted to oxidative perturbations. The concentrations of heavy metals, sulphate and dissolved organic carbon (DOC) were periodically recorded, and cell counts were performed to follow the evolution of several bacterial species. The specific effects of microbial processes were quantified by performing resuspension assays on sterilised samples. Moreover, the effect of an initial acidification of the system was studied. The results showed that metal release was mainly due to oxidation of sulphide minerals contained in the sediment. Sulphur-oxidising bacteria such as Acidithiobacillus thiooxidans were identified to play a major role in the process, by enhancing the oxidation kinetic. However, the acid production resulting from these reactions was almost totally buffered by the dissolution of the calcite present in the sediment. Copper was released to a lesser extent, and a strong association with organic matter was observed. Lead was not observed in solution, because of its low solubility at neutral conditions and of its re-adsorption on the solid phase. The initial acidification of the system resulted in an faster growth of the acidophilic A. thiooxidans. A subsequent pH drop originating from microbial processes was then observed during the first stages of the experiment. As a consequence, drastic increases in metal (Zn, Cd) release were observed. 2004 Elsevier Ltd. All rights reserved. Keywords: Dredged sediments; Heavy metal release; Acidithiobacillus; Thiobacillus; Re-suspension; Bacterial activity
1. Introduction Due to erosion, urban and industrial run-offs, waterways are subject to a constant silting up. As a consequence, it is necessary to undertake regular dredging operations, in order to keep the waterways in good conditions. These operations generate about
*
Corresponding author. Tel.: +33-3-2771-2681; fax: +33-32771-0707. E-mail address: [email protected] (C. Lors). 1 Present address: CEA Cadarache, DED/SAMRA/LTCR, 13108 St. Paul lez Durance.
250 000 m3 of sediments each year for the French NordPas de Calais region (Bogusz, 1997). However, because of their high sorption capacity, sediments act as a sink for toxic substances (heavy metals, organic pollutants) (F€ orstner and Kersten, 1989; Calmano et al., 1996). Consequently, dredged materials are frequently contaminated above safety limits and can represent a secondary source of pollution. As a matter of fact, sediments are generally deposited along streams or in deposit sites because of the cost of other treatment options. However, disposal of dredged material can modify the initial physico-chemical and microbiological conditions prevailing inside the sediments, simply because initially anoxic sediments become progressively
0045-6535/$ - see front matter 2004 Elsevier Ltd. All rights reserved. doi:10.1016/j.chemosphere.2004.04.009
Quartz (SiO2 ), Calcite (CaCO3 ), Albite (NaAlSi3 O8 ), Microcline (KAlSi3 O8 ), Kaolinite, Illite Mineralogical major phases (XRD)
273 5
Mn
2. Materials and methods
50 000 4000
P
3600 200
Fe
18 000 1000
TOC
6.5 0.5
%
53 6
Humidity
pH
oxygenated once they are deposited. This induces major changes in the oxydo-reduction potential, the pH and the speciation of major and trace elements (Tack et al., 1996, 1999). This also promotes the growth of aerobic bacteria playing a role in metal mobilisation, particularly those of genus Acidithiobacillus (Bosecker, 1997). These bacteria are responsible for the oxidation of insoluble metal sulphides to soluble metal sulphates with a high acidification of sediments (Blais et al., 1993; Tichy et al., 1998; Chen and Lin, 2001). As a consequence, this can lead to modifications of biosorption and bioleaching equilibria and to enhanced leaching of the metallic pollutants from the deposit sites (Petersen et al., 1997). The aim of this study is to evaluate the influence of the bacterial activity on the solubilisation of heavy metals contained in dredged sediments. More precisely, the objectives are to investigate the occurrence of indigenous sulphur-oxidising bacteria in sediments and to assess the potential of this microflora to leach toxic metals from dredged materials. The impact of an acidification of the medium (acidic rainfall, acidic run-off, etc.) on the metal solubilisation was also studied.
7.2 0.1
C. Lors et al. / Chemosphere 56 (2004) 619–630
Concentration
620
150 10 190 10 770 15
Cu Cd
6600 300
Pb Zn Ca Unit
Table 1 Heavy metals concentrations and main characteristics of the dredged sediment
The sediment has been sampled using a mechanical shovel (between 0 and 50 cm depth) from the Scarpe channel near the city of Raches (North of France). Once dredged, samples were immediately placed in plastic buckets, and stored at 4 C. The samples were bottled with their initial moisture level (i.e. no drying or draining was performed) and kept saturated in their porewater in order to maintain anoxic conditions during storage. Prior to experiments, the sediments were manually homogenised and wet sieved through a 2 mm mesh. Heavy metal concentrations and the main characteristics of the sediment are reported in Table 1. The Ca, Zn, Pb, Cd, Cu, Mn, P and Fe contents were determined according to the PR NF ISO 14869-1 standard: 0.5 g of dried sediment was heated to 450 C during 4 h and then allowed to cool down. The calcinated solid was then moistened with 4 ml of deionized water and transferred to a Teflon PTFE beaker. 10 ml of hydrofluoric (HF) and 3 ml of perchloric (HClO4 ) acids were added to the sediment, and the mixture was let to settle for 12 h. The beaker was then heated at 150 C during 2 h and evaporated to near dryness. Digests were diluted after filtration (0.45 lm) to 100 ml with a 2% HCl solution. Ca, Zn, Pb, Cd, Cu, Mn, P and Fe in the solution were then determined using Inductively coupled Plasma-Atomic Emission Spectrometry (ICP-AES) (Jobin Yvon, France, JY 138 Ultrace). The total organic carbon (TOC) was determined with a TOC
mg/kg dry sediment
2.1. Sediment sampling and characterisation
C. Lors et al. / Chemosphere 56 (2004) 619–630
analyser (Shimatzu, TOC 5000A with SSM 5000A module) according to the NF ISO 10694 standard. The TOC was obtained by subtracting inorganic carbon (IC) from total carbon (TC). The moisture content was determined by gravimetry, according to the NF ISO 11465 standard. The initial pH of the sediment was obtained following the NF ISO 10390 protocol. However, special care was taken in order to avoid oxidation during the measurement: a 1:5 solid/liquid suspension was prepared with fresh sediment and deaerated deionized water inside a N2 saturated glove box. The pH was measured inside the glove box after a 6 h stirring period, using a combined electrode (Ingold). 2.2. Batch experiments Leaching experiments were conducted both on fresh anoxic sediments, and on sterilised sediments in order to determine the part of abiotic processes in metal mobilisation. For sterilisation, some weighted amounts of sediments were placed in closed bottles and sterilised by gamma-ray irradiation (42 kGy). The slurry for the leaching tests were prepared by mixing fresh anoxic or sterilised sediment with a sterile leaching solution ([CaCl2 ] ¼ 0.1 mM; [NaNO3 ] ¼ 0.2 mM) at a ratio of 2 g wet sediment to 100 ml leaching solution. This solution was used in order to (i) avoid the potential impact of distillated water on bacterial cells due to osmotic factors, (ii) simulate the background concentration of the local rain water. For both fresh and sterilised sediments, two leaching solutions were used: one with neutral pH conditions and the other one with an initial pH previously adjusted to 4 with 0.1 N nitric acid. For sterilised sediments, this preparation was carried out under a sterile laminar hood. Batch experiments were carried out in 250 ml glass bottles and continuously agitated on a horizontal shaker at 20 C. The bottles were capped with porous foam rubbers in order to ensure gas exchange between the bottle contents and the atmosphere and to prevent microbiological contamination. For each of the four conditions studied (sterilised, non-sterilised, neutral solution, acidified solution), six samples were prepared, each sample being prepared in triplicate (i.e. 4 · 6 · 3 ¼ 72 total samples). The release of metals in solution was monitored as a function of time (respectively at day 1, 7, 14, 27, 62, 98 for biotic conditions and at day 1, 7, 16, 23, 28, 76 for abiotic conditions). At each date, the triplicate microcosms of each of the four conditions were sacrificed for
621
analysis. For each microcosm, a 1 ml suspension sample was withdrawn under agitation for counting of total, sulphur-oxidising and sulphate-reducing bacterial populations. The rest of the microcosm was centrifuged at 13 000 rpm for 30 min and the supernatant was sampled for physico-chemical analyses (pH, dissolved organic carbon (DOC), SO2 4 , metals). The pH of the leachate was measured using a pHmeter (Consort C833) equipped with a temperature correction sensor and a combined Ingold electrode. The total Zn, Cd, Cu, Fe, Pb, Mn, Al, Ca, DOC and sulphate contents in the supernatant were determined after a 0.45 lm additional filtration. Metal concentrations were measured by ICP-AES (Jobin Yvon 138 Ultrace), the DOC was determined using a Shimatzu TOC-5000A using catalytic oxidation with an infrared CO2 quantifier in accordance to the NF T 90-102 standard. Sulphate ions were quantified by a colorimetric method according to the NF T 90-040 reference protocol. 2.3. Microbiological analyses Enumeration of total bacterial and sulphur-oxidising bacterial populations were carried out on microplates of 96 wells from Nunc (Nunclon Delta, Germany), on which each well contained 250 ll of specific medium. The Nutrient broth (NB, Difco, USA) medium (diluted to 1/10) was used for total bacterial count. The medium of Razzel and Trussel (1963) in addition with a solution of FeSO4 Æ 7H2 O (c ¼ 20 g/l) was used for Acidithiobacillus ferrooxidans enumeration. For acidophilic sulphur-oxidising bacteria, the modified medium 317 (Chen and Lin, 2001) was combined with a color indicator (bromocresol green, c ¼ 0:02 g/l). The complete compositions of these two media are described in Table 2. The pH of acidophilic sulphur-oxidising bacteria and A. ferrooxidans medium was adjusted respectively to 4.5 and 1.8 with H2 SO4 (1 N). Each well was inoculated with 25 ll of suspension sample diluted to 100 –108 with Ringer solution (Ramsay, 1984). The inoculated microplates were incubated at 30 C during 15 days and then evaluated for colored products. Enumeration of bacteria was carried out using the Most Probable Numbers (MPN) method (De Man, 1977). Sulphate-reducing bacteria were counted in ready-touse culture tubes (Fisher Scientific, ref. A80 29406) containing a solid medium specific to sulphate-reducing bacteria (American Petroleum Institute, 1965). The
Table 2 Composition of medium (g/l of deionized water) for acidophilic sulphur-oxidising (ASO) bacteria and A. ferrooxidans growth ASO bacteria A. ferrooxidans
KH2 PO4
K2 HPO4
(NH4 )2 SO4
MgSO4 Æ 7H2 O
FeSO4 Æ 7H2 O
CaCl2
Na2 S2 O3
– 0.4
3.5 –
0.3 0.8
0.5 0.2
0.018 20
0.25 –
10 –
622
C. Lors et al. / Chemosphere 56 (2004) 619–630
tubes were previously heated to 100 C to liquefy the medium and then cooled down to 55 C for inoculation. Each of them was inoculated with 1 ml of sample diluted to 100 –102 . Three repetitions were performed. After incubation at 30 C during 15 days, the number of sulphate-reducing bacteria of the tested sample was determined from the black coloration of the medium using the MPN method (De Man, 1977).
considered that metals are associated with sulphidic and organic compounds. As a consequence, oxidation of dredged material is generally considered to cause dramatic modifications of the heavy metal mobility (Calmano et al., 1993; F€ orstner, 1995; Tack et al., 1996; Stephens et al., 2001). More recently, the use of spectroscopic methods has directly evidenced the occurrence of heavy metal sulphides in anoxic river sediments (Dood et al., 2000; Large et al., 2001; Isaure et al., 2002). More specifically, FeS2 (pyrite), ZnS (sphalerite) and PbS (galena) have been clearly identified in sediments originating from the same sampling spot as the one chosen for this study (Isaure et al., 2002). As shown in Fig. 1, aerating the sediment in biotic conditions leads to a marked release of sulphate in solution, which is probably due to the oxidation of sulphide solid phases. Although the oxidation of sulphidic minerals is generally accompanied by a large proton release (F€ orstner, 1995), the pH decrease observed in our case is quite low (from about 8 to 7). As a matter of fact, the acidification is buffered by the dis-
3. Results and discussion 3.1. Batch experiments in neutral medium 3.1.1. Metal mobilisation in leaching experiments Different studies based on chemical extraction schemes have shown that heavy metals present in anoxic sediments are generally associated with phases depicted as ‘‘residual’’ or ‘‘oxidable’’ (Kersten and F€ orstner, 1986; Giani et al., 1994; Cauwenberg and Maes, 1997; Yu et al., 2001). From these observations, it is generally
300
10
Biotic condition
Biotic condition Abiotic condition
Abiotic condition
250
[SO4 ] (mg/l)
8
200 150
2-
pH
9
100
7 50 0
6 0
20
40
60
80
0
100
20
Time (days) 50
60
80
100
80
100
200 Biotic condition
Biotic condition
Abiotic condition
40
Abiotic condition
150
[Ca] (mg/l)
[DOC] (mg/l)
40
Time (days)
30
20
100
50 10
0
0 0
20
40
60
Time (days)
80
100
0
20
40
60
Time (days)
Fig. 1. Evolution of pH and sulphate, organic carbon, calcium concentrations in leachates for experiments in neutral biotic and abiotic conditions.
C. Lors et al. / Chemosphere 56 (2004) 619–630
solution of calcite, which is present in large amounts in our sediment. Consequently to this dissolution, calcium is released in solution with a kinetic comparable to that of sulphate ions. Beside, it has been observed the monitoring of an experimental deposit, that the release of sulphate and calcium in solution can lead to saturation with respect to gypsum (CaSO4 ), and thus, to a limitation of the Ca and SO2 concentrations in solution 4 (Tiffreau et al., 1999). However, the concentrations observed in the present study are far from the saturation equilibrium, which is typically observed for concentration values of 1.5 g/l SO2 and 0.5 g/l Ca. In biotic 4 conditions, the oxidation of the metal-bearing sulphidic phases results in a release of high amounts of zinc in solution during the first 62 days, reaching more than 4 mg/l (Fig. 2). Cadmium follows the same evolution as zinc in proportional amount. In contrast, the proportion of released copper appears to be much lesser than for Zn and Cd. This is consistent with the observations made by Stephens et al. (2001). This lower mobility is attributed to the higher affinity of Cu for organic matter and for
623
ferric (hydr)oxides (haematite, goethite, ferrihydrite) that can form during the oxidation of the system. In abiotic conditions, the release of sulphate and calcium is much lower than in the presence of microorganisms (Fig. 1). This confirms that in the absence of microbial activity, chemical oxidation of sulphide minerals proceeds to a much lesser extent than in biotic conditions (Petersen et al., 1997). Moreover, the concentrations of calcium and sulphate do not show any increase as a function of time. This seems to indicate that only a small fraction of the sulphide pool seems to be readily available for chemical oxidation, compared to the biochemical one. As a result, the release of Zn and Cd in solution is also much lower than in biotic conditions (Fig. 2), illustrating the major role of bacterial activity in the leaching of these metals. Concerning copper, a completely different behaviour can be observed in Fig. 2. In abiotic conditions, a higher release of Cu is observed, which exhibits a linear trend as a function of time. The same discrepancy between biotic and abiotic conditions can be noticed for the DOC
6
0.20 Biotic condition
Biotic condition
5
Abiotic condition
Abiotic condition
4
[Cd] (mg/l)
[Zn] (mg/l)
0.15
3 2
0.10
0.05 1 0.00
0 0
20
40
60
80
0
100
20
Time (days)
60
80
100
80
100
0.10
0.10 Biotic condition
Biotic condition
Abiotic condition
Abiotic condition
0.08
[Pb] (mg/l)
0.08
[Cu] (mg/l)
40
Time (days)
0.06
0.04
0.06
0.04
0.02
0.02
0.00
0.00 0
20
40
60
Time (days)
80
100
0
20
40
60
Time (days)
Fig. 2. Evolution of zinc, cadmium, copper and lead concentrations in leachates for experiments in neutral biotic and abiotic conditions.
624
C. Lors et al. / Chemosphere 56 (2004) 619–630
adsorption is supposed to proceed via pure adsorption or coprecipitation with (hydr)oxides of Fe, Al or Mn which form at neutral pH conditions. In our case, this is supported by the fact that neither Fe, nor Mn, nor Al were detected in solution, indicating that the solution was saturated with respect to the (hydr)oxides of these elements.
evolution (Fig. 1). Moreover, when the Cu concentration in solution is plotted as a function of the DOC concentration, both parameters appear to be strongly linearly correlated (R2 ¼ 0:97). This suggests that Cu is associated with some organic compounds, which are released in solution in sterile conditions. Although the association of copper with organic matter present in dredged sediment is frequently reported (Kersten and F€ orstner, 1986; Calmano et al., 1993), the reason for the DOC release in solution remains obscure. Luo et al. (2001) have reported a similar phenomenon (i.e. release of DOC and copper in solution) for a 10 kGy gammaray sterilisation performed on a sandy loam. They suggested that the DOC release was due to the decomposition of killed organisms by surviving species, or that the sterilisation could render some parts of the nonmicrobial soil organic matter more decomposable. In our case, it is to note that such a release has not been observed in another series of batch experiments conducted on sediments from the Scarpe River sterilised with a lower dose (35 kGy) (Caille, 2002). Moreover, since no bacterial activity was detected in sterile samples during the whole experiments (see next paragraph), it is more likely that gamma-ray sterilisation performed for this study might have caused a radio-chemical degradation of the organic matter contained in the sediment. Compared to Zn, Cd and Cu, Pb was not detected during the experiment in both sterile and non-sterile conditions (Fig. 2). These results support the observations made by F€ orstner (1995) and Caille et al. (2003), who have reported that lead is released only during the very early hours of the re-suspension experiments and is then quickly re-adsorbed on the solid phases. This re-
1.E+12
3.2. Evolution of bacterial populations during the bioleaching experiment The results of total bacterial microflora and specific microflora countings (sulphur-oxidising and sulphatereducing bacteria) are reported in Fig. 3. The results concerning sterile microcosms are not shown, since they indicated that no microbial activity was detected during the whole experiments. This allows to ensure that sterile conditions were maintained during the re-suspension experiment. In non-sterile experiments, the total bacterial microflora is relatively abundant (7 · 108 bacteria/g of dry sediment) and stable during the experiment. Obviously, this indicates that the sediment possesses nutritive qualities for optimal bacterial activity. During the re-suspension experiment, the sulphatereducing bacterial population decreases progressively over time, down from 104 to 101 bacteria/g dry sediment. This can be attributed to the establishment of aerobic conditions, which are unfavourable to the growth of these strict anaerobic bacteria. For sulphur-oxidising bacteria, A. ferrooxidans is not detected during the experiment whereas an increase in acidophilic sulphur-oxidising bacteria is observed during
Total bacterial microflora A. thiooxidans Sulfate-reducing microflora
Bacteria / g dry sediment
1.E+10
1.E+08
1.E+06
1.E+04
1.E+02
1.E+00 0
20
40
60
80
100
Time (days) Fig. 3. Evolution of total bacterial and specific (sulphur-oxidising and sulphate-reducing) populations during the neutral leaching experiment.
C. Lors et al. / Chemosphere 56 (2004) 619–630
time, after a latency time of 7 days. Based on the study of 21 different sewage sludges, Blais et al. (1993) have shown that among all the acidophilic sulphur-oxidising bacteria, A. thiooxidans is the only species involved in the leaching of heavy metals. As a consequence, it can be considered that the growth observed in our study is due to the development of A. thiooxidans. The latency period seems to be an adaptation period of these microorganisms to the initial medium conditions (pH > 7), which are not favourable to optimal growth of these acidophilic bacteria. Moreover, the stabilisation of the growth of A. thiooxidans observed after 20 days may also be the consequence of the medium neutrality. Indeed, optimal growth of A. thiooxidans has been reported to occur between pH 2.5 and 4.0 (Blais et al., 1992). However, it can be noticed in Fig. 1 that the release of sulphate and calcium starts immediately after the beginning of the resuspension. Hence, it is probable that another sulphuroxidising bacterial population developing preferentially in neutral conditions is promoted, or that chemical oxidation occurs during this period of time. However, when comparing the curves obtained in sterile and nonsterile conditions (Fig. 1), it appears that the chemical oxidation alone cannot explain the oxidation rate observed during the 7 first days. This rather implies that other microbial species are responsible for the early oxidation of the sediment. As a matter of fact, Lascourreges-Berde€ u (1996) has shown that the bioleaching of metals contained in lake sediments seems to result from the effect of different sulphur-oxidising bacterial groups, depending upon optimal pH. Moreover, Blais et al. (1993) and Chen and Lin (2001) have observed that Thiobacillus thioparus is able to actually oxidise sulphide minerals in conjunction with A. thiooxidans. These authors have shown that T. thioparus is mainly responsible for the oxidation at neutral pH conditions and that its activity drops when the medium turns acidic. In contrast, the activity of A. thiooxidans is reduced in neutral conditions and increases when the pH of the medium drops. As a consequence, it is likely that in our experiments, the bacterial oxidation of the sediment is due to the concomitant activity of T. thioparus and A. thiooxidans, T. thioparus being responsible for the oxidation occurring during the first days of the resuspension assay.
4. Batch experiments in acidic medium 4.1. Metal mobilisation in leaching experiments In biotic conditions, the use of an acidic leaching solutions results in an initial decrease of the solution pH (Fig. 4). However, this acidification is rapidly buffered by calcite dissolution, since the pH reaches a stable value
625
of 6 after only 14 days. As a consequence of the enhanced dissolution of CaCO3 , the release of calcium in solution is shifted to higher values, compared to the leaching experiments conducted with a neutral solution (Fig. 1). Besides, the acidification also causes a dissolution of a higher amount of iron (Fig. 5). As a matter of fact, Fe reaches 30 mg/l after one day of re-suspension and decreases rapidly to be undetectable after 14 days. This can be attributed to the dissolution of iron-rich solid phases such as iron (oxi)hydroxides (ferrihydrite, haematite, goethite) in the early stages of the experiment, and to the re-precipitation of these phases when the pH reaches neutral values again. Concerning sulphate, one has to note that the release of SO2 4 is slightly less than the release observed in neutral conditions (Fig. 1). However, due to the uncertainties associated to the measured values, it is difficult to determine if this discrepancy is really significant. As a consequence, it seems that the initial acidification has little if no impact on the oxidation patterns of the sulphide mineral present in the sediment. However, compared to the experiments conducted in neutral conditions, Fig. 5 shows that the initial acidification causes a much higher release of Zn and Cd in solution. As a matter of fact, the release of both elements is 10-fold increased. Since the sulphate release curves have shown that the oxidation of the sediment does not seem to be enhanced by the initial acidification, the metal release is mainly due to the dissolution of other bearing phases like iron (oxi)hydroxides, as stated above. As a consequence, a small quantity of lead is also observed in solution during the first days of the experiment (Fig. 4). Due to its high affinity toward iron solid phases, the dissolved Pb vanishes when iron phases start to re-precipitate, due to adsorption and/or coprecipitation reactions. These results are strengthened by works of Lee and Touray (1997), who showed that a decrease in pH from 7 to 4 induced 5–10 times higher solubilisation rates for Zn and Cd than for Pb. Concerning the release of copper, no significant influence is observed after the initial acidification of the system (Fig. 5). This can be linked to the fact that copper was observed to be associated with the organic matter contained in the sediment. In abiotic conditions, one can note that the initial pH decrease is much lower than in biotic conditions, and is rather similar to the one observed in neutral and abiotic conditions (Fig. 1). In the same way, the release of Zn, Cd, Cu, Ca2þ and SO2 4 are nearly identical to the one observed in neutral abiotic conditions (Figs. 1 and 2). This induces that, in the absence of microorganisms, the initial acidification of the medium has no noticeable influence on the intrinsic chemistry of the system. As a consequence, this demonstrates that the initial pH drop, and the resulting higher metal release observed in biotic conditions are mainly due to the microbial activity in the
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C. Lors et al. / Chemosphere 56 (2004) 619–630 10
250
Biotic condition
Biotic condition
Abiotic condition
8
Abiotic condition
[SO4 ] (mg/l)
200
2-
pH
6
4 2
150
100
50 0
0 0
20
40
60
80
0
100
20
Time (days) 50
60
80
100
80
100
600 Biotic condition
Biotic condition
Abiotic condition
40
500
30
[Ca] (mg/l)
[DOC] (mg/l)
40
Time (days)
20
10
Abiotic condition
400 300 200 100
0
0 0
20
40
60
80
100
Time (days)
0
20
40
60
Time (days)
Fig. 4. Evolution of pH and sulphate, organic carbon, calcium concentrations in leachates for experiments in acidic biotic and abiotic conditions.
sediment. Since slightly acidic conditions are observed in the early stages of the experiment, it is unlikely that neutrophilic species such as T. thioparus may play a role in these processes, since their range of pH growth is located between 6 and 8 (Lascourreges-Berde€ u, 1996). Besides, since the sulphate production does not seem to proceed to a greater extent either in biotic neutral or biotic acidic conditions, it appears that the microbial acidification is due to other biochemical pathways than the sole sulphide oxidation. Indeed, the oxidation of ferrous to ferric iron is frequently reported to cause a noticeable acidification of the medium, and ironoxidating microorganisms such as A. ferrooxidans or L. ferrooxidans are frequently used for bioleaching assays of metals contained in sludge or soils (Tyagi et al., 1993; Zagury et al., 1994; Solisio et al., 2002). Another possibility is that the initial pH drop observed in biotic and acidic conditions is merely due to the conjunction of the initial acid input and of the additional acid production originating from the microbial oxidation of sulphide and ferrous compounds.
As observed in neutral medium, copper is largely released in solution and very well correlated to the DOC evolution (r2 ¼ 0:98). Moreover, the release of copper and organic carbon are quantitatively identical to the one observed in neutral medium (Figs. 1 and 2). The fact that the release of DOC appears to be independent of the pH supports the hypothesis that this release is probably due to a physical degradation of the organic matter caused by the gamma-ray sterilisation (radiolysis of organic compounds), rather than a chemical desorption. 4.2. Evolution of bacterial population during the bioleaching experiment Microbiological analyses show that during the first days of the experiment, the total bacterial microflora is lower than in neutral conditions (around 7 · 105 bacteria/g dry sediment in acidic conditions, compared to 7 · 108 bacteria/g in neutral conditions) (Figs. 3 and 6). This lower population can be explained by the acidifi-
C. Lors et al. / Chemosphere 56 (2004) 619–630 50
1.0 Biotic condition
Abiotic condition
0.8
[Cd] (mg/l)
[Zn] (mg/l)
Biotic condition
Abiotic condition
40
30
20
10
0.6
0.4
0.2
0
0.0 0
20
40
60
80
100
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20
Time (days)
40
60
80
100
80
100
Time (days) 0.10
0.10 Biotic condition
Biotic condition
Abiotic condition
Abiotic condition
0.08
[Cu] (mg/l)
0.08
[Pb] (mg/l)
627
0.06
0.04
0.02
0.06
0.04
0.02
0.00
0.00 0
20
40
60
80
100
Time (days)
0
20
40
60
Time (days)
40 Biotic condition Abiotic condition
[Fe] (mg/l)
30
20
10
0 0
20
40
60
80
100
Time (days) Fig. 5. Evolution of zinc, cadmium, lead, copper and iron concentrations in leachates for experiments in acidic biotic and abiotic conditions.
cation of the medium, which exerts a selection pressure on non-specific microorganisms. As a matter of fact, when pH rises to about 6, the global bacterial population increases and reaches the stable value 5 · 109 bac-
teria/g dry sediment, as observed for the experiments conducted at neutral conditions (Fig. 3). As observed when no initial acidification was performed, the sulphate-reducing bacteria population
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C. Lors et al. / Chemosphere 56 (2004) 619–630 1.E+12
Total bacterial microflora A. thiooxidans Sulfate-reducing microflora
Bacteria / g dry sediment
1.E+10
1.E+08
1.E+06
1.E+04
1.E+02
1.E+00 0
20
40
60
80
100
Time (days) Fig. 6. Evolution of total bacterial and specific (sulphur-oxidising and sulphate-reducing) populations during the acidic leaching experiment.
continuously decreases, from 104 down to 5 · 101 bacteria/g of dry sediment, due to the establishment of aerobic conditions, which are unfavourable for these strict anaerobic bacteria. For sulphur-oxidising bacteria, A. ferrooxidans are not detected during the experiment. As a consequence, the role of these bacteria in the initial pH drop of the system (see above) cannot be clearly highlighted. As suggested above, the pH drop observed in biotic and acidic conditions seems to be rather due to the conjunction of the initial acid input and of the additional acid produced by the sulphur-oxidising bacteria. In contrast, the increase of A. thiooxidans appears to be much faster than in neutral conditions, and no latency phase is observed (Fig. 6). Hence, the initial acidification of the medium allows an optimal growth of these acidophilic bacteria in the early stages of the experiment. However, one can note that as the system is again progressively buffered, the same plateau value is nearly attained for the A. thiooxidans population compared to the experiment performed without acidification. As in the same time, the sulphate are steadily released in solution, the stabilisation of the A. thiooxidans population is apparently not due to a substrate limitation. This is rather a consequence of the neutral pH of the system which is not favourable for the growth of these bacteria.
5. Conclusion The experiments carried out in this study have evidenced the major role played by indigenous microorganisms in the mobilisation of metals from aerated
dredged sediments. For experiment conducted in neutral medium, the results showed that the release of Zn and Cd was mainly due to the bacterial activation of sulphidic bearing phases. The implication of indigenous acidophilic sulphur-oxidising bacteria such as A. thiooxidans in this process was clearly highlighted. However, it likely that other species such as T. thioparus might have also played a role, at least in the early days stages of the oxidation. The initial acidification to pH ¼ 4 of a sterilised sediment suspension has no significant influence on the release of Zn, Pb and Cd, due to the buffering effect of the calcite contained in the sediment. For non-sterilised samples, this acidification resulted in an initial pH drop of the solution, which was rapidly buffered by the calcite dissolution. It appeared that this initial pH drop was due to the indigenous microbial activity, although no increase of the oxidation kinetic of sulphide was observed. Nevertheless, this initial acidification was sufficient to increase the release kinetics of Zn and Cd by a factor 10. This implies that care has to be taken to the conditions used for re-suspension experiments and their subsequent use in risk assessment, even if the sediment studied exhibits a high buffering capacity. The release kinetic of copper was greatly influenced by the sterilisation process, since the gamma-ray treatment resulted in a large release of DOC and copper, these two concentrations being linearly correlated. Although some uncertainties remain about the origin of this DOC release (likely due to radiolysis of organic compounds), this highlighted the strong association of copper with the organic matter present in dredged sediments.
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