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Science of the Total Environment 563–564 (2016) 302–319

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Science of the Total Environment

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Does bioleaching represent a biotechnological strategy for remediation ofcontaminated sediments?

Viviana Fonti ⁎, Antonio Dell'Anno, Francesca BeolchiniDepartment of Life and Environmental Sciences, Università Politecnica delle Marche, Via Brecce Bianche, 60131, Ancona, Italy

H I G H L I G H T S G R A P H I C A L A B S T R A C T

• Bioleaching may represent a sustainablestrategy for contaminated dredgedsediments

• The performance is greatly influencedby several abiotic and biotic factors

• Geochemical characteristics and metalpartitioning have a key role

• Sulphide minerals in the sediment are afavorable element

• Microorganisms other than Fe/S oxidisersmay open new perspectives

⁎ Corresponding author.E-mail address: [email protected] (V. Fonti).

http://dx.doi.org/10.1016/j.scitotenv.2016.04.0940048-9697/© 2016 Elsevier B.V. All rights reserved.

a b s t r a c t
a r t i c l e i n f o
Article history:Received 9 November 2015Received in revised form 30 March 2016Accepted 1 April 2016Available online 30 April 2016

Editor: F.M. Tack

Bioleaching is a consolidated biotechnology in the mining industry and in bio-hydrometallurgy, where microor-ganisms mediate the solubilisation of metals and semi-metals from mineral ores and concentrates. Bioleachingalso has the potential for ex-situ/on-site remediation of aquatic sediments that are contaminated with metals,which represent a key environmental issue of global concern. By eliminating or reducing (semi-)metal contam-ination of aquatic sediments, bioleaching may represent an environmentally friendly and low-cost strategy formanagement of contaminated dredged sediments. Nevertheless, the efficiency of bioleaching in this context isgreatly influenced by several abiotic and biotic factors. These factors need to be carefully taken into account be-fore selecting bioleaching as a suitable remediation strategy. Here we review the application of bioleaching forsediment bioremediation, and provide a critical view of the main factors that affect its performance.We also dis-cuss future research needs to improve bioleaching strategies for contaminated aquatic sediments, in view oflarge-scale applications.

© 2016 Elsevier B.V. All rights reserved.

Keywords:bioleachingbioremediationaquatic sedimentsmetal contamination

Contents

1. Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3032. Metal bioleaching: Microorganisms and mechanisms . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 303

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303V. Fonti et al. / Science of the Total Environment 563–564 (2016) 302–319

2.1. Microbes in sediment bioleaching . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3032.2. Metal-sulphide oxidation mechanisms . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 308

3. Sediment geochemical properties and their effects on bioleaching . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3103.1. General sediment characteristics . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3103.2. Fate of metals in sediment bioleaching . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3113.3. Other sediment characteristics to be considered in bioleaching applications . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 311

4. Effects of the main operating variables . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3124.1. pH and oxidation/reduction potential. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3124.2. Growth substrates . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 312

4.2.1. Elemental sulphur . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3124.2.2. Thiosulphate . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3134.2.3. Ferrous ion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3134.2.4. Pyrite . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3134.2.5. Choice of growth substrate . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 313

4.3. Sediment concentration . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3134.4. Temperature . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3134.5. Oxygen . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3134.6. Considerations on experimental procedures . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 314

5. Main drivers influencing metal bioleaching from sediment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3146. Feasibility and sustainability . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3157. Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 315Acknowledgements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 316Appendix A. Supplementary data . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 316References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 316

1. Introduction

The management of contaminated aquatic sediments and dredgedmaterials represents an environmental problem of major concern. InEurope, 300 to 400 million m3 of contaminated sediment is dredgedevery year from marine and freshwater ecosystems (such as for mainte-nance of river embankments and of navigational depth; Bortone et al.,2004; http://www.ceamas.eu). With their high contamination levels,dredged sediments often need to be specifically relocated. Legislationfor the handling of dredged materials is complex and there are noharmonised regulations at the European level. Dredged materials aredealt with at the intersection of the Water, Waste and Marine StrategyFramework Directives (i.e., directives 2000/60/EC, 2008/98/EC and2008/56/CE, respectively, of the European Parliament, and valid withinEU countries).

The main management options for contaminated sediments arelandfill disposal and confined aquatic disposal, although alternativesare needed because of the limited sites available, the high cost and thelow environmental sustainability that characterise such solutions(Bortone et al., 2004; Adriaens et al., 2006; Agius and Porebski, 2008).An alternative is decontamination of the dredged materials, in view oftheir potential beneficial use in the building industry, for beach nourish-ment and for other applications (Lee, 2000; Ahlf and Förstner, 2001;Barth et al., 2001; Siham et al., 2008).

Bioleaching is the application of acidophilic microbes with Fe/Soxidising metabolism to promote solubilisation of metals from solidmatrices, and it is one of the most well-established and industriallyapplied biotechnologies. Consortia of Fe/S oxidising bacteria and otheracidophilic microbes are applied in large-scale plants to improve theextraction of precious and non-precious metals from sulphides or(Fe)-bearing ores (Olson et al., 2003; Rawlings and Johnson, 2007;Brierley and Brierley, 2013). More recently, bioleaching has been inves-tigated as an environmental technology for the recovery of valuablebase metals from urban and industrial waste (e.g., printed circuitboards, spent batteries, cathode ray tubes, spent refinery catalysts,sewage sludge; Brandl et al., 2001; Zhao et al., 2008; Pathak et al.,2009; Beolchini et al., 2012; Johnson, 2014). Moreover, bioleaching isoften considered as a promising ex-situ/on-site bioremediation strategyfor eliminating/reducing metal contamination in dredged sediments, inview of their potential beneficial use (Blais et al., 1993; White et al.,1998; Bosecker, 2001; Tabak et al., 2005; Akcil et al., 2014).

Although bioleaching is often assumed to be an environmentallyfriendly and low-cost technique, its feasibility and sustainability as asediment bioremediation strategy have not been studied sufficientlyto date. There have been investigations into the factors and operationalvariables that can influence sediment bioleaching performance(e.g., microorganisms, growth substrates, temperature, concentrationof sediment, type of bioreactor). However, the majority of these studieshave been based on a trial-and-error approach, and the geochemicalproperties of the sediments themselves have barely been considered.Sediments appear to be very challengingmatrices compared tomineralores. In particular, metal contaminants tend to associate with sedimentcomponents other than the crystalline lattice of primary minerals(e.g., adsorption on organic molecules, association as exchangeableions). As a consequence, bioleaching know-how and principles thatare so well-established in biomining and bio-hydrometallurgy canonly be applied partially to sediment bioleaching (e.g., metabolicpathways, metal solubilisation mechanisms; microbial adaptation tominerals; see Rohwerder et al., 2003, Rawlings and Johnson, 2007,Vera et al., 2013 and references within).

In this review we provide a critical analysis of the main constraintsthat influence the effectiveness of bioleaching as a sediment remedia-tion strategy. In particular, we aim to identify themechanisms that reg-ulate the potential of metal removal, the factors that are the mostrelevant, how these interact, and which aspects can limit realbioleaching applications. Here we thus: 1) explore the potential ofbioleachingmicroorganisms in view of real sediment bioleaching appli-cations; 2) discuss sedimentmetal interactions; and 3) provide a criticalanalysis of the scientific literature relating to the application ofbioleaching to metal-contaminated sediments. Our analysis highlightsthe major gaps that need to be filled in the future, and provides thetools for facilitating the decision-making processes in view of largescale applications. Fig. 1 shows a roadmap for a better orientation inthis review paper.

2. Metal bioleaching: Microorganisms and mechanisms

2.1. Microbes in sediment bioleaching

The main microbes that can ‘bioleach’ metals from solids are acido-philic bacteria and archaea that have dissimilatory metabolism that isbased on the oxidation of S0, reduced inorganic sulphur compounds

http://www.ceamas.eu

Fig. 1.Manuscript roadmap.

304 V. Fonti et al. / Science of the Total Environment 563–564 (2016) 302–319

(RISCs), or Fe(II) ions (Quatrini et al., 2006, 2009; Bonnefoy andHolmes,2012; Liu et al., 2012). High concentrations of chemical species withhigh leaching power (mainly H2SO4 and Fe3+) are produced. Moreover,prokaryotes within this non-phylogenetic group share advantageouscharacteristics for bio-hydrometallurgical applications: (i) tolerance tohigh concentrations of metal contaminants; (ii) tolerance to extremelyacidic conditions; and (iii) acidification of the environment and/or in-creased oxidation/reduction potential. A complete list of the acidophilicFe/S oxidising strains that have already been applied in bio-hydrometallurgy, or that might have relevant applications, is providedin SupplementaryMaterials (Table SM1). Bioleaching bacteria are distrib-uted among the α- β- γ-Proteobacteria (Acidithiobacillus, Acidiphilium,Acidiferrobacter, Ferrovum), Nitrospirae (Leptospirillum), Firmicutes(Alicyclobacillus, Sulfobacillus), and Actinobacteria (Ferrimicrobium,Acidimicrobium, Ferrithrix). Bioleaching archaea mostly belong toCrenarchaeota (Sulfolobus, Acidianus, Metallosphaera, Sulfurisphaera), al-though two acidophilic Fe(II)-oxidisers are affiliated with Euryarchaeota(Ferroplasma acidiphilum and Ferroplasma acidarmanus; Rohwerder et al.,2003; Norris, 2007; Schippers et al., 2010; Hedrich et al., 2011).Mesophilic strains (i.e., those that can grow below 35–40 °C) are exclu-sively Gram-negative bacteria, although Acidithiobacillus caldus can be ac-tive at 50 °C (Hallberg and Lindstrom, 1994). Moderate thermophiles(i.e., those that can grow at about 45 °C or a little higher) are almost ex-clusively Gram-positive bacteria (e.g., S-oxidisers: Sulfobacillus andAlicyclobacillus; Fe-oxidisers: Ferrimicrobium and Acidimicrobium), withthe sole exception of the archaeal genus Thermoplasma (Euryarchaeota).Similar to Acidithiobacillus ferrooxidans, Sulfobacillus thermosulfidooxidansand Sulfobacillus acidophilus can live on the oxidation of either Fe or S,although their optimal growth occurs under mixotrophic conditions(Supplementary Materials Table SM-1). S-oxidising (e.g., generaSulfolobus andMetallosphaera) and Fe-oxidising (e.g., genus Ferroplasma)archaea are thermophilic (i.e., they can grow at 70 °C or higher) (Dopsonand Johnson, 2012; Johnson et al., 2014; Vera et al., 2013).

Aerobic chemolithotrophic growth on S0, RISCs or Fe(II) is not thesole metabolism of these microorganims. Many acidophiles aremixotrophs and they need a source of organic carbon to support the

oxidation of S, while others are obligately heterotrophs. Several S-oxidising acidophiles are obligate aerobic (e.g., Acidithiobacillusthiooxidans, At. caldus), while others can use Fe(III) as an alternativeelectron acceptor and grow under anoxic conditions (e.g., At.ferrooxidans and Acidiferrobacter thiooxydans) (Mangold et al., 2011;Valdés et al., 2008). The dissimilatorymetabolism is often specialised ei-ther for the use of S and RISCs or of Fe as the electron donors, althoughsome species can grow both on Fe-oxidisation and S-oxidisation, likeAcidithiobacillus ferrooxidans. Indeed,At. ferrooxidans can live aerobicallyusing either Fe(II) or reduced inorganic S compounds as electron do-nors, and anaerobically using oxidation of hydrogen (or S) coupledwith Fe(III) reduction (Ohmura et al., 2002; Hallberg et al., 2011;Johnson, 2012; Osorio et al., 2013). Due to this particular ability to useboth S and Fe, At. ferrooxidans has beenwidely used as amodelmicroor-ganism to investigate metal solubilisation mechanisms mediated byleaching bacteria.

Recently, a new bacterial strain with similar metabolic characteris-tics was identified within the Acidithiobacillus genus and namedAcidithiobacillus ferrivorans (Hallberg et al., 2010). Other species withinthe Acidithiobacillus genus can oxidise S but not Fe. Fe/S oxidising bacte-ria adhere to and colonise the mineral surface by creating a biofilm,where the majority of the cells attach to the sulphide surface by theproduction of extracellular polymeric substances, while planktonic bac-terial cells remain floating in the solution (Sand et al., 1995, 2001;Rohwerder and Sand, 2007; Diao et al., 2014).

Although several acidophilic strains are used largely in variousbioleaching applications, a very limited number of these have been in-vestigated for the bioleaching of contaminated sediments. In this field,the most studied strains are the mesophiles At. ferrooxidans, At.thiooxidans, Thiobacillus thioparus, Leptospirillum ferrooxidans andLeptospirillum ferriphilum (Table 1). With the exception of T. thioparus,which is active also at circumneutral pH, the choice of these strainsarises from the first biomining and bio-hydrometallurgical applications,where they were found to dominate the microbial communities of bio-reactors (Rawlings and Johnson, 2007; Rohwerder et al., 2003).T. thioparus appears to adapt better to the sediment, due to oxidisation

Table 1Summary of the main bioleaching treatments on aquatic sediments, with details about metal contamination, experimental plans and results. Studies listed in the table have been selected for their particular contribution to the knowledge aboutsediment bioleaching.

Ref. Sediment sample Extra treatments Strainsa Operating conditions Energy source Metal content(ppm)

Solubilisationyieldsb

Blais et al., 2001 Freshwater sediment: short acidic washing steps At. ferrooxidans Pilot scale: 350 L batch reactor Fe2SO4 (Fe 0.3 g/L) Cd n.a. 82–100%Zn n.a. 80–87%Cu n.a. 44–70%Pb n.a. 14–33%

Lachine Canal (Québec, Canada) Pre-adaptation stage Sediment 30 g/L Ni n.a. 12–21%Cr n.a. 6–16%Inoculum 20% (v/v)

Initial pH = 4 with H2SO4

Room temperatureBeolchini et al., 2009 Marine sediment: none Autotrophs: Lab-scale: flask S0 2.5–5 g/L

Glucose 0.01 g/LCd 1.2 ca. 100%Cu 410 ca. 100%

Ancona's port (Italy, Adriatic Sea) Zn 500 ca. 90%Sediment 20 g/L Hg 5.4 ca. 90%

At. ferrooxidans Cr 141 ca. 50%Pb 76 ca. 40%Ni 105 ca. 40%

At. thiooxidans Inoculum 10% (v/v) As 21 ca. 5%L. ferrooxidans Initial pH = 2 with H2SO4

Room temperatureHeterotrophs:A. cryptum

Chartier et al., 2001 Freshwater sediment:Québec, Canada:

– Lachine Canal– Aylmer Lake– Iles-aux-Chats

Post-treatment:rinsing with NaCl

At. ferrooxidansPre-adaptation stage

Lab-scale: flaskSediment 30 g/LInoculum 20% (v/v)Initial pH = 4 with H2SO4

Room temperature

Fe2SO4 (Fe 0.3 g/L) Lachine Canal:Cd n.a. 100%Zn n.a. 75%Cu n.a. 52%Pb n.a. 33%Ni n.a. 22%Cr n.a. 8%Aylmer Lake:Cd n.a. 100%Zn n.a. 59%Cu n.a. 52%Ni n.a. 18%Pb n.a. 13%Cr n.a. 11%Iles-aux-Chats:Zn n.a. 67%Cd n.a. 62%Pb n.a. 33%Cu n.a. 30%Ni n.a. 20%Cr n.a. 3%

Chen and Lin, 2001a Freshwater sediment:Ell Ren River (Taiwan)

none At. ferrooxidansT. thioparusPre-adaptation stage

Lab-scale: 3 L batch reactorSediment 10÷70 g/LInoculum 5% (v/v)No pre-acidificationT = 20 °C

S0 5 g/L Cu 191 95%Mn 424 73%Ni 50 65%Zn 401 60%Pb 143 47%Cr 74 18%

Fang et al., 2011 Freshwater sediment:Hai Bo River (China)

no pre-treatmentBio-precipitationafter treatment

At. thiooxidansPre-adaptation stage

Semi-pilot: 50 L batch reactorSediment 97 g/LInoculum 50% (v/v)Room temperature

S0 3 g/L Zn 732 78%Cu 174 63%Cr 206 29%

Fonti et al., 2013a Marine sediment:Piombino's port,

Pre-treatment:washing for

Autotrophs:At. ferrooxidans

Lab-scale: flaskSediment 10 g/L

FeSO4 (Fe 0.9 g/L) Zn 1030 38%Cd 1.8 16%

(continued on next page)

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Table 1 (continued)

Ref. Sediment sample Extra treatments Strainsa Operating conditions Energy source Metal content(ppm)

Solubilisationyieldsb

Tyrrhenian Sea (Italy) removing salts At. thiooxidansL. ferrooxidansHeterotrophs:A. cryptum

Inoculum 10% (v/v)Initial pH = 2 with H2SO4

Room temperature

Cr 140 0%As 48 0%

Fonti et al., 2013b Marine sediment:Italy:

– Piombino's port (Tyrrhenian Sea)– Livorno's port (Tyrrhenian Sea)– Ancona's port (Adriatic Sea)

Pre-treatment:washing for removing salts

Autotrophs:At. ferrooxidansAt. thiooxidansL. ferrooxidansHeterotrophs:A. cryptum

Lab-scale: flaskSediment 100 g/LInoculum 10% (v/v)Initial pH = 2 with H2SO4

Room temperature

FeSO4 (Fe 8.9 g/L)S0 1 g/LGlucose 0.1 g/L

Piombino's port:Zn 1030 76%Cd 1.8 39%Pb 200 5%Ni 29 35%Cr 140 1%As 48 1%Livorno's port:Zn 170 55%Ni 70 35%As 11 23%Cd 0.5 8%Cr 120 8%Pb 28 0%Ancona's port:Zn 83 52%Ni 40 35%As 10 20%Cr 70 6%

Gan et al., 2015a Freshwater sediment:Xiangjiang River (Xiawan port, China)

air-dried At. ferrooxidansAt. thiooxidansL. ferriphilum

Lab-scale: flasksLab-scale: 1.5 L batch reactorSediment 50 g/LInoculum 10% (v/v)Initial pH = 4 with H2SO4

(flasks only)T = 30 °C

S0 (2.5–10 g/L)Fe

Zn 4294 96%Mn 2417 92%Cd 84 88%As 84 ca. 30%Pb 581 ca. 20%

S2 (2.5–10 g/L) Hg 81 ca. 10%

Kim et al., 2005 Synthetic sediment: 10% sand, 43% organictop soil, 47% clay (75% illite +25% chlorite)Spiked with Cd and Ni sulphides

none At. ferrooxidans Lab-scale: 1 L batch reactorSediment 15 g/L

FeCl2 (Fe 16 g/L) Cd 200 ca. 80%Ni 200 ca. 60%

Löser et al., 2001 Freshwater sediment:Weisse Elster River (Germany)

made soil-like(6 year at open)

Indigenous S- oxidizersBiostimulation

Lab-scale: flaskSediment 100 g/LNo pre-acidificationT = 20 °C

S0 (0.5÷5 g/L) Zn 3320 ca. 90%Ni 287 ca. 80%Cd 39 ca. 80%Cu 352 ca. 70%Cr 536 ca. 65%Pb 324 ca. 15%

Löser et al., 2006c, 2007;Seidel et al., 2004; 2006b

Freshwater sediment:Weisse Elster River (Germany)

made soil-like(6 year at open)


Pilot scale: solid bed reactor(1000 Kg)

S0 (20 g/Kg) Zn 3170 80–90%Cd 37 70–90%Ni 289 65–90%Cu 307 20–50%Cr 503 b10%Pb 303 b10%

Löser et al., 2006a Freshwater sediment:Germany:

- Radeburg dam (Röder River)- Kleindalzig detritus trap (Weisse Elster River)- Kleindalzig detritus trap (Weisse Elster River)- Meuschau sluice (Saale River)- Süßer Lake- Bremen's port (Weser River)

Pre-treatment:Radeburg dam: noK. detritus trap:

– 6 years at open (oxic)– Dewatering (anoxic)

Meuschau: noSüßer Lake: noBremen: aged (disposal site)


Lab-scale: flaskSediment 100 g/LNo pre-acidificationT = 20 °C

S0 (2 g/L) Röder River:Zn 1750 ca. 90%Cd 83 ca. 90%Ni 90 ca. 80%Cu 380 ca. 50%Cr 480 ca. 5%Pb 720 ca. 0%Weisse Elster (oxic):Zn 3291 ca. 80%Cd 36 ca. 80%Ni 286 ca. 70%


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Cu 322 ca. 50%Cr 515 ca. 5%Pb 312 ca. 0%Weisse Elster (anoxic):Zn 22 ca. 85%Cd 23 ca. 80%Ni 36 ca. 60%Cu 136 ca. 50%Cr 146 ca. 2%Pb 231

140ca. 0%

Saale River:Zn 871 ca. 10%Cd 3 ca. 0%Ni 70 ca. 2%Cu 11 ca. 0%Cr 144 ca. 0%Pb 114 ca. 0%Süßer LakeZn 5448 ca. 20%Cd 13 ca. 5%Ni 43 ca. 0%Cu 785 ca. 0%Cr 84 ca. 0%Pb 886 ca. 0%Bremen's portZn 361 ca. 0%Cd b2 ca. 0%Ni 41 ca. 0%Cu 99 ca. 0%Cr 106 ca. 0%Pb 93 ca. 0%

Seidel et al., 2004 Freshwater sediment:Weisse Elster River (Germany)

Pre-conditioning:planted with reedcanary grass


Pilot scale: solid bedreactor (1000 Kg)

S0 (20 g/Kg) Zn 1958 81%Mn 1326 71%Cd 9 70%Ni 136 65%Co 42 61%Cu 142 21%Cr 270 2%Pb 135 1%

Tsai et al., 2003a Freshwater sediment:Ell Ren River (Taiwan)

none Indigenous S- oxidizersBiostimulation

Lab-scale: flaskSediment 5.2–32.3 g/LInoculum 2% (v/v)No pre-acidificationT = 25 °C

Na2S2O3 (S 2.15 g/L) Cu 1088 99% (non-residual)c

Zn 1241 99% (non-residual)c

Ni 199 95% (non-residual)c

Cr 337 94% (non-residual)c

Pb 374 86% (non-residual)c

a : A.: Acidiphilium; At.: Aciditiobacillus; P.: Pseudomonas; T.: Thiobacillus.b : solubilisation efficiencies for the best operating conditions; data refer only to bioleaching treatment.c : authors reported solubilisation efficiencies excluding the amount of each metal in the residual fraction.


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of S0 under circumneutral pH conditions (Chen and Lin, 2000a, 2001a).In view of large-scale applications, the use of pH-tolerant strains canoffer the advantage of avoiding or reducing the need for sediment pre-acidification. Similarly, Zhu et al. (2013) reported that sediment pre-acidification can be avoided in the presence of a consortium of autotro-phic and heterotrophic strains (e.g., At. thiooxidans with Pseudomonasaeruginosa). On the other hand, thermophilic strains can offer kineticadvantages over mesophiles, and they might represent better choicesfor large-scale applications.

The temperatures in heap leaching can sometimes be up to 50- °C to60 °C, due to heat from the oxidation reactions (Watling, 2006; Zhanget al., 2015). To date, only one study has tested non-mesophilic bacteriaS. thermosulfidooxidans and At. caldus for sediment remediation (Ganet al., 2015a). Fingerprinting techniques, like FISH (fluorescent in situhybridisation) and PCR-based techniques, have shown that autochtho-nous S-oxidisers are present in certain sediments as a minor fractionof the microbial community (Chen and Lin, 2004; Zhu et al., 2013).Some studies have investigated the biostimulation of indigenous S-oxidisers in sediments, despite inoculation of allochthonous strains,with high solubilisation levels of Zn, Cd and Ni achieved (Table 1).

A variety of heterotrophic acidophiles have been isolated from thesame environments where Fe/S oxidisers have been isolated. Some ofthese have been reported to favour metal bioleaching from mineralores (Fournier et al., 1998; Johnson, 1998, 2001, 2008; Vera et al.,2013). Heterotrophic Fe-reducing bacteria, like Acidiphilium cryptum,can reduce ferric ions generated by Fe-oxidisers, which feeds an Fe(II)cycle. In particular, although Fe(II) ions are generated during metal-sulphide dissolution (Section 2.2, Eqs. (1)-(2),(12)-(13)), the co-precipitation of Fe(III) is common during bioleaching, with the conse-quent decrease in Fe availability in the medium. Acidiphilium strainscan regenerate Fe(II) ions also from insoluble amorphous or crystallineminerals, like Fe(OH)3 and jarosites (Johnson, 1998; Küsel et al., 1999;Daoud and Karamanev, 2006; Rawlings, 2007). However, sedimentbioleaching studies have provided evidences that A. cryptum does notalways significantly improve metal removal from contaminated sedi-ments. A. cryptum appears to be very sensitive to the presence of highconcentrations of sediment, and even if Fe-reducers are active in trig-gering the Fe3+/Fe2+ cycle, the effects on metal solubilisation canevary among sediment samples (Beolchini et al., 2009, 2012, 2013;Fonti et al., 2013a).

Considering the limited number of strains that have been tested forthe removal ofmetal contamination frompolluted sediments, the exploi-tation of different microbes might open new perspectives in this field.Particular functions can be obtained from extreme environments(e.g., S0/S= oxidisers mine run-off sediments; Salo-Zieman et al., 2006;S-oxidisers isolated in karst caves; Jones et al., 2012) and the potentialfor Fe/S oxidising archaea or Gram-positive bacteria can be also investi-gated. Microorganisms from different environments (e.g., hydrothermalvents, the deep seafloor, anoxic or oxygen-depleted environments)have been shown to contribute significantly to mineral weatheringthrough modification of the macroscopic structure of several mineralsalso under circumneutral pH conditions (Edwards et al., 2005; Staudigelet al., 2008; Fru et al., 2012). S and Fe bio-oxidation also has importantroles in mineral weathering processes at the seafloor interface zones be-tween anoxic and oxic conditions (Edwards, 2004; Donati and Sand,2007). In the future, these might represent new frontiers for bio-hydrometallurgy, and especially for the extraction of metals from neutraland slightly alkaline ores. However, Sklodowska and Matlakowska(2007) provided evidence that, though 77% to 93% of Cu can be extractedfrom either tailings or ores, a real industrial applicationwould require upto 4 months.

Filamentous fungi, like Aspercillus niger and Penicillium chrysogenum,can produce organic acids (e.g., citric, oxalic, gluconic acids) that favourmetal extractions from solids (i.e., “fungal leaching” or “heterotrophicleaching”; Gadd, 2007, 2010). Metals are solubilised by forming water-soluble complexes (“complexolysis”; Bosecker, 1997; Burgstaller and

Schinner, 1993) or by adsorption with functional groups on the surfaceof the cell wall (e.g., carboxyl, carbonyl, amine, amide, hydroxyl, phos-phate groups; Baldrian, 2003). Moreover, carboxylic acids produced byfungi can attack mineral surfaces and lead to release of the associatedmetals (“acidolysis”; Gadd, 2007). Fungal bioleaching has beenmostly in-vestigated for metal extraction from low-grade ores and mine tailings(Mulligan et al., 2004) and metal-rich industrial waste (e.g., spent refin-ery catalysts; Aung and Ting, 2005; Santhiya and Ting, 2006). Investiga-tion into sediment bioleaching mediated by fungi have provided lowersolubilisation percentages compared to bioleaching with Fe/S oxidisingbacteria (Sabra et al., 2011, 2012), although further investigations areneeded before general conclusions can be drawn here.

2.2. Metal-sulphide oxidation mechanisms

Fe/S oxidising bacteria have been used in industrial-scale processesfor metal extraction from sulphide ores since the early 1950s(Zimmerley et al., 1958). Several investigations have clarified themech-anisms of metal-sulphide dissolution, which can now be considered aswell characterised. In the past, the solubilisation of metal sulphides byFe/S oxidising bacteria was described as a process based on two inde-pendent mechanisms: 1) a “direct mechanism”, in which microorgan-isms would oxidise the metal sulphides by the attack of the mineralsurface, which lead to their dissolution without a soluble electronshuttle; and 2) an “indirect mechanism”, in which the dissolution ofsulphides occurs by the leaching agents produced by Fe/S oxidisers.However further investigations have provided evidence that this indi-rect mechanism represents the sole mechanism of metal-sulphide dis-solution (Donati and Sand, 2007; Fowler and Crundwell, 1998; Sandet al., 2001; Schippers and Sand, 1999; Tributsch, 2001). Dependingon the electronic configuration of the metal sulphides (i.e., acid solubleVS acid insoluble) and the geochemical conditions of the environment(e.g., pH, oxidant availability), different RISCs can accumulate(Schippers, 2004; Schippers and Sand, 1999). The metal sulphidesFeS2 (pyrite), MoS2 (molybdenite) andWS2 (tungestenite) are not sen-sitive to proton attack, due to their electronic configuration. They can besolubilised only by a combination of proton and oxidative attack, ac-cording to the following overall reactions (i.e., “thiosulphate pathway”):

FeS2 þ 6Fe3þ þ 3H2O→S2O32− þ 7Fe2þ þ 6Hþ ð1Þ

S2O32− þ 8Fe3þ þ 5H2O→2SO4

2− þ 8Fe2þ þ 10Hþ ð2Þ

At pH 2, the main product of Eqs. (1) and (2) are sulphate (90%), al-though side products are also formed, as sulphite, sulphate trithionate,tetrathionate and other polythionates can be detected (1%–2%). In par-ticular, tetrathionate forms from thiosulphate oxidation (Eq. (3)). Thisreaction occurs across a wide pH range, due to the different oxidants(Schippers and Jørgensen, 2002; Schippers et al., 1996). Contemporane-ously, thiosulphate decomposes to sulphite and elemental S, although atleast at pH 2, this reaction is very slow; thus, tetrathionate is the mainproduct from thiosulphate reactions. Tetrathionate decomposes to sul-phate (Eq. (4)) and trithionate (Eq. (5))

2S2O32− þ 2Fe3þ þ 5H2O→S4O

462− þ 2Fe2þ ð3Þ

S4O62− þH2O→HS3O

33− þ SO4

2− þHþ ð4Þ

S3O32− þ 1:5O2→S3O

362− ð5Þ

Pentathionate, elemental S and sulphite can also form as sideproducts. As for tetrathionate (Eq. (4)), trithionate is hydrolysed tothiosulphate and sulphate (Eq. (6)) (Wentzien et al., 1994):

S3O62− þH2O→S2O

232− þ SO4

2− þHþ ð6Þ


At circumneutral pH, Fe(III) ions are not in the solution phase, butthiosulphate can still be formed if O2 or MnO2 are present. In particular,it has been suggested that O2 can promote the chemical dissolution ofpyrite as Fe(II)/Fe(III) shuttle if Fe(III) complexing organic substancesreduce the pyrite oxidation rate (Schippers, 2004). Similarly underanoxic conditionsMnO2 can promote the chemical dissolution of pyrite(Schippers, 2004).

Eq. (5) can occur with Fe3+ as an alternative oxidant. This series ofreactions results in the cyclic degradation of thiosulphate to sulphate(Schippers et al., 1996; Schippers et al., 1999).

Metal sulphides, like As2S3 (orpiment), As4S4 (realgar), CuFeS2(chalcopyrite), FeS (troilite), Fe7S8 (pyrrhotite), MnS2 (hauerite), PbS(galena) and ZnS (sphalerite) are acid-soluble and can be solubilisedjust by proton attack (i.e., the “polysulphide pathway”). In this case, sul-phide dissolution occurs by proton attack:

MSþ Hþ→M2þ þH2S ð7Þ

In the presence of Fe(III) ions hydrogen sulphide is oxidisedconcomitant to the proton attack:

H2Sþ Fe3þ→H2S�þ þ Fe2þ ð8Þ

The radical cation H2S*+ dissociates to the radical HS*, anddisulphide formation is favoured (Schippers and Sand, 1999):

H2S�þ þH2O→H3O

þ þHS� ð9Þ

2HS�→HS2− þHþ ð10Þ

Disulphide can be further oxidised to HS* by Fe3+ and then it candimerise to tetrasulphide or reactwithHS* to trisulphide,with reactionssimilar to Eqs. (8) to (10). Chain elongation of polysulphide's (H2Sn) can

Fig. 2. Sediment composition in terms of the physico-chemical phases, wit

proceed by analogous reactions. Polysulphides decompose to rings ofelemental S as S8 (Schippers, 2004):

HS9−→HS− þ S8 ð11Þ

These series of reactions result in the formation of elemental S, withyields of N90%. The side products can include thiosulphate (Schippersand Sand, 1999; Zhang and Millero, 1993). The overall resultingreactions are:

MSþ Fe3þ þHþ→M2þ þ 0:5H2Sn þ Fe2þ n≥2 ð12Þ

0:5H2SnþnFe3þ→0:125S8 þ Fe2þ þ Hþ ð13Þ

Protons are generated in both the “thiosulphate” and “polysulphide”pathways although in the first phases of the polysulfide pathway, pro-tons are consumed and the presence of S-oxidisers is needed to regen-erate these and to stabilise the pH (Schippers, 2004). If acidophilic S-oxidisers are not present during this process, S0 can accumulate inboth the polysulphide and thiosulphate pathways, while variouspolythionates can accumulate in the thiosulphate pathway.

Fe/S oxidisers are often applied as a consortium of more strains, toenhance metal dissolution from minerals. For instance, the S-oxidisingautotroph At. thiooxidans cannot dissolve pyrite in pure culture, al-though but it can improve pyrite dissolution in co-culture with the Fe-oxidisers L. ferrooxidans or Ferrimicrobium acidophilum (Kelly andWood, 2000; Okibe and Johnson, 2004). Acidophilic Fe-oxidisers canalso control the oxidative conditions of the medium by controlling theFe(III)/Fe(II) ratio (Gonzalez-Toril et al., 2003; Rowe et al., 2007). Incontrast, S-oxidisers can have important roles by: (i) removing S-richlayers on the mineral surface that can hinder metal dissolution(Fowler and Crundwell, 1999); and (ii) generating sulphuric acid, thatmaintains the acidic conditions required by the Fe-oxidisers (pH 1÷3).

h examples of reactions and processes that occur among and therein.

Box 1Selective sequential extraction: protocols for metal mobilityassessment.

Selective sequential extraction procedures provide informationabout the metal distribution among the geochemical fractions withdifferent degrees of mobility in environmental matrices, like soilsand sediments. Since metal mobility depends strictly on the metalspeciation, selective sequential extraction protocols also provide in-direct information about metal bio-availability and toxicity (Hlavayet al., 2004; Wang et al., 2010).Up to eight geochemical fractions can be considered (Filgueiraset al., 2002; Hlavay et al., 2004; Monterroso et al., 2007; Ureet al., 2006; Zimmerman and Weindorf, 2010):

(i) The “mobile fraction” (i.e., easily exchangeable), whichincludes the water-soluble, non-specifically adsorbedmetals;

(ii) The “easily mobilisable fraction”, which includes themetal species that are bound specifically or surface oc-cluded (in some cases metals bound to carbonates andmetallo-organic complexes can fall in this fraction);

(iii) The “bound to carbonates fraction” (metals associatedwith highly acid volatile sulphides are recorded in thisfraction);

(iv) The “organic fraction”, which includes metals associatedto functional groups of high molecular weight organiccompounds;

(v) The “bound to Mn-oxide fraction”;(vi) The “bound Fe- and Al-oxide fraction”, which includes

metals bound both amorphous and crystalline oxides;(vii) The “sulphide fraction”, which includes metals within

acid-insoluble sulphides;(viii) The “residual fraction”, which includes metals within

the crystalline lattice of primary and secondaryMinerals

Metals in the residual fraction are ideally very stable and are usuallyconsidered to reflect the natural background level of ametal due togeological and non-anthropogenic causes. On the contrary,metalsin the non-residual fractions (especially those in the exchangeableand carbonates fraction) can be mobilised or solubilised becauseof biotic and abiotic processes (e.g. changes in ionic strength,pH, oxidation/reduction potential), and thus they can becomemore bioavailable and enter the foodweb,with potential detrimen-tal effects on the health of the ecosystem (Akcil et al., 2014;Eggleton and Thomas, 2004; Gleyzes et al., 2002; Toes et al.,2004).One of themost applied selective sequential extraction proceduresis the three-step protocol of the European StandardMeasurementsand Testing Programme (also known as the European Communi-ties Bureau of Reference), which was designed on the basis ofcontributions by Salomons and Förstner (Förstner, 1993;Salomons, 1993). Four sediment macro-fractions are defined:(i) an “acid-soluble fraction”, which includes exchangeable andcarbonate-bound fractions; (ii) a “reducible fraction”, which in-cludes metals associated with Fe/Mn oxides; (iii) an “oxidisablefraction”, which includes metals bound to sediment organic mat-ter and to sulphides; and (iv) the residual fraction (Quevauviller,1998a, 1998b, 2002). Other widely applied procedures are theTessier scheme and its modified versions (Tessier et al., 1979;van Hullebusch et al., 2005), which includes five and four frac-tions, respectively: (i) exchangeable; (ii) carbonatic; (iii) reducible;(iv) organic; and (v) residual; and (i) exchangeable; (ii) carbonatic;(iii) oxidizable; and (iv) residual, respectively.


As there is no “direct mechanism” and as metal solubilisation islargely dependent on the microbial metabolic products, the main roleof microbes in any sediment bioleaching treatment is to establish andmaintain the loop of generation and re-generation of the leachingagents.Moreover, there is noneed to limit the external supply of growthsubstrates, unless this is useful to reduce cost and impact. On the con-trary, a pre-adaptation step of the microbes to the sediment matrixcan improve bioleaching rates (Chen and Lin, 2001a; Hong et al., 2015).

3. Sediment geochemical properties and their effects on bioleaching

3.1. General sediment characteristics

Sediment particles originate through a variety of processes, whichinclude continental run-off, erosion, biological production, depositionof organic particles, weathering of primary minerals, mineral precipita-tion and bio-precipitation. As a consequence, sediments are veryheterogenic environmental matrices, even at very fine spatial scales(Warren and Haack, 2001; Salomons and Brils, 2004). Aquatic sedi-ments can be described as consisting of three types of phases (Fig. 2):(i) the aqueous phase, that is the porewater between the sediment par-ticles (i.e., the interstitial water), where salts, gases and organic mole-cules are dissolved; (ii) several solid phases: these form the solid partof the sedimentmatrix, including themineral and organic components;and (iii) the boundary phases: these are the transitional interfacesamong the phases. Examples of boundary phases are the exchangeableions that reside in the exchange phase of the sediment, or those that areadsorbed onto solid phases (Essington, 2004; Grabowski et al., 2011).The sediment phases are not sealed off from each other, and several re-actions can occur among phases that are regulated both chemically(i.e., changes in the environmental conditions) and biologically, throughthe living component. Almost all of the chemical reactions in the sedi-ment are mediated by (or occur in) the aqueous phase. At the sameway, the behaviour and fate of substances in the water column (includ-ing contaminants) are strongly affected by reactions that occur in thesediment.

In all the aquatic ecosystems, sediments are a major repository forcontaminants. Metals, like Cd, Hg, Zn, Ni, Cr, Pb and Cu, and semi-metals, like As, enter aquatic ecosystems through multiple sources,which include the release of industrial effluents, atmospheric deposi-tion and continental run-off (Perrodin et al., 2012; Salomons and Brils,2004). Due to their high affinity with sediment components(e.g., organic matter, Fe/Al/Mn oxides, sulphides), (semi-)metals canreach concentrations that are much higher than in the water column(Yu et al., 2001; Eggleton and Thomas, 2004; De Jonge et al., 2012).The mineralogical composition, grain size and content of total organicmatter (i.e., TOM) can heavily influence the degree of sediment contam-ination (Tessier et al., 1982; Buckley and Cranston, 1991; Bergamaschiet al., 1997). Muddy sediments are typically characterised by highmetal content due to high surface to volume ratios and high TOM con-tents than for sandy sediments (Karickhoff et al., 1979; Caron, 1989;Zagula and Beltinger, 1993).

Metals and semimetals associate with the solid phases in the sedi-ment through different mechanisms (i.e., particle surface adsorption,ion exchange, complexation with organic substances, co-precipitationand precipitation) and with different level of retention (i.e., mobility).Their behaviour in sediment environments depends on the equilibriaamong the processes that lead to metal immobilisation (i.e., redoxtransformations, precipitation, adsorption and intracellular uptake)and processes leading to metal mobilization (i.e., redox reactions,leaching, volatilisation by methylation and chelation/complexation;Toes et al., 2004; van Hullebusch et al., 2005; Fonti et al., 2015a). Inaddition, as sediment properties can vary from site to site (due todifferences in geological, sedimentological and hydrological, conditions;Preda and Cox, 2005; Perrodin et al., 2012), (semi-)metals are distribut-ed among the various geochemical phases in a site-specific way.


Procedures for selective sequential extraction of (semi-)metals fromsediments can be used to investigate their partitioning among differentgeochemical fractions (Filgueiras et al., 2002; Gleyzes et al., 2002;Hlavay et al., 2004; Ure et al., 2006; Zimmerman and Weindorf, 2010).A brief explanation of this is given in Box 1.

Although selective sequential extraction procedures have intrinsiclimitations (e.g., lack of specificity in some steps), they can be also usefulto provide insights about the mobility of metals in the sediment(Eggleton and Thomas, 2004; Toes et al., 2004; Yuan et al., 2004;Bacon and Davidson, 2008).

3.2. Fate of metals in sediment bioleaching

In sediment bioleaching, the rate and the extent to which a metalsolubilises is closely related to its speciation/partitioning (Chartieret al., 2001; Chen and Lin, 2009; Fang et al., 2011; Sabra et al., 2011).As a consequence, the determination of total metal concentrations isnot sufficient to provide reliable information about metal mobility andremoval potential during sediment bioleaching treatments (Brils,2008; Peng et al., 2009; Wang et al., 2010; Chon et al., 2012). In partic-ular, bioleaching microbes influence equilibria among processes ofmetal immobilisation/mobilization by having a biological control of(i) the operating conditions (i.e., lowering pH to around 2 and increas-ing the oxidation reduction potential) and (ii) the availability of theleaching chemical species (i.e., H+, Fe3+, products of S-oxidation).Low pH and highly oxidative conditions can favour the release of(semi-)metals associated with the sediment organic matter, althoughthe degree of dissociation will be dependent on site-specific properties,like the ratio of fulvic-to- humic acids, the composition in terms of func-tional groups (e.g., carboxyl, phenolic, alcoholic, carbonyl groups), andthe aging state of the sediment organic matter (McBride et al., 1997;Yin et al., 2002; Guo et al., 2006; Tang et al., 2010).

Metals that are weakly absorbed on mineral surface (i.e., the ex-changeable fraction) and those that are bound to carbonates are easilysolubilised under low pH conditions, through ion-exchange reactions(i.e., protonation) and dissolution reactions, respectively (Gleyzeset al., 2002). Metal contaminants in the oxidisable fractions(i.e., organic and sulphide fractions) appear to solubilise only underhighly oxidative conditions (Chartier et al., 2001; Fonti et al., 2013b).

A number of studies have provided evidence that bioleaching can re-sult in the release of (semi-)metals from the reducible fraction (i.e., Feand Mn oxides; Chartier et al., 2001; Chen and Lin, 2004; Fang et al.,2011; Fonti et al., 2013a). This can be explained in term of the low pHvalues, whereby hydrous Fe(III) oxides can partially dissolve at theedge of the lattice structure through protonation, which results in therelease of (semi-)metals from the reducible fraction (Filgueiras et al.,2002). In contrast, Tsai and co-authors (2003a) reported that Mn-oxides become an important binding pool after bioleaching, especiallywhen the S0 available for bioleaching strains is limited).

Metals in the residual fraction are expected to be very stable(Schippers and Jørgensen, 2002; Yuan et al., 2004), although somebioleaching studies have reported metal solubilisation even from thisgeochemical fraction (Chartier et al., 2001; Chen and Lin, 2004; Fanget al., 2011; Sabra et al., 2011; Fonti et al., 2013a, 2013b). This mightbe partially due to intrinsic procedure bias (Usero et al., 1998; Gleyzeset al., 2002; Hlavay et al., 2004; Goh and Lim, 2005), although smallamounts of metals in the residual phase can be released through edgedissolution of some acid-sensitive minerals.

Although there is evidence that the initial partitioning of (semi-)metals does significantly influence the solubilisation of these contami-nants from the sediment during bioleaching, each (semi-)metal be-haves according to its chemistry in the aqueous phase (McBride et al.,1997; Sauvé et al., 2000; Fonti et al., 2013a, 2015b; Chen and Lin,2001a). For example, low pH increases the solubilisation of Zn, Mn, Cuand Pb (for Pb only if no precipitation occurs), while thus has less influ-ence on the release of Cd, Ni and Cr. Zn is known to be one of the most

mobilemetals, and its solubilisation ismore dependent on oxidation/re-duction potential and pH, rather than the presence of bioleachingstrains per se (Escrig and Morell, 1998; Reddy and DeLaune, 2004;Atkinson et al., 2007; Yao et al., 2012; Fonti et al., 2015a). Ni, Cd andCu can be bioleached with high efficacy, but they can also be poorlydissolved (Bae et al., 2001; Blais et al., 2001; Chartier et al., 2001). Cu re-moval is dependent mainly on its partitioning in the sediment. Cd, un-like Pb and Cu, adsorbs to complex organic molecules via outer-spherecomplexation, and thus its solubilisation is greatly affected by theionic strength (e.g., competitionwith Ca2+ ions for complexation of dis-solved organicmatter; Guo et al., 2006; Kinniburgh et al., 1999;McBrideet al., 1997; Sauvé et al., 2000). Pb solubilisation by bioleaching is greatlyaffected by precipitation processes, as it can easily speciate into highlyinsoluble compounds (e.g., PbSO4, Pb5(PO4)3Cl, Pb5(PO4)3OH), even atlow pH. Low solubilisation of Pb is generally observed after bioleachingtreatment of sediments (Chartier et al., 2001; Chen and Lin, 2004; Fontiet al., 2013a; Gan et al., 2015b). The addition of NaCl or FeCl2 to sedi-ments after bioleaching has been proposed to improve Pb solubilisationthrough its complexation with Cl− ions (Chartier et al., 2001; Kim et al.,2005). However, organic compounds and Fe/Mn oxy/hydroxides canform surface coatings on minerals, with the consequent decrease inmetal bioleaching efficiency (McBride et al., 1997; Warren and Haack,2001; Yin et al., 2002).

3.3. Other sediment characteristics to be considered in bioleachingapplications

As multiphasic and heterogenic matrices, sediments are more com-plex than the mineral ores that are commonly treated by bioleaching.Its properties (e.g., mineralogical composition, redox conditions, poros-ity) are highly site specific, and this can affect the effectiveness of the re-mediation technologies (USEPA, 1994; Batley et al., 2005; Vallero, 2010;Fonti et al., 2013b, 2015a). In view of large-scale application (e.g., heapbioleaching, solid-bed bioleaching), several geotechnical properties ofsediments have to be considered, like plasticity, compressibility, un-drained shear strength and sensitivity (Lee, 1982; Schultheiss, 1982).Dredgedmaterials are usually very fine-grained and this can complicatethe engineering operations during sediment treatment. Freshly dredgedsediments are nearly impermeable to water and air, and thus they re-quire a pre-treatment. The conversion of dredged aquatic sediments(as nearly impermeable) into soil-like permeable material occurs spon-taneously when the sediments are stored in the open for years(Vermeulen et al., 2003), or can occur over a few months by plantingsediments with helophytes (Löser et al., 2001; Löser and Zehnsdorf,2002). Reed canary grass (Phalaris arundinacea) has been shown to beparticularly suitable for conditioning sediments (Zehnsdorf et al.,2013). Freshly dredged and anoxic sediments usually have highbuffering capacity, while acidification appears to be easier in dredgedsediments after being exposed in the open for a long time, and thusafter becoming oxidised (Löser et al., 2006c).

With respect to sediment bioleaching treatments, the acid-neutralising capacity (i.e., buffering capacity against acidification) isparticularly relevant in determining (semi-)metal removal percentages.High carbonate contents result in high acid-neutralising capacity butthe organic matter can contribute to neutralising protons: under acidicconditions, the functional groups in sediment organic matter(e.g., carboxyl, phenolic, amino groups) behave as weak acids and con-sumeprotons throughprotonation. Under acidic conditions, aluminosil-icate minerals can contribute at the acid-neutralising capacity ofsediments by consuming protons in mineral weathering reactions, al-though the kinetics are much slower than for carbonate dissolution(Essington, 2004; White et al., 2001). The buffering capacity of sedi-ments can be measured by titration of a sediment suspension with1 M H2SO4, although determination of the carbonate content providesa useful indirect estimation (Seidel et al., 2004; Löser et al., 2005;Löser et al., 2006a). If dredged sediments are characterised by high


carbonate concentrations, this process would require excessiveamounts of leaching agents and might well not be economical. Tsaiet al. (2003a) reported the need for S0 at 360 g/kg sediment to bioleachmetals from a sediment sample with 11.2% carbonate content. In thesameway, the treatment can have high environmental impact, becauseof the production of large amounts of CO2 from the reactions betweencarbonates and sulphuric acid.

Some bioleaching studies have addressed the effect of certain sedi-ment properties, although very few studies have compared sedimentsamples with different geochemical characteristics. Löser et al.(2006a) investigated qualitatively the effects of the acidic neutralisingcapacity on: (i) solubilisation of Zn, Ni, Cu, Cd, Cr, and Pb; (ii) the aciddecreasing rate; and (iii) bio-production of sulphuric acid. They com-pared six sediment samples from different water bodies in Germany(Table 1) and they observed that acidification of the sediments wasmainly affected by their carbonate content, and in a minor way bytheir oxidation state.

The development of efficient sediment bioleaching strategiesrequires a deep understanding of the sediment geochemistry and thesediment-metal interactions (e.g., presence and type ofmetal scavengercomponents, types of reactions in which metals are involved). Never-theless, the geochemical properties of the sediments have often been ig-nored or barely defined in many bioleaching studies.

4. Effects of the main operating variables

The first study of the application of bioleaching as a sediment biore-mediation strategy was published in 1993 (Couillard and Chartier,1993). This study was inspired by a biological process for metalsolubilisation from sewage sludge that had been developed a fewyears before (Couillard and Chartier, 1991; Blais et al., 1992; Couillardand Mercier, 1993). Subsequently, new studies investigated the factorsand the operating conditions (e.g., growth substrates, temperature, sed-iment concentration, bioreactor type). Table 1 provides an overview ofthe most representative studies and their main results. In particular,the large majority of these studies were based on bench-scale experi-mentations (volume, 50–250 mL) or small batch bioreactors (volume,≤5 L). Most of the studies were performed on sediment samples fromfreshwater systems, while marine sediments were investigated in onlya few studies. Zn, Pb, Cu, Cd, Cr, Ni, Mn, As, Co and Hg have been themain target (semi-)metals (in order of decreasing investigation fre-quency. An analysis of the main operating variables that influence sed-iment bioleaching studies, and of the main conclusions obtained by thescientific community is provided in the following paragraphs.

4.1. pH and oxidation/reduction potential.

pH is one of the most important variables that influences (semi-)metal bioleaching from solids: 1) pH determines the rates and extentof metal solubilisation (Sauvé et al., 2000; Chen and Lin, 2001a;Kumar and Nagendran, 2007); 2) low pH is needed for metals to be sta-ble in the solution phase (Blais et al., 1992; Sreekrishnan et al., 1993;Chen and Lin, 2000b, 2001a); and 3) low pH is essential for the activityof themajority of the Fe/S oxidising bacteria and archaea (Tuovinen andKelly, 1973; Johnson, 1998; Schrenk et al., 1998).

As low pH is needed both for the activity of the bioleaching strainsand for the (semi-)metal stability in the solution phase, different strat-egies have been applied to favour pH lowering, for example: 1) a stageof microbial strain pre-adaptation, 2) a pre-acidification stage of thesediment slurry (to neutralise acid-consuming substances), and 3) theuse of key substrates (like S0 and Fe2+) and optimisation of their con-centrations as a function of the sediment concentration and of the sed-iment buffering capacity. However, although pH is an important factor,the relationship between pH and (semi-)metal solubilisation from sed-iments is not linear, and not even the same among the (semi-)metals

(Chen and Lin, 2001a), as other factors can affect (semi-)metalsolubilisation in an element-specific way.

High oxidation/reduction potential favours solubilisation of(semi-)metals associated with the oxidisable fraction, especiallywith non-acid soluble sulphides. The oxidation/reduction potentialduring bioleaching is not significantly affected by the amount of S0 pro-vided (Tsai et al., 2003a), while the presence of Fe2+, pyrite or other Fesources can result in very high oxidative conditions due to bio-oxidation by Fe-oxidising strains (Eh values even up to +700 mV ormore).

4.2. Growth substrates

Due to the nature of bioleaching, growth substrate selection and op-timisation of their concentrations and application methods are hotpoints in planning bioleaching treatments. Growth substrates are cho-sen according to the metabolism of the microbial strains to be used.RISCs and Fe2+ are key substrates, because they are both energy sourcesfor Fe/S oxidising strains and precursors of key leaching species (i.e., H+

and Fe3+; Rohwerder et al., 2003; Fonti et al., 2013a). In the scientificliterature, S0 and Fe sulphate have been the most applied growth sub-strata, although organic substances (e.g., glucose, sucrose, yeast extract)can also be addedwhen heterotrophs ormixotrophs are used (Beolchiniet al., 2009, 2012, 2013; Fonti et al., 2013a).

4.2.1. Elemental sulphurS0 has been the most used substratum in investigations into sedi-

ment bioleaching (Table 1). S0 is added alsowhen the bioleaching strat-egy is based on biostimulation of indigenous microbes in the sediment,instead of augmentation with allochthonous bacteria (Chen and Lin,2001a; Tsai et al., 2003b; Seidel et al., 2004; Seidel et al., 2006b).

During batch reactor experiments with S-oxidising strains(e.g., T. thioparus, At. thiooxidans) using 1% and 2% (w/v) sediment con-centrations, Chen and Lin showed that the optimal S0 concentrationsranged from 3 g/L to 5 g/L, respectively (Chen and Lin, 2001b, 2004).Fang and co-authors used 3 g/L S0 in suspension systems with the sed-iment at 10/to 12% (Fang et al., 2009a, 2009b; Fang et al., 2013).Beolchini et al. (2009) used 5 g/L S0 in their 2% marine sedimentbioleaching. For sediment treatment in a solid-bed bioreactor, theconcentration of S0 has been shown to be a function of the amount ofsediment to be treated (e.g., 20 g S0/Kg sediment; Löser et al., 2001;Seidel et al., 2004, Seidel et al., 1998).

There is evidence that in sediment bioleaching the rate of acidifica-tion, and the rate of sulphate bio-production, increase with the amountof available S0 (Tichy et al., 1998; Chen and Lin, 2001a, 2001b; Tsai et al.,2003a, 2003b; Löser et al., 2005). However, Löser and co-authors (2005)reported that too high amounts of S0 can result in solubilisation of someof theminerals in sediments, likeMgO, CaO, Al2O3. In contrast, Chen andLin (2001b) reported that a S0 concentration higher than 5 g/L can neg-atively affect the rate of metal solubilisation. Differentmetals are affect-ed differently by the S0 concentration of: while Zn appears not to besignificantly affected by different S0 concentrations in the rangebetween 0.5 g/L and 7.0 g/L, Cu and Cr solubilisation are positively influ-enced at up to 3 g/L and 5 g/L S0, respectively (Fang et al., 2009a).

The main S source used in bioleaching investigations is commercialorthorhombic S0 powder, although it is known to have disadvantagesthat are mainly related to its very low solubility. This results in residuesthat are difficult to remove from the sediment after treatment, and it cancause sediment re-acidification. Alternatively, S pastilles can be used,with the additional advantage that they can be re-used easily for furthersediment bioleaching treatments (Chen et al., 2003a). Due to its hydro-philic properties, waste biogenic S0 that is produced by Acidithiobacillusstrains inwet desulphurisation plants for the purification of waste gases(e.g., in paper mills) also represents a valid substrate for bioleachingprocesses, and its low cost favour its application to bioleaching (Chen


et al., 2003a; Seidel et al., 2006b; Fang et al., 2009b; Fang et al., 2013;Ilyas et al., 2014).

4.2.2. ThiosulphateTwo studies have tested the use of thiosulphate as a S source during

bioleaching of contaminated sediments from freshwater systems (Tsaiet al., 2003a, 2003b).These studies reported rapid rates of sulphate pro-duction and pHdecrease, alongwith high removal of Ni, Zn and Cu. Con-versely, lower solubilisation was observed for Cr, Co and Pb. No otherstudies have tested the use of thiosulphate as a substrate for bioleachingof contaminated sediments, very likely because S0 is a cheaper S source,even in its commercial form. In addition, rates of thiosulphate bio-oxidation are slower than those of S0 during bioleaching processes ap-plied onto sewage sludge (Tyagi et al., 1994). Nevertheless, the use ofdifferent types of S sources (correctly based on theworking parameters,sediment properties and microbial consortia used) might represent acontrol strategy for chemical and biological processes that are triggeredduring sediment bioleaching treatments.

4.2.3. Ferrous ionThe available information indicates that Fe(II) is less used compared

to S0 for bioleaching of contaminated sediments. Fe2+ is provided onlywhen Fe-oxidisers are used to bioleach metals. For the biostimulationof indigenous bioleaching strains, Fe is not usually provided, becauseaquatic sediments are circumneutral or slightly basic, and a rapid abioticoxidation of Fe(II) occurs. Fe is usually added as FeSO4 at concentrationsranging from 0.3 g/L to 16 g/L (Table 1), although in some cases Fe hadalso been used as FeCl2 (Kim et al., 2005).

When bioleaching strategies are applied to sediments that are rich inacid-consuming substances (like those from seaports) S0 can havereduced effects on metal solubilisation, while Fe2+ is a more suitablegrowth substrate. Also, after bio-oxidation, Fe2+ is a key element inthe acidification of the experimental environment, especially in thepresence of high content of sediment (Beolchini et al., 2012; Fontiet al., 2013b). On the other hand, high concentrations of Fe cancomplicate further steps after bioleaching, because of the formation ofFe co-precipitates, which are seen mainly as jarosites (Fonti et al.,2013b).

4.2.4. PyritePyrite (FeS2) dissolution occurs by oxidative attack and under acid

conditions. This leads to the release of Fe2+ and RISCs (the thiosulphatepathway), and thus it represents a potential growth substrate in metalbioleaching from solid matrices (Bas et al., 2013). At the same time, ithas been recently reported that pyrite addition can improve the acidifi-cation potential of sediments during bioleaching treatments in stirrertank reactors (Gan et al., 2015b). Since S0 is known to favour acidifica-tion in the early stages of sediment bioleaching, Gan et al. (2015b)hypothesised that the addition of S and pyrite in equal amount can pro-vide more efficient acidification.

4.2.5. Choice of growth substrateThere have been few studies that have compared the effects of S0

and Fe2+ as the key substrates for bioleaching of contaminated sedi-ments, and contrasting results have been reported. In the bioleachingof river sediment samples, the highest removal percentages of Mn, Cu,Cd and Zn were obtained using 10 g S0 per kg dry sediment (Sabraet al., 2011). In contrast, in the bioleaching of marine sediments, thehighest metal solubilisation percentages were obtained using Fe(II)rather than S0 (Beolchini et al., 2012, 2013; Fonti et al., 2013a, 2013b).Similarly, Chen et al. (2003b) observed that the addition of 1 g/LFe(III) can significantly improve the solubilisation of Pb and Cr inbioleaching based on S-oxidising bacteria (i.e., At. thiooxidans,T. thioparus). Unfortunately, the choice of the best growth substratesand the optimisation of their concentration are not only function ofthe microbial strains used in a bioleaching treatment, as it depends on

other factors, like sediment concentration and its content of acid-consuming substances (e.g., organicmatter, carbonates and acid solubleminerals). Further investigations into growth-substrate use are needed,particularly as excess S or Femight compromise the re-use of the treatedsediment. On the other hand, substrate concentrations should be set tomaximise the metal bioleaching performance.

4.3. Sediment concentration

In sediment bioleaching, the rate of acidification does not only de-pend on the rate of S0 bio-oxidation, as it also decreases with increasingsediment concentrations (Chen and Lin, 2000a, 2001a; Tsai et al.,2003a). Sulphuric acid, ferric ions and the other products of Fe/S oxidis-er metabolism are consumed by the sediment components. Moreover,the acid-neutralising capacity of sediments opposes the metalsolubilisation and might favour metal precipitation after solubilisation(Fonti et al., 2013b). On the other hand, sediments with high acid-neutralising capacity can have inhibitory effects on Fe/S oxidiser activi-ty, even at low concentrations (Bae et al., 2001; Beolchini et al., 2012).Consequently, the solid to liquid ratio (i.e., the concentration of sedi-ment) is a very important factor in real applications, and it need to bevaried according to the geochemical properties of the sediments. Dueto these problems, sediment concentrations are often relatively low inbioleaching studies (often b30 g/L) and this factor has been poorlyinvestigated as a dependent variable.

4.4. Temperature

The effects of temperature on bioleaching have been widely investi-gated in the field of metal recovery from low-grade ores and mineralconcentrates. In general, metal bioleaching is more rapid at higher tem-peratures, due to both an increase in the kinetic rates and to the fastermetabolism of S-oxidising microbes (with the consequent rapid de-crease in pH). However, as most microorganisms used in bioleachingare mesophilic bacteria, temperatures that are too high can lead toinhibition. The highest leaching activity in the presence of mesophilicS-oxidising strains is around 30 °C to 35 °C, while at 50 °C almostcomplete inhibition is observed (Rawlings and Silver, 1995; Bosecker,1997; Krebs et al., 1997; Rawlings, 2005). Some other studies haveconfirmed these findings also for bioleaching of contaminated sedi-ments (Tsai et al., 2003b). However, in large-scale applications(e.g., solid-bed bioleaching), the heat that is generated by the exother-mic reactions can accumulate, which results in increase of tempera-tures. The greatest temperature fluctuations appear to occur in thefirst days of treatment, and these cause delays in S oxidation (Löseret al., 2005, 2006c; Seidel et al., 2004).

4.5. Oxygen

Bioleaching processes require oxygen. In the presence of sediments,the oxidation of S0 requires higher levels of oxygen (Blais et al., 2001;Seidel et al., 2004). Reduced oxygen availability results in a delay (ortemporal suppression) in acidification, sulphate production and metalsolubilisation, although slow rates of S0 and Fe2+ oxidisation occureven under strong oxygen limitation (Rawlings, 2005; Löser et al.,2006b; Seidel et al., 2006a). The O2 consumption can also increase,especially in the first stages of a bioleaching process with contaminatedsediments that have not undergone a pre-acidification stage. In thiscase, organic compounds are both chemically degraded and used by in-digenous heterotrophic microorganisms (Fonti et al., 2013a; Seidelet al., 2004). Thus, the quantity and bio-availability of sediment organicmatter have important roles in determining the total amount of oxygenconsumed during bioleaching treatments.


4.6. Considerations on experimental procedures

A comparison of the data obtained across different studies is difficult,as the analytical techniques and experimental conditions are often toodifferent. Sediment pre-treatment stages represent an additional im-portant source of variability. Pre-treatments can include: (i) pre-washing with water remove salts (particularly for marine sediments);(ii) autoclaving; and (iii) conditioning pre-treatments to improve thepermeability of sediments to oxygen and water (Beolchini et al., 2013;Fonti et al., 2013a; Löser et al., 2001; Seidel et al., 2004). However,while sediment washingwith water does not result in significant varia-tions in the properties of sediments while simulating large scale sedi-ment treatments (Beolchini et al., 2012, 2013; Fonti et al., 2013a), theautoclaving of sediment should be considered as a mild thermal-pre-treatment. This can greatly influence the nature of themetal contamina-tion and the geochemical sediment properties, and it thus does notallow comparisons of data obtained across different studies.

Some studies have not included control treatments in the experi-mental design, or the controls have been set up under experimentalconditions that are different to the corresponding experimental treat-ments. The initial pH in controls has often been higher than in the cor-responding experimental treatments. In this way the variance due tothe bioleaching processes has not be assessed correctly. The metal ex-traction performances in a bioleaching treatment (i.e., Me%) are thesum of the chemical solubilisation processes (dependant on initial pH,sediment properties, nature of contamination;Me%chemical) and the bio-logically mediated leaching (dependant on environmental conditionsestablished and biogeochemical cycles powered by Fe/S oxidisingstrains; Me%biological); i.e.:

Me% ¼ Me%chemical þMe%bio logical ð14Þ

As well as the inclusion of control treatments in statistical analysesof the data obtained, control treatments are needed to define the

Fig. 3. Conceptual model of the main biogeochemical processes that are involved in sedibetween acid-consuming reactions and the availability of key chemical species (e.g., H+ and Fefractions of the sediment are involved in the release of metals with different ways and differendergo further processes, like re-adsorption, (co-)precipitation and complexation with soluble l

biological effects of bioleaching treatments and to obtain a better under-standing of the chemical and biological mechanisms involved.

An additional source of experimental error is the heterogeneous dis-tribution of metals in the sediment (even at very fine spatial scale),which cannot be completely avoided by sample homogenisation(Fonti et al., 2013a). Thus, in order to provide robust estimates of thereal extents of metal solubilisation during bioleaching experiments,the metal extraction performances should be calculated according toEq. (15):

Me% ¼ MesolutionMesolution þ Mesediment �W=Volð Þ ð15Þ

where Mesolution is the concentration of the metal in the solution phase(μg/L), Mesediment is the concentration of the metal that remains in thesediment (μg/g), Vol is the experimental volume used (L), and W isthe amount of sediment used (g, dry weight). This implies that metalsshould be measured both in the soluble phase and in the residue fromthe sediment, at least at the beginning and at the end of the experiment.

The development of a sediment bioleaching process should includestrategies for the maintenance of low pH conditions through: (i) pre-acidification by acid addition; (ii) bioaugmentation with S-oxidisingstrains that are active at almost neutral pH (e.g., Thiobacillus thioparus,see also Table 1); (iii) biostimulation of indigenous S-oxidisers in thesediment; (iv) addition of Fe3+ ions; and (v) bio-augmentation withFe-oxidising strains using Fe2+ as an energy source, to stimulate thebio-production of Fe(III) ions.

5. Main drivers influencing metal bioleaching from sediment

As discussed above, in sediment bioleaching the dissolution ofmetalsulphides as soluble sulphates is not the sole process that leads to metalsolubilisation. The operating conditions and the bioleaching microbesneed to be defined to favour (i) metal desorption from high-

ment bioleaching. The stability of metals in the solution phase depends on the balance3+). Fe/S oxidising microorganisms have a biological control on this balance. Geochemicalt amount. Once that a metal is released from its major scavenging components, it can un-igands. Modified from Fonti et al. (2013a).


molecular-weight organic molecules, (ii) complexation with dissolvedorganic particles, and (iii) metal solubilisation by protonation and ionexchange (Sauvé et al., 2000; Warren and Haack, 2001; Jain, 2004;Monterroso et al., 2007; Prica et al., 2010). According to our analysis,the development and implementation of bioleaching strategies for ex-situ remediation of contaminated dredged sediments need to take intoaccountmanymore variables than previously considered. The potentialof (semi-)metal removal depends on the possibility of establishing con-ditions that promote their chemical stability in the solution phase. Sev-eral factors and variables have important roles in this context: (i) thegeochemical properties of the sediment; (ii) the solid-to-liquid ratio;(iii) the type of metal contaminants; (iv) the partitioning of the ofmetal contaminants; and (v) the presence/concentration of key growthsubstrates. These can all greatly influence the effectiveness of sedimentbioleaching treatments.

We can state that the limiting step is very likely to be the bio-production rate of H+ and Fe3+ by the Fe/S oxidising strains, due tothe consumption of the leaching agents by a variety of sediment compo-nents (see Section 3). However, the amount of solubilisation of the(semi-)metals is highly element specific and speciation specific. Thegeochemical properties of the sediments also control the availability ofthe leaching agents, the pH and oxidation/reduction potential, whichare responsible for the stability of the dissolved (semi-)metal speciesin the solution phase. It follows then that these biological, geochemicaland chemical factors that influence bioleaching can also affect eachother, with cascade-like effects expected (Fig. 3).

According to our analysis, bacteria adaptation to sediments and sed-iment acid-neutralising capacity represent key factors that can compro-mise bioleaching feasibility in the treatment of contaminated dredgedsediments. Among the metals, Zn, Mn, Cu, Cd and Ni appear to be themore sensitive to bioleaching treatments, although wide fluctuationsdue to their speciation need to be considered.

Further steps towards the development and implementation ofbioleaching processes specific for dredged aquatic sediments shouldfocus on kinetic analysis of the electrochemical aspects of the microbio-logical, chemical and geo-chemical sub-processes defined here. Thiswould lead to the identification of the factors and variables that are in-volved in the control of the rate of metal leaching according to differentprocess configurations and at different stages during the operations. Atpresent, no studies have addressed this point. Important insights hascome from kinetic studies and conceptual mathematical models in thebiomining field (Ojumu et al., 2006; Vargas et al., 2014), althoughthese are simplified for specific categories of minerals (e.g., metalsulphides).

6. Feasibility and sustainability

Although bioleaching is considered to be a promising environmentaltechnology for aquatic sediment reclamation, studies on its applicationhave focused on the identification of the best operating conditions inlaboratory-scale studies (e.g., substrata concentration, substratasubministration, microbial consortium, temperature). Feasibility andsustainability have hardly been investigated to date. Low energy andchemical consumption (compared to other technologies) have beenhypothesised, but not yet demonstrated. Nonetheless, real scale appli-cations imply serious engineering efforts, which will also depend onthe reactor configuration (e.g., tank, heap, dump leaching). Moreover,good performances in the laboratory do not guarantee similar effectsfor large-scale applications, as inhomogeneity and parameter fluctua-tions or gradients can result in loss of efficiency.

To date, the studies by Seidel, Löser and co-authors on dredged sed-iments from Weisse Elster River (Germany), represent the main pilotscale investigation in this field (Löser et al., 2001, 2007; Seidel et al.,2003, 2004; Löser et al., 2006a; Seidel et al., 2006a; Seidel et al.,2006b). In particular, they investigated the performances of a sedimentbioleaching treatment in a solid-bed bioreactor (2000L), where the

sediment was percolated with sulphuric acid generated by biostimula-tion of indigenous S-oxidising strains with S0 (at 20 g/kg sediment).High Zn, Cd and Ni solubilisation was obtained (60%–90%), and thesolid-bed process was demonstrated to treat high amounts of sedimentin relatively small volumes compared to stirred tank bioreactors(i.e., 0.5 kg/L vs. 0.03 kg/L in Blais et al., 2001). To be treated in a solid-bed reactor, the dredged materials need pre-conditioning to incraesetheir permeability to water and air (i.e., conversion into soil-like mate-rial by storing the sediment in the open for some years, or by plantingthe sediment with helophytes; Löser and Zehnsdorf, 2002; Vermeulenet al., 2003; Zehnsdorf et al., 2013). According to their simulations, a de-contamination plant that treats 10,000 ton/y of sediment would pro-duce soil-like material that is potentially re-usable, at a cost of €37 to€110 per ton (Seidel et al., 2004). However, their cost estimations ex-cluded the dredging, transport and conditioning costs. Although solid-bed plants would simplify the further steps after bioleaching(e.g., solid-liquid separation) and although this appears to be the mostsuitable solution, comparisons with other reactor configurations arenot possible at present because the other pilot-scale investigationshave not considered feasibility and sustainability analyses. New pilotscale studies should be performed to this end.

7. Conclusions

Our analysis suggests that several points need to be carefully takeninto account before claiming that any bioleaching is an ecologicallycompatible and sustainable environmental biotechnology for the recla-mation of contaminated dredged sediments for beneficial reuse. One ofthe main limitations is that our knowledge in this field focuses mainlyon the investigation of factors and parameters that influence metal-removal performances. Sediment properties are often neglected andstudies are often based on trial-and-error approaches, where the mainchemical and biological mechanisms are viewed as “black boxes” andare not considered in the experimental planning phases. The develop-ment of a sediment bioleaching treatment that is integrated into asound management strategy needs to take into account manymore as-pects than previously expected. In particular, a comprehensive and inte-grated understanding is need in terms of: (i) the geochemicalcharacteristics of the sediment; (ii) the type, concentration andpartitioning of the metal contaminants; (iii) chemical and biologicalmechanisms involved; and (iii) the potential of appropriate microbialstrains/consortia. This would thus represent the first tool in the devel-opment of a true sediment management strategy.

According to our analysis, remediation or management options fordredged aquatic sediments other than bioleaching need to be consid-ered when: (i) the sediment is very rich in carbonates or has a highacid-neutralising power; (ii) Pb and Cr are the main contaminants;(iii) the metal contaminants in general show low solubilisation(assessed by bioleaching pilot tests); and (iv) the parameter controlstrategies require additions of high amounts of S0, Fe or acids.

Additional research efforts are needed before bioleaching will be-come an effective technology for sediment-reclamation purposes. Onlya few acidophilic microorganism strains have been tested, which repre-sent those that are mainly used in the mining industry. Moreover, mi-croorganisms from several environments (e.g., hydrothermal vents,the deep seafloor, anoxic or oxygen-depleted environments) havebeen shown to contribute significantly to mineral weathering, andexploring their metabolic pathways might open up new ways for theimprovement of bioleaching performances in contaminated aquaticsediments. On the other hand, further investigations should focus onkinetic and thermodynamic studies, where metal speciation andleaching agent bio-production (including chemical consumption bythe sediment components) can be used to predict the final metal solu-bility equilibria. This would facilitate the decision-making processesfor the selection of bioleaching as a suitable strategy for the manage-ment of dredged contaminated sediments. Finally, the feasibility and


sustainability of sediment bioleaching have not been demonstrated yet,and there remains the need for scale-up investigations, with estima-tions of cost and environmental impact. These represent crucial aspectsin the development of truly sustainable and ecologically compatibleprocesses of bioleaching of contaminated aquatic sediments.

Acknowledgements

This researchwas supportedfinancially byUniversità Politecnica delleMarche (Ancona, Italy)within the research project entitled: “Strategie in-novative di recupero e valorizzazione di sedimenti marini” (translatedtitle: “Innovative strategies for marine sediment remediation”).

Appendix A. Supplementary data

Supplementary data to this article can be found online at http://dx.doi.org/10.1016/j.scitotenv.2016.04.094.

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