Biosorption of Lanthanides Actinides and Related Materials

Biosorbents forMetal Ions

Edited byDR JOHN WASE

School of Chemical Engineering, University of Birmingham, UK

and

DR CHRISTOPHER FORSTER

School of Civil Engineering, University of Birmingham, UK

UK Taylor & Francis Ltd, 1 Gunpowder Square, London EC4A 3DEUSA Taylor & Francis Inc., 1900 Frost Road, Suite 101, Bristol, PA 19007

This edition published in the Taylor & Francis e-Library, 2003.

Copyright © Taylor & Francis Ltd 1997All rights reserved. No part of this publication may be reproduced, stored in a retrievalsystem, or transmitted, in any form or by any means, electronic, electrostatic, magnetictape, mechanical, photocopying, recording or otherwise, without the prior permission ofthe copyright owner.

British Library Cataloguing in Publication Data

A catalogue record for this book is available from the British Library.

ISBN 0-203-48304-9 Master e-book ISBN

ISBN 0-203-79128-2 (Adobe eReader Format)ISBN 0 7484 0431 7 (Print Edition)

Library of Congress Cataloging Publication Data are available

Cover design by Jim Wilkie

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Contents

List of Contributors ix

1 Biosorption of heavy metals: an introduction (C.F.Forster; D.A.J.Wase) 1Introduction 1Toxic metals 2Control 4Treatment 5References 9

2 The use of algae as metal biosorbents (G.W.Garnham) 11Introduction 11Biosorption by algae and the mechanisms involved 12Factors affecting the biosorption of metals by algae 20Production and cost of algal biomass for metal removal 23Immobilised algae and derived products 26Algal biosorption processes and engineering considerations 27Commercial algal biosorption 30References 33

3 General bacterial sorption processes (M.M.Urrutia) 39Introduction 39Bacterial surface 39Biofilms 43Charge of bacterial cell surfaces 43Sorption of metal cations and mechanisms 46Sorption of metal anions and mechanisms 51Binding constants 52Modelling 57Applications in biotechnology 58Summary 59References 59

Contents

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4 Fungi as biosorbents (A.Kapoor; T.Viraraghavan) 67Introduction 67Modes of metal ion uptake 67Modelling of biosorption 68Biosorption by living cells 69Biosorption of metal ions by non-living cells 73Regeneration of fungal biomass and elution of biosorbed metals 77Use of immobilised fungal biomass in biosorption 78Biosorption mechanism 78General considerations in the use of fungi as biosorbents 79References 80

5 Biosorption of lanthanides, actinides and related materials (M.Tsezos) 87Introduction 87The mechanism of biosorption/bioaccumulation 89The lanthanides and actinides 96Application of biosorption 97Uranium biosorption 97Thorium biosorption 106Radium biosorption 106Closing comments 109References 110

6 Scavenging trace concentrations of metals (C.J.Banks) 115Introduction 115Coincidental sorption systems 116Biosorption systems specifically for metal removal 121Exposure of biosorbent surfaces to metal-laden wastewaters 123Immobilisation matrices 128Comparisons of reactor designs 136References 136

7 Low-cost biosorbents: batch processes (D.A.J.Wase; C.F.Forster; Y.S.Ho) 141Introduction 141Peat 141Other biosorbents 146Novel activated carbons 147Copper 148Nickel and lead 148Chromium 149Zinc 151Manganese 153Cobalt and cadmium 153Competitive adsorption 154Practical aspects of using peat 155References 158

Contents

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8 Biosorption using unusual biomasses (R.G.J.Edyvean; C.J.Williams;M.M.Wilson; D.Aderhold) 165Introduction 165Types of biomass 166Performance 170Factors affecting adsorption 171Industrial scale systems 177Conclusions 178References 179

9 Low-cost adsorbents in continuous processes (G.McKay; S.J.Allen) 183Introduction 183Peat, lignite and chitosan as sorbents for metal ions 188Sorption column design 195Regeneration and metal recovery 216References 217

10 Biosorption: the future (C.F.Forster; D.A.J.Wase) 221Introduction 221Algal biosorption 222Fungal biosorption 222Bio-wastes 223The future 225Conclusions 226References 227

Index 229

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5

Biosorption of Lanthanides, Actinides

and Related Materials

M.TSEZOS

Introduction

Background

Over the past two decades an increased interest in the phenomenon of metal ionssequestering by living or inactive microbial biomass has been seen in the scientificand engineering community. This phenomenon has potential application inenvironmental pollution control as a biochemical process on which correspondingunit operations can be designed and operated by industry.

Previous publications on the subject have proposed the adoption of two differentterms for the description of the two mechanistically different types of metalssequestering by microorganisms. The term ‘bioaccumulation’ has been proposed forthe sequestering of metal ions by metabolically mediated processes (livingmicroorganisms), and the term ‘biosorption’ for the sequestering by non-metabolically mediated process (inactive microorganisms) (Diels et al., 1995). As ourunderstanding of the above processes has increased, the mechanistic differencesbetween biosorption and bioaccumulation have proved to be so significant that theuse of the two terms has become a necessity (Tsezos and Volesky, 1982a, 1982b;Macaskie and Dean, 1984; Diels, 1989; Diels et al., 1995). The two processes can co-exist and can also function independently as, for example, in the case where aconsortium of microorganisms is exposed to metal-bearing solutions. Literature onboth biosorption and bioaccumulation is extensive, including, for example, work on: • the use of Alcaligenes eutrophus strains in bioreactors for the bioaccumulation of

Cd, Zn and other heavy metals and radionuclides (Diels, 1989)

• the use of Citrobacter species in the bioaccumulation of heavy metals (Macaskie,1991; Macaskie and Dean, 1984)

• the use of Methylobacillus species for uranium biosorption (Glombitza et al.,1984) and of other bacterial species for silver biosorption (Pumpel and Schinner,1986)

• uranium, thorium and radium biosorption from mine waters (Tsezos andMcCready, 1989).

Biosorbents for Metal lons

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This chapter will focus more on the phenomenon of biosorption, and, in particular, onthe removal of lanthanides, actinides and related elements.

Biosorption of metals is generally characterised by high selectivity as compared toion-exchange resins or other adsorbents. This selectivity is considered to be adesirable feature in designing processes for pollution control and/or metal valuerecovery (Tsezos and Volesky, 1982b; Tsezos, 1985; Tsezos and McCready, 1991;Diels, 1989; Diels et al., 1995; Macaskie, 1991; Glombitza et al., 1984; Pumpel andSchinner, 1986; Gadd, 1992). In addition to selectivity, biosorptive processes have thefollowing advantages: • solution toxicity does not inhibit microbial biosorptive uptake

• microbial biomass growth requirements need not be met

• culture purity maintenance is not a concern. Biosorptive processes are excellent candidates for use for the recovery of metalvalues from dilute industrial complex aqueous solutions, the extraction ofradionuclides, e.g. uranium, thorium or radium from mine leachates, and similarmetal value recovery or water pollution control applications (Macaskie, 1991;Pumpel and Schinner, 1986; Diels et al., 1995; Tsezos and Volesky, 1982b; Tsezos,1990; Gadd, 1992).

Technological considerations

The engineering applications of biosorption or bioaccumulation commonly involve adilute complex ionic matrix and large volumes of aqueous process or waste solutionsfrom which the selective extraction and, occasionally, recovery of targeted elementsvia the use of the microbial biomass is intended. Regardless of the detailedengineering configuration of such a process, a stage which significantly affects theoverall efficiency and the economics of the technology is the separation of themicrobial biomass from the waste or process waters following contact (SENESConsultants, 1985).

As a result of this constraint, contact systems making use of microbial biomassimmobilised on a support medium have been developed and proposed for use. Twogenerically different types of immobilised biomass contact systems have beenproposed. The first type is based on the use of immobilised biomass particles whichare produced via the use of a wide range of biomass binding agents, such as syntheticpolymers (e.g. polysulphones), natural polymers (e.g. alginates) or chemical biomasstreatment (Brierley et al., 1986; Kiff and Little, 1986; Tobin et al., 1994; Tsezos andDeutschmann, 1990). The second type is based on the use of microbial biomass films,immobilised on support media such as membrane sheets, disks or inorganic particles(Diels, 1989; Diels et al., 1996–1999; Harel et al., 1995; Tobin et al., 1994; Brierleyand Vance, 1988; Darnall et al., 1989). Each one of the two types of immobilisedbiomass necessitates the implementation of different contact reactor design, such asupflow or downflow packed-bed reactors, rotating biological contactors, membranesheet or tubular reactors, etc. Figure 5.1 shows a typical example of an immobilisedbiomass particle of the first type in two different magnifications.

It is interesting to note the highly porous structure of the particles shown in Figure

Lanthanides, Actinides and Related Materials

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5.1 which is required in order to facilitate and improve the kinetics of metal ionsdiffusion into the inner particle active biosorption sites (Tsezos et al., 1988; Tsezosand Deutschmann, 1990).

The mechanism of biosorption/bioaccumulation

Although a large volume of work has been published and reported on theassessment of the uptake capacities of several microbial biomass types for a variety

Figure 5.1 Electron micrographs (TEM) of immobilised biomass particles: (a) general view;(b) magnification of the particle porous structure


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of metallic elements, systematic effort to elucidate the underlying mechanisms hasbeen limited. The way in which elements bind or are retained by specific microbialbiomass species is understood in detail only for limited combinations of biomass/metal ion pairs.

The mechanistic understanding of biosorption is considered essential in order tooptimise the process application potential of biosorption. More specifically, thisunderstanding is essential in order to exploit optimally the selectivity and efficiencyof the process and to overcome ionic competition and interference effects by otherionic species which exist along with the targeted element in the ionic matrix of thecontact solution (Tsezos and Volesky, 1982a, 1982b; Tsezos et al., 1995, 1996a;Georgousis, 1990; Huang et al., 1991; Avery and Tobin, 1993; Beveridge and Murray,1980).

Biosorptive uptake sites can be intracellular or extracellular and are microbialspecies and element dependent (Tsezos and Volesky, 1982a, 1982b; Tsezos et al.,1996b; Avery and Tobin, 1992; Lovley et al., 1991; Lovley and Phillips, 1992; Tolleyet al., 1991). Reported mechanisms of biosorption are briefly presented below,illustrating the wide variety of physical-chemical phenomena which are involvedduring biosorptive uptake.

The biosorption of uranium by R. arrhizus takes place inside the mycellial cellwall. Retained uranium is taken up via three independent but interrelated processes(Tsezos and Volesky, 1982b). The first process involves the coordination of uranylions by the mycellial cell wall chitin nitrogen. The second process involves thephysical adsorption of uranyl ions within the chitin three-dimensional network. Thethird process involves the hydrolysis of the uranyl ion-chitin complex and theprecipitation of additional uranium hydrolysis species within the cell wall chitinnetwork. Figure 5.2 shows typical transmission electron micrographs of the R.arrhizus mycellial cell wall before and after contact with uranium. The electron-denseareas on the post-contact micrograph are the uranium-bearing zones.

The mechanism of thorium biosorption by the same organism is different (Tsezosand Volesky, 1982a). Thorium is retained primarily by adsorption on the externalsurface of the mycellial cell wall. Chitin involvement in thorium biosorption is ofsubstantially reduced significance as compared to its role during uranium biosorption.Figure 5.3 shows typical transmission electron micrographs of R. arrhizus cells afterthorium biosorption. The electron-dense areas on the outer cell wall are the thorium-bearing zones.

The biosorption of strontium by inactive yeast cells (S. cerevisiae) has beenreported to be primarily an electrostatic attraction of the Sr2+ by the yeast cells,while living cells sequester Sr2+ by a more complex mechanism involving ionexchange with strontium residing primarily within the cell vacuoles (Avery andTobin, 1992).

Work involving the use of EXAFS and XANES techniques reported on thebiosorption of Au by the algal biomass of C. vulgaris has demonstrated the binding ofgold to be primarily the result of ligand exchange reactions leading to the formationof bonds between Au(I) and sulphur/nitrogen sites contained within the algae cells(Watkins et al., 1987).

A combination of biosorption equilibrium and electron microscopy studies on thebiosorption of metals by bacterial species has been reported recently (Tsezos et al.,1995, 1996a, 1996b). In this work, the biosorption loci of Arthrobacter spp.,Alcaligenes spp. and Pseudomonas spp., selected for their high biosorptive uptake


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capacities, were examined using EM and EDAX microprobe analysis. It was reportedthat the locus of biosorption for palladium, silver, nickel and yttrium appears to bemore metal dependent than microbial species dependent. Silver was mostly locatedon the external surfaces of the cells (Figure 5.4). Palladium was mostly located inside

Figure 5.2 Electron micrographs (TEM) of R. arrhizus cell wall thin section before (a) and after(b) uranium biosorption


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the cells (Figure 5.5), while yttrium occupied mostly cellular membrane sites, and toa substantially lesser extent inner specific sites (Figure 5.6).

The mechanism of the metabolically mediated bioaccumulatory metal uptake hasbeen studied and has been reported for the cases of Alcaligenes spp. (Diels, 1989),Citrobacter spp. (Macaskie, 1991) and Desulfovibrio spp. (Diels et al., 1995).These mechanisms involve the metabolically mediated production of a chemicalagent which precipitates the element of interest in the near-cell area. Thus, forexample, the Citrobacter species continuously produce inorganic phosphate by the

Figure 5.3 Electron micrographs (TEM) of R. arrhizus cell wall thin section after thoriumbiosorption

Figure 5.4 Electron micrograph (TEM) of AS302 cells following Ag biosorption (a), EDAXconfirmation of Ag retained (b)

Figure 5.5 Electron micrograph (TEM) of AS302 cells following Pd biosorption (a), EDAXconfirmation of Pd retained (b)

Figure 5.6 Electron micrograph (TEM) of AS302 cells following Y biosorption (a), EDAXconfirmation of Y retained (b)


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action of an acid-phosphatase type enzyme on an organic phosphate ‘donor’molecule, to precipitate heavy metals as cell-bound metal phosphate (Macaskie,1991). The technique has been applied for the sequestering of strontium,lanthanum, americium and Plutonium (Macaskie and Dean, 1985; Tolley et al.,1991).

Under specific physiological circumstances Alcaligenes eutrophus can alsoprecipitate metal species, leading to the bioaccumulation of these species. Thisaccumulation is the result of the progressive alkalinisation of the cell periplasmicspace by the action of a metal efflux system which continuously generates OH- ionsin the periplasm. Metal hydroxides thus precipitate on the cell envelopes usingmembrane components as a support (Diels, 1989; Diels et al., 1995).

Desulfovibrio bacteria can reduce sulphate to sulphide, thus providing a sulphide-rich environment in their immediate space, leading to metal sulphide precipitation.The system requires the supply of a sulphur or sulphate substrate and leads to thebioaccumulation of the metal species via the precipitation of their low-solubilitysulphides (Diels et al., 1995).

The dissimilatory metal reduction of uranium (VI) to insoluble uranium (IV) andthe corresponding removal and potential recovery of the uranium from dilutesolutions by microorganisms of the Shewanella alga type have also been reported. Asa result of this enzymatically mediated reduction the bioaccumulation of uranium isobserved. Similar work has been reported for uranium (VI) reduction byDesulfovibrio desulfuricans. The above processes can be classified in thebioaccumulatory process category as they rely on the activity of enzymes to carry outtheir metal sequestering function through the precipitation of the metal species ofinterest. The use of the above process in association with a bicarbonate extractionstage has been proposed for the bioremediation of uranium-contaminated soils(Lovley and Phillips, 1992; Lovley et al., 1991; Phillip et al., 1995).

The lanthanides and actinides

The lanthanide elements (rare earths) are a group of elements characterised by strongsimilarities in their chemistry with atomic numbers ranging from 58 to 71. This is thelargest naturally occurring group of elements in the periodic table (with the exceptionof the unstable Pm147, half life of 2.62 years). The lanthanides are not rare: over 100minerals are known to contain lanthanides (Greenwood and Earnshaw, 1993). Theirchemistry is dominated by the +3 oxidation state; they are electropositive and reactivemetals. They primarily form ionic type bonds and their cations display a typicalClass-A preference for O-donor ligands, a property which will be discussed laterwhen we will deal with the subject of competing ion effects.

The actinides are 14 chemically related elements with atomic numbers from 90 to103. Of these, only the first three are naturally occurring: thorium, protactinium anduranium. The rest are the transuranium elements which are artificially produced. Theyare naturally radioactive elements existing in mixtures of isotopes. They are closelyrelated to the uranium nuclear fuel cycle, hence their environmental significance.Also closely linked to the uranium nuclear fuel cycle are radioactive isotopes of otherelements, such as those of radium-224, 225, 226, radon-222, lead-210, 211, 214, etc.,which are daughter products of the thorium or uranium radioactive decay series.Under unusual conditions, such as those postulated to have occurred during the ‘Oklo


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phenomenon’, natural nuclear reactors can operate, generating fission products ofactinides within small regions and at elevated concentrations (West, 1976;Greenwood and Earnshaw, 1993).

Actinides are electropositive and reactive, with most current knowledgeconcentrated on the chemistry of uranium and, to a lesser extent, thorium. For thefirst three elements of the actinides the most stable oxidation state is the oneinvolving all the valence electrons. Additional oxidation states are possible. Thecommon oxidation state is +6 for uranium and +4 for thorium.

Application of biosorption

The application of biosorption for the sequestering of the lanthanides, the actinidesand related elements was primarily motivated by environmental concerns over therelease to the environment and the subsequent fate of radioactive isotopes from theuranium nuclear fuel power generation cycle. Therefore, interest has focused mostlyon uranium, thorium, radium and, to a lesser extent, other elements associated withnuclear activities, such as cobalt and strontium. Interest in the application ofbiosorption for rare earths sequestering is more recent and originated, primarily, withindustrial interest in scavenging and recovering rare earth metal values from aqueousdilute process or waste streams.

Information on biosorption will be presented separately for elements of interest inthese groups.

Uranium biosorption

In examining the biosorptive uptake of uranium by microbial biomass, theequilibrium and the rate of the process need to be defined. The equilibrium ofbiosorption has been successfully described by the use of the Langmuir andFreundlich relationships which show the equilibrium distribution of the biosorbedelement between the solution (liquid phase) and the microbial biomass (solid phase).Both models have been used and reported on (Tsezos and Keller, 1983; Tsezos, 1985,1990; Tsezos et al., 1995, 1996a; Glombitza et al., 1984; Georgousis, 1990).Attention must be paid to the fact that these models cannot be attributed anymechanistic significance and should only be interpreted as mathematical tools fordescribing the distribution of the element between the solid and the liquid phases inbiosorption. The effects of parameters such as the solution pH, the biomass growthconditions and the solution ionic matrix on the microbial biomass biosorptive uptakehave been discussed in detail and have been presented in other publications by severalauthors (Tsezos, 1985; Tsezos and McCready, 1989; Tsezos, 1990; Ehrlich andBrierley, 1990). Therefore, the detailed discussion on the effects of the aboveparameters on the biosorptive uptake of the metals of interest will not be discussed inthis chapter.

Most of the uranium biosorptive uptake studies have been conducted utilisingsynthetic uranium solutions, i.e. single-element solutions. The correspondingsolution ionic matrices have, therefore, been kept simple, well defined andcontrollable. Less work has been carried out and reported on industrial or complexmatrix solutions.


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Several different microbial biomass types have been examined for their uraniumuptake capacity. Figure 5.7 shows typical reported uranium biosorption isotherms forsimple uranyl nitrate solutions at moderately acidic pH values (pH=4) (Tsezos andVolesky, 1981).

The isotherms in Figure 5.7 demonstrate that the biosorptive uptake of uraniumcan be significant (up to about 20% of the biomass dry weight). They also suggestthat the uranium biosorptive uptake can be efficient and ‘aggressive’ since selectedbiomass types may demonstrate high uranium uptake capacities at low equilibriumuranium solution concentrations. This is a very desirable characteristic for theprocesses application potential of biosorption, as it secures significant biomassuranium loadings at low residual solution uranium concentrations. Table 5.1

Figure 5.7 Comparison of uranium uptake capacities for selected sorbent materials


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summarises reported uranium biosorption uptake capacities by a variety of microbialbiomass types (Tsezos and Volesky, 1981; Tobin et al., 1994).

Similar order uranium uptake capacities have been reported for several biomasstypes as, for example, Saccharomyces cerevisiae (15% w/w), Aspergillus niger(21.4% w/w) and Penicillium Cl (17% w/w) at moderately acidic pH values (Tobin etal., 1994).

Very few kinetic experiments on the rate of uranium biosorption by microbialbiomass have been reported (Tobin et al., 1994). The results available have shownthat the intrinsic rate of uranium biosorption by R. arrhizus is a very rapid processand will likely not be the rate limiting step in any engineering application ofbiosorption (Tsezos and Volesky, 1982b; Tsezos et al., 1988; Tsezos and McCready,1989; Tsezos, 1990; Tobin et al., 1994; Ryon et al., 1982). Figure 5.8 shows a typicalintrinsic uranium biosorption rate curve for native R. arrhizus biomass and confirmsthe above conclusion. The use of immobilised R. arrhizus microbial biomass,however, results in a completely different kinetic behaviour as diffusional processessuperimpose on the intrinsic uranium biosorption rate resulting in substantiallyslower kinetics. Figure 5.9 is a typical example of the rate of uptake of uranium byimmobilised R. arrhizus biomass from synthetic uranyl nitrate solutions. Comparisonof the curves in Figures 5.8 and 5.9 clearly shows the effects of diffusion on theobserved overall uranium biosorption rate when the biomass is immobilised intoparticulate form (Ehrlich and Brierley, 1990; Tsezos and Volesky, 1981; Tsezos andMcCready, 1991; Tsezos and Deutschmann, 1992; Ryon et al., 1982).

The technical application potential of uranium biosorption is substantiallydependent on the recovery of the uranium which has been sequestered by themicrobial biomass as well as the potential for re-using the regenerated biomass inmultiple biosorption-desorption cycles. The recovery of the adsorbed uranium can beachieved by the use of an appropriate elution solution capable of effectively strippingthe adsorbed uranium from the exhausted biomass and bringing it back to a solution.The elution must be complete, with no damage to the microbial biomass structure. Asystematic study on the elution of uranium which has been sequestered by microbialbiomass has been reported (Tsezos, 1984). The work has suggested that sodium

Table 5.1 Reported uranium biosorption uptake capacities (atvarious pH values)


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Figure 5.8 Uranium concentration in solution during biosorption by R. arrhizus at pH 4: kineticdata

Figure 5.9 Comparison of experimental (?) and model-predicted (line) uranium solutionconcentration profiles


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bicarbonate solutions are the most appropriate eluents for uranium, as theycompletely strip the biosorbed uranium while maintaining intact the microbialbiomass uranium biosorption characteristics. Mineral acids and sulphate-richsolutions have been shown to damage the microbial biomass re-use potential (Tsezos,1984). Table 5.2 summarises the effect and performance of a variety of elements usedfor uranium elution on R. arrhizus biomass.

Implementation of optimised solid to liquid ratios in elution enables the generationof highly concentrated uranium eluates with concentration factors of over 103

(Tsezos, 1984).

Engineering applications of uranium biosorption

Studies on the engineering application of biosorption for the recovery of uraniumfrom industrial process or waste solutions in batch form and at laboratory scalecontinuous pilot installations have been reported. The solutions treated have been thebiological leachates of uranium-bearing pyritic ore from the Elliot Lake district ofCanada (Tsezos, 1990; Tsezos and McCready, 1991). The above leachates aretypically dilute, very complex solutions with a pH value in the range of 1–2 anduranium concentrations in the range of 200–500 mg/l.

The continuous laboratory pilot testing of uranium biosorption as a process for theremoval/recovery of uranium from the above complex waste or process solutions hasconfirmed that biosorption is a very selective process and that uranium can beselectively sequestered by the microbial biomass out of the complex leachate solutionmatrix. Figures 5.10 and 5.11 show, respectively, typical pilot plant performance datafor the biosorption stage (breakthrough curve for uranium) and the elution stage(uranium concentration profile) for typical biosorption-elution cycles reported.Figure 5.12 summarises the uranium elution efficiency reported for the first 11 cycles

Table 5.2 Optimal uranium reloading of R. arrhizus following elution


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Figure 5.10 Typical uranium biosorption breakthrough curve

Figure 5.11 Typical uranium elution curve


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of one continuous pilot plant operation, suggesting the complete recovery of allbiosorbed uranium for each cycle (Macaskie, 1991; Tsezos, 1990; Tsezos andMcCready, 1991; Tsezos et al., 1996c).

Ionic competition effects

In the course of the continuous pilot plant testing of the biosorptive uranium recoveryfrom mine leach solutions by immobilised microbial biomass of R. arrhizus, agradual reduction of the uranium biosorptive uptake capacity of the biomass has beenreported (Tsezos and McCready, 1991; Tsezos et al., 1996c). These results aresummarised in Table 5.3.

Although the recovery of uranium, in each sorption/elution cycle, was complete,the total mass of uranium sequestered in a given cycle by a specific immobilisedbiomass quantity gradually declined. The phenomenon was investigated via the useof experimental techniques involving electron microscopy, microprobe analysis andequilibrium studies. The results of this work have suggested an interestingmechanism of interference between uranium and aluminium co-existing within thesame solution during their biosorption by the microbial biomass of R. arrhizus. Theinterference mechanism operates via a shift in the contact solution pH, caused bythe microbial biomass. This shift is more prominent in the immediate region of themicrobial cell and brings the contact solution within the cell wall chitin networkclose to neutral solution pH values. Aluminium is an element which hydrolysesextensively at near-neutral pH. It generates a complex range of low-solubility

Figure 5.12 Summary of uranium elution efficiency observed during the first 11 cycles of thepilot plant operation


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hydrolysis products. Aluminium is sequestered by the microbial biomass as shownin the typical aluminium biosorption breakthrough curve (Figure 5.13) which hasbeen observed and reported on in the course of the operation of a continuousuranium recovery biosorption pilot plant which was fed by uranium mine leachate(Figure 5.14).

The hydrolysis of aluminium within the fungal cell wall leads to the precipitation ofmetastable amorphous aluminium hydrolysis species within the cell wall. Thisprecipitate gradually fills the voids of the chitin cell wall network and limits the abilityof the fungal cell to biosorb uranium by primarily affecting the second of the threeprocesses active in the uranium uptake mechanism (Tsezos, 1984; Georgousis, 1990).

Table 5.3 Loading/elution cycling results

Figure 5.13 Typical Al breakthrough curve


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This mechanism of interference is a typical example of what we can call the ‘sterichindrance’ type of competition among elements in biosorption. Our systematic workon the subject of ionic competition in biosorption has suggested the existence of asecond type of mechanism of interference in biosorption among metals co-existing incomplex solutions, which we can call the ‘binding competition’ type of mechanism(Georgousis, 1990; Tsezos, 1984; Tsezos et al., 1995, 1996a).

Microbial biomass provides ligand groups on which metal species may bind bydifferent mechanisms. Major classes of microbial biopolymers, such as proteins,nucleic acids and polysaccharides, provide sites on which metal ions may bind. Theligand groups available include negatively charged groups, such as carboxylate,thiolate, or phosphate and groups such as amines, which often coordinate to the metalthrough lone pairs of electrons. The metal ionic species should exhibit a preferencefor the ligand binding sites of the biomass based on their chemical coordinationcharacteristics. Different ionic species of the same element can potentially exhibitpreference for different binding sites.

If the preference of one metal ion for a ligand is similar to that of another metalion, a competition effect could result between the metals for that given binding site.As a result of this type of competition when two metal species compete, thebiosorptive uptake capacity for the targeted metal can be lower than thatcorresponding to single metal solutions of the targeted element. If, however, the metalions species exhibit preferences for different biomass binding sites, theirsimultaneous presence in solution may not significantly affect their individual uptakecapacities by the microbial biomass used. In order to understand such competitioneffects, it has often been suggested that the differentiation of metals’ expected

Figure 5.14 Laboratory scale immobilised R. arrhizus biomass pilot plant treating uranium minewastewaters


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behaviour according to Pearson’s classification is a successful tool (Georgousis,1990; Avery and Tobin, 1993; Brady and Tobin, 1994; Tsezos et al., 1996b).

The effects of ionic competition in the biosorption of metals have been reportedfor two strains of microbial biomass and the metals palladium, gold, uranium,yttrium, silver and nickel on the basis of their Pearson classification (Georgousis,1990; Tsezos et al., 1996b). The selection of appropriate pairs of metals permitted theexamination of combinations of metals representative of each class (A, B,borderline). The biosorption results obtained from solutions containing each pair ofmetals have been compared to the corresponding single metal biosorption results.These results have shown that elements belonging to either the hard or soft classexhibit binding competition effects among members of their own class. Borderlineelements were affected by the presence of either hard or soft elements. Pearson’sreasoning appears to be a useful tool in interpreting aspects of the ‘bindingcompetition’ mechanism, but needs to be assisted by a detailed examination of metalsolution (hydrolysis behaviour, stereochemical) and biomass characteristics.

Thorium biosorption

The interest in the biosorption of thorium, as evidenced by the number of paperspublished on the subject, is substantially less than that in the biosorption of uranium,perhaps because thorium does not have the same economic significance as uranium.Thorium, however, commonly exists along with uranium in nature and, from anenvironmental point of view, the biosorption of thorium is of interest (Tsezos andVolesky, 1981).

In general, thorium appears to be sequestered well by microbial biomass (Tsezosand Volesky, 1981; Tobin et al., 1994). The locus of thorium biosorption in the caseof R. arrhizus has been reported to be different from that of uranium (Figure 5.3).Although both elements are retained primarily by the fungal cell wall, uranium islocalised within the cell well chitin network while thorium is localised on the externalsurface of the cell wall. This difference in the biosorptive loci enables thesimultaneous biosorption of uranium and thorium from the same solution by the samebiomass without immediate competition effects. Reported results on the operation ofa biosorption pilot plant utilising immobilised R. arrhizus biomass and treating acidicmine waters from an uranium mine in Canada have shown both uranium and thoriumto be biosorbed by the immobilised biomass particles (Ehrlich and Brierley, 1990;Tsezos and McCready, 1991). The biosorptive uptake of thorium was very efficient(Tsezos and McCready, 1991).

The intrinsic kinetics of thorium biosorption has also been reported for single-element solutions and for the biomass of R. arrhizus (Tsezos and Volesky, 1981,1982a). The intrinsic kinetics is very rapid, as for the case of uranium. Systematicstudies on the elution of thorium are not available. Table 5.4 and Figure 5.15summarise representative information available on the biosorptive uptake of thoriumby several biomass types from single-element solutions at the optimal solution pH.

Radium biosorption

Radium as an element does not belong to the lanthanides or actinides groups. It is,however, closely associated with them, as radium isotopes are daughter products of


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the uranium-thorium radioactive decay series. Radium-226 is of particularenvironmental interest because it has a long half life and generates radon, a gaseousradioactive daughter product (Ryon et al., 1982; Tsezos, 1985).

Most of the work reported on radium sequestering refers to several types ofinorganic adsorbents such as ion exchange resins or zeolites (Greenwood andEarnshaw, 1993). Limited information is available on the biosorption of radium.Early work by the Czech Atomic Energy Commission reported radium biosorptiveuptake by Penicillium chrysogenum to the order of 103 pCi/1 of wet biomass(Stamberg et al., 1975). In another publication, municipal sludge originating fromtwo Canadian wastewater treatment plants was reported to have sequestered radiumup to 1024 pCi/kg (Durham and Joshi, 1979).

In a systematic evaluation of radium biosorption, selected samples of wastemicrobial biomass, used in industrial fermentation processes and wastewaterbiological treatment plants, were studied for their radium biosorption ability fromaqueous solutions. Equilibrium biosorption isotherms were used to quantify theradium uptake capacity of the various types of biomass, which were also compared totwo types of activated carbon. Solution pH was shown to affect the observed uptakesignificantly. In general, the biomass types which showed appreciable sorptioncapacity exhibited maximum uptake between pH 7 and 10. The uptake was reducedconsiderably at pH 4, and little or no uptake was observed at pH 2. Radiumbiosorptive uptake capacities of the order of 4.5×104 nCi/g at pH 7 and at anequilibrium radium concentration of 1000 pCi/1 were determined for a mixed culture,while the biomass of Penicillium chrysogenum adsorbed 5×104 nCi/g radium underthe same conditions.

Figure 5.16 shows typical examples of linearised radium biosorption isotherms forthe biomass of Rhizopus arrhizus, demonstrating the effect of solution pH on theobserved radium biosorptive uptake (Tsezos and Keller, 1983; Tsezos et. al., 1986c).Competitive radium biosorption equilibrium uptake studies have also been reportedfor Penicillium chrysogenum and a mixed culture from a municipal wastewatertreatment installation (Tsezos et. al., 1986c). The IIA group of elements was reportedto be the most effective radium cationic competitors. Iron was also reported to act as

Table 5.4 Reported thorium biosorption uptake capacities (atvarious pH values)


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a competing element. Fine FeO(OH) precipitates formed at near-neutral pH valueshave been reported to coat the surface of the microbial biomass cells, limiting accessof radium to the biomass biosorption sites. A similar phenomenon has been reportedfor the case of uranium biosorption (Tsezos et al., 1986a).

The potential of eluting the biosorbed radium by washing the loaded microbialbiomass with a wide spectrum of potential eluants has been reported (Tsezos et al.,1986b). In that report mineral acids and EDTA solutions were shown to be the mostefficient radium eluants. The rate of radium elution is reported to be very rapid, withcomplete elution achieved within one or two minutes (Tsezos et al., 1986b).

Figure 5.15 Comparison of thorium uptake capacities for selected sorbent materials


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The radium re-adsorption capacity of the microbial biomass following elution wasreported to be reduced substantially as the acidic elements damaged the microbial cellarchitecture (Tsezos et al., 1986b).

Immobilised microbial biomass has been used in a laboratory-scale continuouspilot plant for the treatment of radium-bearing waste waters from the Elliot Lakedistrict of Canada (Tsezos et al., 1987). In that report, the equilibrium radium uptake(~ 200 nCi/g), the kinetics of radium uptake and the regeneration/re-use of theimmobilised biomass were reported, suggesting that biosorption can be an efficientprocess for the selective extraction of radium from the waste streams. The subsequentelution of radium in a concentrated form and the re-use of the biomass in a limitednumber of cycles have been reported as possible (Tsezos et al., 1986b). Table 5.5summarises the reported re-use potential of the immobilised biomass particles, wherea mixed culture of predominantly bacterial organisms from a municipal wastewatertreatment plant was used.

The work reported on radium biosorption has suggested that the biosorptivesequestering of radium could be a reasonable alternative to the Ba-Ra sulphateprecipitation technology as it does not produce, as a by-product, large volumes ofradioactive sludge and it is affected less than ion exchange resins by the presence of IIAelements.

Closing comments

The information presented in this chapter summarises some of the work and theexperience accumulated over the past 20 years on the biosorption of members of thelanthanides, actinides and related elements. One could potentially include more

Figure 5.16 Linearised radium-226 adsorption isotherms by inactive biomass of Rhizopusarrhizus


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elements such as Sr or Co, which are related to nuclear applications, or includedaughter products of the radioactive decay series of some of the elements discussedabove. However, these are outside the scope of the present chapter.

It is important to note that 20 years ago the mechanistic understanding ofbiosorption was quite nebulous, and biosorption was mostly an interestingphenomenon related mainly to microorganisms. Since then a substantial volume ofsystematic work has been added. The engineering applications potential of thephenomenon is being investigated, and numerous scientists and engineers areworking on the subject. The differentiation of the ‘biosorptive’ versus the‘bioaccumulatory’ process has also been a positive step in the direction of the betterunderstanding of the underlying mechanisms in biosorptive phenomena. Thespecificity of biosorption makes it an excellent candidate technology for industrialapplications where large volume, low concentration, complex ionic matrix waste orprocess solutions need to be treated for the purpose of sequestering targeted elements.

Tertiary or polishing treatment applications are therefore good candidateapplication areas. Finally, it should be noted that the phenomenon of biosorptionexists not only for inorganic ionic species but also for organic molecules, and thisobservation opens up the opportunity to study the interactions between biosorptionand biodegradation for organic molecules of interest (Tsezos and Wang, 1991).

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Biosorption of Lanthanides Actinides and Related Materials