editors: gra 17. bioremediation of soils contaminated with...

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Trends in Bioremediation and Phytoremediation, 2010: 283-299 ISBN: 978-81-308-0424-8 Editors: Grażyna Płaza

17. Bioremediation of soils contaminated with metalliferous mining wastes

H. Cortez1, J. Pingarrón1, J.A. Muñoz1, A. Ballester1, M.L. Blázquez1 F. González1, C. García1 and O. Coto2

1Departamento de Ciencia de los Materiales e Ingeniería Metalúrgica, Facultad de Ciencias Químicas Universidad Complutense, Av. Complutense s/n, 28040 Madrid, Spain; 2Departamento de Microbiología

Facultad de Biología, Universidad de La Habana, Calle 25 # 462, 10400 La Habana, Cuba

Abstract. Historically, microorganisms have been used to remediate sites contaminated with organic compounds. Specifically, soils provide an incalculable number of ecological niches for the growth of microbial consortia which play a decisive role in that natural process. Despite that biodegradation of organic contaminants is a process relatively well documented and has been successfully implemented at contaminated sites, the technical basis for natural attenuation of heavy metals are still undefined. It is known, however, that microbial processes play a key role in mobilization or immobilization of metals and that, although heavy metals cannot be biodegraded, they can be bioreduced or biooxidized to less toxic and less mobile forms. Thus, the success of bioremediation of soils contaminated with heavy metals requires a comprehensive understanding of the factors that drive or restrict the desirable biotransformations. This work is focus on the environmental impact caused by metalliferous mine wastes containing toxic metals and their possibilities of bioremediation. Further information is also given on the effect of bioaugmentation on metal mobilization in soil microcosms using a contaminated soil from a mining site located in southeast Spain.

1. Introduction Environmental pollution is one of the biggest concerns of today’s society. The huge amount of organic and inorganic wastes generated annually in the materials cycle and the Correspondence/Reprint request: Dr. J.A. Muñoz, Departamento de Ciencia de los Materiales e Ingeniería Metalúrgica Facultad de Ciencias Químicas, Universidad Complutense, Av. Complutense s/n, 28040 Madrid, Spain E-mail: [email protected]

H. Cortez et al. 284

improper disposal practices used have ended in a worldwide problem. These contaminants pose serious ecological threats to natural ecosystems with fatal consequences for human health. Bioremediation of contaminated soils is a natural attenuation process that takes place almost everywhere in our planet and since life first appeared on the Earth about 3,500 million years ago. Indeed, the initial hypothesis of the bioremediation process is that behind the idea formulated by Lovelock in 1979 in his Gaia theory [1]. According to such theory, the Earth, in its outermost part, resembles a single organism that under the action of external factors, mainly human activities, activates mechanisms to restore environmental conditions suitable for life. Clearly, the problematic of soil contamination with heavy metals seems to follow this pattern. In such a way that natural ecosystems automatically activate biochemical processes led to minimize the harmful effects provoked. These natural processes include the action of the microbial population present in the soil, the sorption of contaminants by specific soil particles or its agglomeration into nuclei. The potential of soils for bioremediation of contaminants is based on the large number of microorganisms present in this terrestrial habitat. It has been estimated that each gram of soil may contain up to 1010 microbial cells [2]. However, the effectiveness of this process depends on several factors among which the most relevant are: the types of contaminants and its concentration and the physicochemical, mineralogical and microbiological characteristics of the host soil [3]. Although monitored natural attenuation was initially focused on organic contaminants, there is a growing interest in understanding the fate of inorganic contaminants, especially heavy metals, in soils. In the case of heavy metals, the characteristics of the soil matrix have a great influence on their mobility and bioavailability through sorption and leaching bioprocesses. For instance, iron and manganese (hydr)oxides are a major sorbent of contaminant metals in soil [3,4]. Furthermore, since these mineral phases can be formed during biogenic processes, microorganisms can play a key role in decontamination of soils. One of the biggest challenges in this field is how to deal with the huge amount of residues generated in mining processing operations over long periods of time and in many cases located in the earth’s crust. For instance, many of the existing abandoned mine sites that can be catalogued as potential contamination sources of heavy metals are clearly affected by weathering factors. Thus, it is important to optimise the design of mine wastes storage in order to prevent weathering and mobilisation of contaminants. 1.1. Environmental impact of metalliferous mining activities in Europe Over the last century the generation of overburden material has grown dramatically with the increase demand of minerals and metals and the decrease of metal ore grades being mined. Consequently, metal mining activities have generated large volumes of toxic and harmful residues for long periods of time, especially when wastes management was not liable under the existent environmental regulations. Metalliferous mine wastes are generated from mechanical and chemical processing of run of mine ores. Since the efficiency of this extraction process is always less than 100%, there is a high proportion of discarding uneconomic material that contains variable amounts of metallic minerals which can mobilize metals into the environment. It has been estimated that since prehistoric times mankind has mined about 1,150 million tons of heavy metals (Cu, Pb, Co, Zn, Cd and Cr) and that of the whole mineral extracted from

Bioremediation and mining wastes 285

metal mines, a small proportion, usually less than 2%, corresponds to the desired valuable metal while the rest is discharged as different overburden wastes in areas nearby the mine operation [5]. One of the biggest concerns about metalliferous mine wastes is their ability to mobilize metals into the environment. Some facts can help to understand the magnitude of the problem in Spain:

• The metal mining industry in Spain has been focus mainly on the treatment of complex sulfides. Thus, the residues generated in mineral processing installations constitute a potential contamination source of acid drainages and heavy metals. The most known case is the Iberian Pyrite Belt, which extends along the southern part of the Iberian Peninsula from Portugal to Spain, 250 km length and 75 km width, and constitutes one of the world’s largest metallogenic reserves of massive sulfides with over 400 millions tons of massive mineral sulfides and 2,000 millions tons of low-grade ores [6]. Metal exploitation since ancient times has generated a great number of abandoned metalliferous wastes accumulations in this area that include: dumps, open-pits, tailing dams, mine ponds, etc. [7]. These installations, active or abandoned, are one of the main sources of environmental contamination of natural water streams (The Tinto and Odiel rivers) affected by acid mine drainages (AMD). The Andalusian Environmental Department has estimated that the generation of pyrite, a potential soil contaminant, was of 774,000 t in 1994 [8]. Additionally, the Rio Tinto surroundings have proven to be of great geomicrobiological importance due to the high diversity of microbial life [9]. Despite that, the knowledge about the effect of heavy metals on soils is rather scarce.

• Similar environmental issues have been detected in other mining areas in Spain. For instance, in the mining district of Sierra de Cartagena-La Unión (Murcia, southeast Spain), with 2,500 years of mining activities over an area of 100 km2, it has been recorded the existence of: 12 open-pits, 2,363 wastes deposits that include tailing dams and waste rock. The mine wastes occupy an area of approximately 11 km2 and an estimated volume of 200 Mm3 [10,11].

• According to the "Inventory of potentially contaminated sites" emitted by the Environmental Department of the Autonomous Community of Madrid [12], the principal pollutants found in contaminated soils are heavy metals (averaging 54% of the total), high above the environmental impact caused by organic compounds. Far from being an exception, this seems to be a general trend.

• Finally, the Geological and Mining Institute of Spain [13] has estimated that around three fourth parts of the industrial residues produced annually result from open-pit mining exploitations: of the 63,176 millions of tons of residues produced in year 2000, 41,777 corresponded to open-pit mining.

The collapse of mining wastes stockpiles and tailings dams occurred in Europe (Aberfan (Wales, 1966), Stava (Italy, 1985) and, the more recent of Aznalcóllar (Spain, 1998) [14] and Baia Mare (Romania, 2000)) was the seed for starting an internal debate on the environmental impact caused by mining residues in European Union countries which was materialized in the following Communiqués emitted by the European Commission in the year 2000: “Promoting sustainable development in the EU non-energy extractive industry” and “ Safe operation of mining activities: a follow-up to recent mining accidents”. Both Communiqués were the origin of the mining waste Directive 2006/21/EC on the management of wastes from extractive industries to prevent or minimize contamination of waters and soils by acid mine drainages and by heavy


metals leached. Further valuable information on this subject at European level can be consulted at the web page [15]. In the European Union, the mining industry deals with serious economic problems related to the exploitation of low-grade ores. However, the search for new natural metalliferous resources continues (e.g. Las Cruces copper mine near Seville, Spain). Complex sulfides, a crucial raw material in the non-ferrous metallurgy, usually contain pyrite (FeS2) as the main metallic constituent that incorporates to the residues of the process.

Table 1. Data on non-ferrous base metal mining activities in the EU [16].

Mine Site Mining technique

Tailings produced (t/year)

Tailings used in

backfill (%)

Metal discharges (*) (kg/year)

Aitik (Sweden)

Open pit

17,700,000

0

446 Al; 1.7 As; 5.3 Co; 0.2 Cr; 36 Cu; 0.1 Hg; 5.1 Ni; 0.1 Pb; 34.6 Zn

Almagrera (Spain)

Underground 900,000 0 n.a.

Boliden (Sweden)

Open pit Underground

1,457,000

29

156 As; 1 Cd; 72 Cu; 191 Pb; 1,070 Zn

Garpenberg (Sweden)

Underground

910,000

50

18 As; 0.8 Cd; 25 Cr; 40 Cu; 0.3 Hg; 52 Pb; 586 Zn

Hitura (Finland) Underground 518,331 0 24 Fe; 107 Ni

Legnica-Glogow (Poland)

Underground

27,000,000

0

26,164 Ca; 422 As; 591 Cd; 1,160 Cr;

1,435 Cu; 9,495 Fe; 6.33 Hg; 3,376 Pb; 949 Zn

Lisheen (Ireland)

Underground

910,000

50

2,465 Al; 8.1 Cd; 17 Co; 28.5 Cu; 1,412 Fe; 565

Mn; 0.6 Hg; 311.9 Ni; 263 Pb; 2,321 Zn

Mina Reocín (Spain)

Open pit Underground

950,000

94

n.a.

Neves Corvo (Portugal)

Underground 1,370,000 30 n.a.

Pyhäsalmi (Finland)

Underground

213,816

16

4,727 Ca; 7 Cd; 309 Cu; 9,141 Fe; 1,464 Zn

Tara (Ireland) Underground 1,680,000 52 n.a.

Zinkgruvan (Sweden)

Underground

850,000

50

3.6 As; 0.5 Cd; 1.5 Cr; 4.8 Cu; 0.1Hg;

48.3 Pb; 390 Zn (*) Data from Almagrera, Hitura, Lisheen, Mina Reocín, Neves Corvo, Pyhäsalmi and Tara correspond to year 2000 and from Aitik, Boliden, Garpenberg, Legnica-Glogow and Zinkgruvan to year 2001.


Under acidic and oxidizing conditions, favoured by the presence of iron- and sulfur-oxidizing microorganisms, sulfides are solubilized provoking the generation of acid mine drainages (ADM) and the release of metals contained in the mineral matrixes and its subsequent migration to deeper soil horizons or its incorporation to subsurface waters. Table 1 shows data of different European mining sites where non-ferrous base metals are under exploitation. In 2001, the production capacity of underground mines ranged between 65,000 and 1,100,000 tonnes, and between 1,200,000 and 43,700,000 t/year for open-pit mines. Most of that material extracted constituted the final residue of the mining operation. In underground mining, most of the residues produced were used as backfill in the mine. In contrast, that option was not possible in many of the open-pit mines, with the exception of Reocín mine (Cantabria, Spain) where part of the waste rock produced in the operation was backfilled into the open pit. The table also collects the discharges of metals generated in these mining installations whether in the year 2000 or 2001. Those figures agree with the fact that most of the wastes produced in these mines contain pyrite and, therefore, have the potential to produce AMD and the mobilization of heavy metals. 1.2. Heavy metal bioremediation of soil Environmental contamination issues are closely related to the rapid industrial growth experimented in the most technologically advanced countries during the last three centuries. The impact of anthropogenic activities includes contamination of air, soils and sub-surface waters by different contaminants, especially organic compounds and heavy metals (Hg, Cd, Pb, Ni, Cu, Zn, Fe, Cr, Mo, Mn) [17,18,19]. The adverse effects of these hazardous contaminants both on ecosystems at the beginning of the food chain as on humans have promoted sustainable development policies that force companies to adopt strict cleanup regulations on effluents and soils. However, that was not always the case. Currently, the great interest on environmental contamination of waters and soils is supported by the increasing number of books recently published [20-29]. Practically any metallurgical industrial activity poses significant environmental risk whether by producing contaminant liquid effluents or solid residues containing heavy metals [30,31,32]. Specifically, metalliferous mining residues may be leached in natural ecosystems and heavy metals mobilized into deeper soil horizons or even into subsurface waters. This problem, far from being solved, is becoming more and more important as new countries with emerging economies and with a strong mining and metallurgical industry are incorporated to the global market, as in the case of China, India and Brazil [33]. Three strategies have been considered in the remediation of contaminated soils: physical, chemical and biological. Different physical and chemical technologies have been proposed for the clean-up of soils contaminated with heavy metals [27,34]. Physical separation technologies include: soil washing, vitrification, permeable barrier systems, electrokinetic techniques, mechanical screening, hydrodynamic classification, gravity concentration, froth flotation, attrition scrubbing and magnetic or electrostatic separation. Chemical extraction technologies use different chemical reagents to precipitate metals as insoluble compounds, or ion exchange, flocculation and membrane filter processes. On the other hand, the biological option, known as natural attenuation, bioattenuation or intrinsic bioremediation, refers to the use of (micr-)organisms to reduce or remove the effect of harmful contaminants in the environment and seems an attractive alternative to conventional physical and chemical techniques. In addition, in the past years, phytoremediation and rhizoremediation have been developed as a bioremediation


technology that utilizes plants for the removal of metals. Many literature revisions on the subject support the interest for this biotechnology [35-43]. Essentially, natural attenuation is based on natural processes that occur without human participation reducing the environmental impact of contaminated sites, mainly soils and subsurface waters. This objective can be achieved by means of destructive processes that destroy the contaminant (e.g. organic compounds) or by non-destructive processes that although do not destroy contaminants can reduce their impact (e.g. heavy metals). These in situ processes result from a combination of physical actions and chemical and biological reactions mediated by microorganisms which include: biodegradation, dispersion, dilution, sorption, precipitation, volatilization, chemical and biochemical stabilization of contaminants [44]. However, natural attenuation of soils is a decontamination process with a slow kinetics. Bioremediation can be even more effective when external measurements led to improve biological processes are adopted, such as: biostimulation of indigenous microbial populations by adjusting the pH and redox potential or adding nutrients or electron acceptors/donors, bioaugmentation of specialized microorganisms to favour their growth in the ecosystem and accelerated bioremediation considering both criteria [45-49]. A further step in controlling and understanding these natural processes has been monitored natural attenuation (MNA) [50]. Recently, several authors have reviewed management strategies of contaminated lands using monitored natural attenuation in Germany, Denmark, The Netherlands, United Kingdom and USA [51]. In the case of heavy metals, microorganisms (especially bacteria, fungi, yeasts and algae) have the ability of whether immobilized or mobilized these contaminants in natural environments [35,37,52,53]. Metal uptake by biomass can be achieved whether actively by living cells through energy-dependent processes (bioaccumulation) or passively by dead cells through energy-independent processes (biosorption). In fact, biosorption has been proposed as a biotechnology for the decontamination of industrial effluents contaminated with heavy metals [54]. On the other hand, metal mobilization from soil matrixes is also directly related to natural biological processes. In this way, bioleaching processes for metal extraction from ores have been extensively studied and successfully implemented in industrial scale [55]. Undoubtedly, bioremediation is an emerging technology for the treatment of soils contaminated with a great variety of contaminants, as demonstrated by the success achieved in the decontamination of organic compounds. The main motivation for applying this biotechnology for environmental purposes is economics. On average, in situ bioremediation can lower costs by more than two or three times with respect to soil washing or landfilling [28].

Table 2. Advantages and disadvantages of bioremediation.

Advantages Disadvantages Remediation can be done in situ. Process already implemented for

biodegradation of organic compounds. Low environmental impact. Compatible with the use of other

technologies. Low cost.

Slow kinetics. May require monitoring. The requirements to eliminate efficiently

contaminants may vary from site to site. Some contaminants can be present at high

concentrations that inhibit microbial activity. Risk of accumulation of toxic by-products.


Currently, in situ bioremediation and MNA have been accepted as alternative effective technologies to clean up sites contaminated with a great variety of organic contaminants [56]. At the same time, there is a growing interest in bioaugmentation and genetic engineering as a way to improve natural processes [45,57,58]. The principal advantages and disadvantages of this bioprocess are shown in Table 2. The basic requirements for bioremediation of soils contaminated with heavy metals are depicted in Figure 1. Of course, the process requires the existence of a microbial community provided by the own soil with the ability to transform potential metal contaminants into less innocuous substances. Obviously, sustainable microbial growth involves energy production and cellular synthesis from natural substrates which provide both sources and terminal acceptors of electrons. Besides energy requirements, the process is markedly sensitive to environmental conditions, such as: moisture degree, essential nutrients, pH, redox potential, temperature, bacterial heavy metal resistance, toxic substances (including metabolites and minerals) and the presence of high concentrations of protozoa which are predators for microorganisms responsible for remediation of contaminants [59]. Therefore, heavy metals bioremediation of soils should consider physiological and nutritional needs of degradative microorganisms that assure that desirable biotransformations occur in the ecosystem. This is not an easy task, especially because of mass transfer limitations and difficulties for an adequate distribution of compounds involved in biostimulation [60]. In fact, bioremediation is presently considered as a viable but constraint clean up technology due to the uncertainty of results and times. Nevertheless, this perception is changing mainly because of the successful implementation of bioremediation to treat soils contaminated with organic compounds. Then, the following step should be focus on the development of biological processes for the treatment of soils contaminated with heavy metals.

MICROBIAL CONSORTIA

HEAVY METALBIOREMEDIATION

Energyrequirements

Environmentalrequirements

Energy sourceElectron acceptor

MoistureNutrientspHRedox potentialTemperatureMetal resistanceAbsence of toxicmetabolitesAbsence of predators

MICROBIAL CONSORTIA

HEAVY METALBIOREMEDIATION

Energyrequirements

Environmentalrequirements

Energy sourceElectron acceptor

MoistureNutrientspHRedox potentialTemperatureMetal resistanceAbsence of toxicmetabolitesAbsence of predators

Figure 1. Basic requirements for bioremediation of soils contaminated with heavy metals (modified from Álvarez and Illman [28]).


Metalsulfide

Fe(III)-reducingbacteria

Fe(II)-oxidizingbacteria

Fe(III)

Fe(II)

Metalsulfide

Sulfate-reducingbacteria

Sulfur-oxidizingbacteria

H2SO4

H2S

MeoxidizedMereduced

Metalsulfide

Fe(III)-reducingbacteria

Fe(II)-oxidizingbacteria

Fe(III)

Fe(II)

Metalsulfide

Sulfate-reducingbacteria

Sulfur-oxidizingbacteria

H2SO4

H2S

MeoxidizedMereduced

Figure 2. Interactions between microorganisms and soils contaminated with mining residues containing metal sulfide minerals (modified from Picardal and Cooper[3]). Among different bioremediation scenarios, the contamination of soils with metalliferous sulfide mining residues is of especial relevance. As shown in Figure 2, several microorganisms with an active participation in the biogeochemical cycles of iron and sulfur play a crucial role in processes of mobilization and/or immobilization of heavy metals from metal sulfide matrixes, especially pyrite (FeS2). The exposure of hard rock sulfidic bearing ores to air and water is just one example of the potential problems associated with accelerated weathering that result in a well known problem affecting the mining industry commonly referred as acid mine drainage (AMD) or acid rock drainage (ARD) [61,62]. In addition, acid generation can be responsible for metal mobilisation in the ecosystem. 1.3. Partitioning and fate of heavy metals in soils Unlikely organic contaminants metals cannot be degraded neither biologically nor chemically. Although metallic compounds can be transformed into other forms, the environmental risk of metals is highly dependent on its toxicity, oxidation state and level of concentration [29]. Within the group of metals, heavy metals have a high relevance in contamination processes. These are metals with a density higher than 5x103 kg/m3 which are frequently associated to industrial residues in variable amounts. The European Commission has established a black list of contaminants that includes metals such as: arsenic (As), cadmium (Cd), chromium (Cr), copper (Cu), mercury (Hg), nickel (Ni), lead (Pb) and zinc (Zn) [20]. Similar regulatory policies have been revised by the American Environmental Protection Agency (USEPA), a leading organization in this field.


Humans

Animals

Plants

Heavymetals

AIR

SOIL

WATERAnthropogenicprocesses

Naturalprocesses

Humans

Animals

Plants

Humans

Animals

Plants

Animals

Plants

Heavymetals

AIR

SOIL

WATERAnthropogenicprocesses

Naturalprocesses

Figure 3. Sources of heavy metals and their cycling in natural ecosystems (modified from Brady and Weil [29]). The main sources of heavy metals contamination of soils are both of natural and anthropogenic origin [3]. At present, however, the release of heavy metals into the environment is mainly due to human activities that include among others: agriculture (fertilizers, pesticides, etc.) and metallurgical activities (mining, smelting, metal transformation and finishing, etc.). The harmful effects of these metals on plants or animals are evident when present in concentrations above a threshold value. In contrast, some heavy metals (Fe, Co, Cu, Ni, Mn, Zn), in low concentration, serve as micronutrients and are essential elements for the growth of plant and animals [59]. In the cycle of heavy metals in terrestrial ecosystems, depicted in Figure 3, the content of metals builds up from left to right along the trophic chain. In a first approach, soils consist of a great variety of constituents that include: particles of inorganic minerals, dead organic matter, living microorganisms, aqueous solutions and gases that fill the voids. Nevertheless, the inorganic matter is by far the most abundant fraction in soils. Heavy metals in natural environments are associated to different soil fractions, according to the scheme shown in figure 4. Thus, there is a clear relationship between the distribution of metals in each fraction and the degree of metal mobilization in the natural environment. In this way, metals are more bio-available in the exchangeable than in the residual fraction. Generally, heavy metals associated to the exchangeable or soluble fraction are easily available for plants and animals. The organic fraction includes those metals bound to organic matter that can be released into the environment in longer periods of time. Furthermore, heavy metals can be distributed in the form of carbonates, oxides and hydroxides (carbonates and hydroxides fractions), with a lower degree of mobilization, especially in alkaline soils. Finally, the highest stability of metals is related to sulfides and residual fractions. Nevertheless, metal sulfides can be solubilized in acid soils of oxidizing characteristics, especially in the presence of microbial populations with iron- and sulphur-oxidizing abilities. On the other hand, the residual fraction contains the most


Exchangeable Organic Carbonates

Hydroxides

Sulfides

Residual

Higher Bio-availability

Lower Bio-availability

Figure 4. Schematic diagram of heavy metals distribution in the different fractions contained in soils and its degree of bioavailability based on sequential extraction procedures. inert and stable phases and, in consequence, metals associated to this fraction present the lowest mobilization in the soil. Then, (im)mobilization of metals in soils is closely related to its partitioning between the different fractions that constitute the soil which, finally, is a function of the conditions prevailing in the natural ecosystem (pH, redox potential, salinity, microbial consortia, temperature, presence of minerals with a toxic effect on microorganisms or with high retention metal uptakes, etc.) [29,63]. 1.4. Relevant aspects in the metal-soil-microorganism interaction Although monitored natural attenuation was initially focused on monitoring organic contaminants, there is a growing interest in understanding the fate of inorganic contaminants, especially heavy metals, in soils. Such interest has been extended to the role play by the different solid fractions present in a contaminated soil, in relation with the implications of microbiological processes and the partitioning of contaminants. Thus, one of the key aspects of the overall process resides in the establishment of compatibility or incompatibility criteria between contaminants, microorganisms involved and soil components which could either contribute to a better or worse distribution of heavy metals in the different soil fractions, i.e., reducing the toxic effect of contaminants. Historically, the indiscriminate dumping of mine waste was done on the basis that the purification capacity of soils is almost unlimited. Indeed soils act as real filters exchanging matter with aqueous solutions, but the sorption capacity of soils is limited and varies depending on the type of soil, the nature of the contaminant and the microbiota present. Then, in the case of heavy metals, its mobilization is strongly conditioned by microbiological processes and by the characteristics of the different solid fractions present in the soil (Figure 5). The interaction between heavy metals and soils is widely documented in the literature [63-66]. The distribution of heavy metals in soils takes place between the soil solution (liquid phase) and the different constituents of the solid phase that serve as binding sites for metals. The different pathways of metal interaction with soils are depicted in Figure 6. According to such scheme, metals can be found: 1) in solution, whether as free or complex ions; 2) adsorbed onto solid particles, whether as amorphous or crystalline inorganic compounds (hydroxides, carbonates, sulfides and sulfates) or as organic compounds; 3) bioaccumulated on live biomass (including phytoremediation and rhizoremediation


Hea

vym

etal

sIN

FLO

W

Hea

vym

etal

sO

UTF

LOW

Organic matter Amorphous inorganic substances

Microorganisms Crystalline inorganic substances

Hea

vym

etal

sIN

FLO

W

Hea

vym

etal

sO

UTF

LOW

Organic matter Amorphous inorganic substances

Microorganisms Crystalline inorganic substances Figure 5. Soil basic components affecting mobilization of heavy metals (adapted from Yong and Mulligan [25]).

Inorganic matter(Sorption)

Organic matter(Biosorption)

PrecipitationOcclusion

Bioaccumulation

Men+

(aq)Organic

complexationInorganic

complexation

Inorganic matter(Sorption)

Organic matter(Biosorption)

PrecipitationOcclusion

Bioaccumulation

Men+

(aq)Organic

complexationInorganic

complexation

Figure 6. Possible interactions of heavy metals in soils contaminated with mining residues (redrawn from Suthersan [23]). processes [39]) or biosorbed on dead biomass; 4) occluded within mineral particles or 5) precipitated. Nevertheless, these chemical equilibria are dynamic and, depending on the physico-chemical conditions prevailing (pH, redox potential, salinity, etc.), these metals can be released into the soil solution affecting its availability to the biota and microbiota or being susceptible of migration to subsurface waters [67]. Thus, the bio-availability of metals in terrestrial environments greatly depends on its distribution between the different soil fractions and its interaction with dead or living micro-organisms (Table 3).


Table 3. Relative bio-availability of metals retained in soils.

Metal speciation/distribution Relative availability

Ions in solution Easily available

Inorganic complex ions Relatively available

Organic complex ions (chelates) Less available

Compounds precipitated Only available by chemical transformation

Adsorbed in biological matrixes Available once decomposed

Incorporated into mineral lattices Only available after weathering

Then, in order to evaluate the environmental risk by heavy metals it is not enough to determine its total amount in the soil but also its availability among the different soil fractions. The sequential extraction is an analytical method extensively used in the determination of heavy metals partitioning. This method is based on the use of specific chemical reagents for dissolving preferentially metals from each soil fraction providing, in this way, useful information on how heavy metals are distributed into the different solid fractions of a contaminated soil, such as: interchangeable, organic, carbonates, oxides/hydroxides, sulfides and residual [68,69,70]. 1.5. Case Study: Effect of bioaugmentation on metal mobilization in soil microcosms This microcosm study was performed on a contaminated soil sampled in a mining area in SE Spain (Sierra Cartagena-La Unión, Murcia) (Figure 7). In this mining district, exploited since Roman times for Ag, Pb and Zn, metalliferous mine wastes dumping on natural soils has been a common practice. The soil particle size was conditioned to <1 mm and the main metal contaminants determined by chemical analysis were: 25.01% Fe, 0.52% Zn and 0.80% Pb. The mineralogical characterization by X-ray diffraction revealed the presence of pyrite in a matrix of quartz with aluminium and iron silicates. In addition, sequential extraction chemical analysis, performed in triplicate and considering six soil fractions (exchangeable, organic, carbonates, hydroxides, sulfides and residual), confirmed that the main soil contaminants were in the form of metal sulfides (see figure 11). The microorganisms used in the bioaugmentation process were grown from the own soil. The Fe- and S-oxidizing ability of the strain was achieved by contact between the soil (10 g), 0K nutrient medium (90 mL) and ferrous sulphate (1 g/L Fe2+) or elemental sulfur (1 g), respectively. Before inoculation, strains were grown for 20 days in three consecutive transfers. The morphology of bacterial cells of both strains is shown in figure 8. The microcosm tests were performed in Erlenmeyer flasks, containing 10 g of contaminated soil and 90 mL of deionized water, in an orbital shaker at 150 rpm and 35°C. The contaminated soil was tested both in sterilized (microcosm Control) and non-sterilized conditions containing in the latter case indigenous microorganisms (microcosm Indigenous) or bioaugmented with 10 mL of a microbial strain of Fe-oxidizing or S-oxidizing microorganisms (microcosms Fe-ox and S-ox, respectively). Periodically, pH, ORP and metal concentration (Fe and Zn) were measured in solution. After 49 days, the solid residues were dried and analyzed following a sequential extraction procedure similar as for the as-received contaminated soil.


Figure 7. Site sampled showing weathering impact.

Figure 8. Electron micrographs of Fe-oxidizing (left) and S-oxidizing (right) bacterial cells grown from the contaminated soil. 1.5.a. pH and redox potential measurements pH and ORP are physicochemical parameters that can be used to predict microbial activity. The evolution of both parameters depicted in Figure 9 indicates that the beginning of bacterial activity in the microcosms establishes differences from day 7th on. In fact, the higher acidity associated to more oxidizing conditions was recorded in the bioaugmented microcosms (S-ox and Fe-ox). This behaviour would be related to a greater production of H2SO4 during the chemical oxidation of mineral sulfides mediated by S-oxidizing and Fe-oxidizing microorganisms. 1.5.b. Metals dissolution The aggressive conditions generated by microbial activity would promote the dissolution of metals from the soil. As shown in figure 10a, the iron concentration in


a) b)

1

2

3

4

5

6

0 10 20 30 40 50 Days

pH

Control Indigenous

Fe-ox S-ox

200

300

400

500

600

700

0 10 20 30 40 50Days

OR

P (m

V v

s Ag/

AgC

l)

Control Indigenous

Fe-ox S-ox

Figure 9. Evolution of pH (a) and ORP (b) versus time. a) b)

0

2

4

6

8

10

12

14

0 10 20 30 40 50Days

Fe i

n so

lutio

n (%

)

Control Indigenous

Fe-ox S-ox

0

10

20

30

40

50

60

0 10 20 30 40 50Days

Zn

in so

lutio

n (%

)

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Fe-ox S-ox

Figure 10. Metals dissolved: iron (a) and zinc (b) in the microcosms tested for 49 days. solution increased gradually in both bioaugmented microcosms and the higher iron dissolution was recorded in the microcosms inoculated with S-oxidizing microorganisms. This seems to indicate a better adaptation of the S-oxidizing strain to the contaminated soil in agreement with a more pronounced decrease of pH and increase of ORP, as observed in figure 9. Unlike iron, the zinc dissolution rate was initially faster in all microcosms (figure 10b). The more marked differences were recorded after day 7th. That would be an indication of a dissolution process in two stages: in the first stage, zinc is easily leached from the soil while oxidizing conditions, generated by microbial activity, are required in the second stage. 1.5.c. Metal distribution in solid residues The biochemical reactions that took place in the microcosms affected the composition of both the aqueous and the solid phases. After 49 days, solid residues were analyzed following a sequential extraction procedure, a slight variant of the EPA procedure, to determine the new distribution of metal contaminants (Fe, Zn and Pb) in the different soil fractions.


In the case of iron (Figure 11a), its mobilization from the sulfides to the residual soil fraction and to the leached phase is quite evident because of the low iron concentration in the more labile soil fractions (exchangeable, organic, carbonates and hydroxides). Considering the values of the as-received contaminated soil, the marked difference observed between the biotic microcosms, especially the microcosm S-ox, and the abiotic microcosm, Control, would be attributable to the action of microorganisms (both inoculated and indigenous) present in the contaminated soil. On the other hand, zinc was easily mobilized from the more labile soil fractions to the aqueous phase (Figure 11b) which is an indication that the zinc weakly adsorbed can be dissolved from those soil fractions in mild environmental conditions of pH and ORP. Since more oxidizing conditions are required in order to release zinc from the sulfides soil fraction such effect is clearly detectable in the microcosms with microbiological activity, in agreement with the dissolution of Zn observed during the second stage (Figure 10b). Then, the absence of microbial activity in the microcosm Control would clearly condition the dissolution of Zn from the less labile soil fractions.

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Figure 11. Distribution of metals in the different soil fractions for the residues collected in the microcosms: a) iron, b) zinc and c) lead.


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