Download Biosorbents for Metal Ions

Biosorbents for
Metal Ions
Edited by
DR 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 3DE
USA 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 1997
All rights reserved. No part of this publication may be reproduced, stored in a retrieval
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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
Contents
List of Contributors
1 Biosorption of heavy metals: an introduction (C.F.Forster; D.A.J.Wase)
Introduction
Toxic metals
Control
Treatment
References
ix
1
1
2
4
5
9
2 The use of algae as metal biosorbents (G.W.Garnham)
Introduction
Biosorption by algae and the mechanisms involved
Factors affecting the biosorption of metals by algae
Production and cost of algal biomass for metal removal
Immobilised algae and derived products
Algal biosorption processes and engineering considerations
Commercial algal biosorption
References
11
11
12
20
23
26
27
30
33
3 General bacterial sorption processes (M.M.Urrutia)
Introduction
Bacterial surface
Biofilms
Charge of bacterial cell surfaces
Sorption of metal cations and mechanisms
Sorption of metal anions and mechanisms
Binding constants
Modelling
Applications in biotechnology
Summary
References
39
39
39
43
43
46
51
52
57
58
59
59
v
Contents
4 Fungi as biosorbents (A.Kapoor; T.Viraraghavan)
Introduction
Modes of metal ion uptake
Modelling of biosorption
Biosorption by living cells
Biosorption of metal ions by non-living cells
Regeneration of fungal biomass and elution of biosorbed metals
Use of immobilised fungal biomass in biosorption
Biosorption mechanism
General considerations in the use of fungi as biosorbents
References
67
67
67
68
69
73
77
78
78
79
80
5 Biosorption of lanthanides, actinides and related materials (M.Tsezos)
Introduction
The mechanism of biosorption/bioaccumulation
The lanthanides and actinides
Application of biosorption
Uranium biosorption
Thorium biosorption
Radium biosorption
Closing comments
References
87
87
89
96
97
97
106
106
109
110
6 Scavenging trace concentrations of metals (C.J.Banks)
Introduction
Coincidental sorption systems
Biosorption systems specifically for metal removal
Exposure of biosorbent surfaces to metal-laden wastewaters
Immobilisation matrices
Comparisons of reactor designs
References
115
115
116
121
123
128
136
136
7 Low-cost biosorbents: batch processes (D.A.J.Wase; C.F.Forster; Y.S.Ho) 141
Introduction
141
Peat
141
Other biosorbents
146
Novel activated carbons
147
Copper
148
Nickel and lead
148
Chromium
149
Zinc
151
Manganese
153
Cobalt and cadmium
153
Competitive adsorption
154
Practical aspects of using peat
155
References
158
vi
Contents
8 Biosorption using unusual biomasses (R.G.J.Edyvean; C.J.Williams;
M.M.Wilson; D.Aderhold)
Introduction
Types of biomass
Performance
Factors affecting adsorption
Industrial scale systems
Conclusions
References
165
165
166
170
171
177
178
179
9 Low-cost adsorbents in continuous processes (G.McKay; S.J.Allen)
Introduction
Peat, lignite and chitosan as sorbents for metal ions
Sorption column design
Regeneration and metal recovery
References
183
183
188
195
216
217
10 Biosorption: the future (C.F.Forster; D.A.J.Wase)
Introduction
Algal biosorption
Fungal biosorption
Bio-wastes
The future
Conclusions
References
Index
221
221
222
222
223
225
226
227
229
vii
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 ions
sequestering by living or inactive microbial biomass has been seen in the scientific
and engineering community. This phenomenon has potential application in
environmental pollution control as a biochemical process on which corresponding
unit operations can be designed and operated by industry.
Previous publications on the subject have proposed the adoption of two different
terms for the description of the two mechanistically different types of metals
sequestering by microorganisms. The term ‘bioaccumulation’ has been proposed for
the sequestering of metal ions by metabolically mediated processes (living
microorganisms), and the term ‘biosorption’ for the sequestering by nonmetabolically mediated process (inactive microorganisms) (Diels et al., 1995). As our
understanding of the above processes has increased, the mechanistic differences
between biosorption and bioaccumulation have proved to be so significant that the
use 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 coexist and can also function independently as, for example, in the case where a
consortium of microorganisms is exposed to metal-bearing solutions. Literature on
both 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 and
McCready, 1989).
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Biosorbents for Metal lons
This chapter will focus more on the phenomenon of biosorption, and, in particular, on
the removal of lanthanides, actinides and related elements.
Biosorption of metals is generally characterised by high selectivity as compared to
ion-exchange resins or other adsorbents. This selectivity is considered to be a
desirable feature in designing processes for pollution control and/or metal value
recovery (Tsezos and Volesky, 1982b; Tsezos, 1985; Tsezos and McCready, 1991;
Diels, 1989; Diels et al., 1995; Macaskie, 1991; Glombitza et al., 1984; Pumpel and
Schinner, 1986; Gadd, 1992). In addition to selectivity, biosorptive processes have the
following 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 metal
values from dilute industrial complex aqueous solutions, the extraction of
radionuclides, e.g. uranium, thorium or radium from mine leachates, and similar
metal 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 a
dilute complex ionic matrix and large volumes of aqueous process or waste solutions
from which the selective extraction and, occasionally, recovery of targeted elements
via the use of the microbial biomass is intended. Regardless of the detailed
engineering configuration of such a process, a stage which significantly affects the
overall efficiency and the economics of the technology is the separation of the
microbial biomass from the waste or process waters following contact (SENES
Consultants, 1985).
As a result of this constraint, contact systems making use of microbial biomass
immobilised on a support medium have been developed and proposed for use. Two
generically different types of immobilised biomass contact systems have been
proposed. The first type is based on the use of immobilised biomass particles which
are produced via the use of a wide range of biomass binding agents, such as synthetic
polymers (e.g. polysulphones), natural polymers (e.g. alginates) or chemical biomass
treatment (Brierley et al., 1986; Kiff and Little, 1986; Tobin et al., 1994; Tsezos and
Deutschmann, 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; Brierley
and Vance, 1988; Darnall et al., 1989). Each one of the two types of immobilised
biomass necessitates the implementation of different contact reactor design, such as
upflow or downflow packed-bed reactors, rotating biological contactors, membrane
sheet or tubular reactors, etc. Figure 5.1 shows a typical example of an immobilised
biomass particle of the first type in two different magnifications.
It is interesting to note the highly porous structure of the particles shown in Figure
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Lanthanides, Actinides and Related Materials
Figure 5.1 Electron micrographs (TEM) of immobilised biomass particles: (a) general view;
(b) magnification of the particle porous structure
5.1 which is required in order to facilitate and improve the kinetics of metal ions
diffusion into the inner particle active biosorption sites (Tsezos et al., 1988; Tsezos
and Deutschmann, 1990).
The mechanism of biosorption/bioaccumulation
Although a large volume of work has been published and reported on the
assessment of the uptake capacities of several microbial biomass types for a variety
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Biosorbents for Metal lons
of metallic elements, systematic effort to elucidate the underlying mechanisms has
been limited. The way in which elements bind or are retained by specific microbial
biomass species is understood in detail only for limited combinations of biomass/
metal ion pairs.
The mechanistic understanding of biosorption is considered essential in order to
optimise the process application potential of biosorption. More specifically, this
understanding is essential in order to exploit optimally the selectivity and efficiency
of the process and to overcome ionic competition and interference effects by other
ionic species which exist along with the targeted element in the ionic matrix of the
contact 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 microbial
species and element dependent (Tsezos and Volesky, 1982a, 1982b; Tsezos et al.,
1996b; Avery and Tobin, 1992; Lovley et al., 1991; Lovley and Phillips, 1992; Tolley
et al., 1991). Reported mechanisms of biosorption are briefly presented below,
illustrating the wide variety of physical-chemical phenomena which are involved
during biosorptive uptake.
The biosorption of uranium by R. arrhizus takes place inside the mycellial cell
wall. Retained uranium is taken up via three independent but interrelated processes
(Tsezos and Volesky, 1982b). The first process involves the coordination of uranyl
ions by the mycellial cell wall chitin nitrogen. The second process involves the
physical adsorption of uranyl ions within the chitin three-dimensional network. The
third process involves the hydrolysis of the uranyl ion-chitin complex and the
precipitation of additional uranium hydrolysis species within the cell wall chitin
network. Figure 5.2 shows typical transmission electron micrographs of the R.
arrhizus mycellial cell wall before and after contact with uranium. The electron-dense
areas on the post-contact micrograph are the uranium-bearing zones.
The mechanism of thorium biosorption by the same organism is different (Tsezos
and Volesky, 1982a). Thorium is retained primarily by adsorption on the external
surface of the mycellial cell wall. Chitin involvement in thorium biosorption is of
substantially reduced significance as compared to its role during uranium biosorption.
Figure 5.3 shows typical transmission electron micrographs of R. arrhizus cells after
thorium biosorption. The electron-dense areas on the outer cell wall are the thoriumbearing zones.
The biosorption of strontium by inactive yeast cells (S. cerevisiae) has been
reported to be primarily an electrostatic attraction of the Sr2+ by the yeast cells,
while living cells sequester Sr 2+ by a more complex mechanism involving ion
exchange with strontium residing primarily within the cell vacuoles (Avery and
Tobin, 1992).
Work involving the use of EXAFS and XANES techniques reported on the
biosorption of Au by the algal biomass of C. vulgaris has demonstrated the binding of
gold to be primarily the result of ligand exchange reactions leading to the formation
of 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 the
biosorption 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
90
Lanthanides, Actinides and Related Materials
Figure 5.2 Electron micrographs (TEM) of R. arrhizus cell wall thin section before (a) and after
(b) uranium biosorption
capacities, were examined using EM and EDAX microprobe analysis. It was reported
that the locus of biosorption for palladium, silver, nickel and yttrium appears to be
more metal dependent than microbial species dependent. Silver was mostly located
on the external surfaces of the cells (Figure 5.4). Palladium was mostly located inside
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Biosorbents for Metal lons
Figure 5.3 Electron micrographs (TEM) of R. arrhizus cell wall thin section after thorium
biosorption
the cells (Figure 5.5), while yttrium occupied mostly cellular membrane sites, and to
a substantially lesser extent inner specific sites (Figure 5.6).
The mechanism of the metabolically mediated bioaccumulatory metal uptake has
been 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 chemical
agent which precipitates the element of interest in the near-cell area. Thus, for
example, the Citrobacter species continuously produce inorganic phosphate by the
92
Figure 5.4 Electron micrograph (TEM) of AS302 cells following Ag biosorption (a), EDAX
confirmation of Ag retained (b)
Figure 5.5 Electron micrograph (TEM) of AS302 cells following Pd biosorption (a), EDAX
confirmation of Pd retained (b)
Figure 5.6 Electron micrograph (TEM) of AS302 cells following Y biosorption (a), EDAX
confirmation of Y retained (b)
Biosorbents for Metal lons
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 also
precipitate metal species, leading to the bioaccumulation of these species. This
accumulation is the result of the progressive alkalinisation of the cell periplasmic
space by the action of a metal efflux system which continuously generates OH- ions
in the periplasm. Metal hydroxides thus precipitate on the cell envelopes using
membrane components as a support (Diels, 1989; Diels et al., 1995).
Desulfovibrio bacteria can reduce sulphate to sulphide, thus providing a sulphiderich environment in their immediate space, leading to metal sulphide precipitation.
The system requires the supply of a sulphur or sulphate substrate and leads to the
bioaccumulation of the metal species via the precipitation of their low-solubility
sulphides (Diels et al., 1995).
The dissimilatory metal reduction of uranium (VI) to insoluble uranium (IV) and
the corresponding removal and potential recovery of the uranium from dilute
solutions by microorganisms of the Shewanella alga type have also been reported. As
a result of this enzymatically mediated reduction the bioaccumulation of uranium is
observed. Similar work has been reported for uranium (VI) reduction by
Desulfovibrio desulfuricans. The above processes can be classified in the
bioaccumulatory process category as they rely on the activity of enzymes to carry out
their metal sequestering function through the precipitation of the metal species of
interest. The use of the above process in association with a bicarbonate extraction
stage 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 strong
similarities in their chemistry with atomic numbers ranging from 58 to 71. This is the
largest naturally occurring group of elements in the periodic table (with the exception
of the unstable Pm147, half life of 2.62 years). The lanthanides are not rare: over 100
minerals are known to contain lanthanides (Greenwood and Earnshaw, 1993). Their
chemistry is dominated by the +3 oxidation state; they are electropositive and reactive
metals. They primarily form ionic type bonds and their cations display a typical
Class-A preference for O-donor ligands, a property which will be discussed later
when we will deal with the subject of competing ion effects.
The actinides are 14 chemically related elements with atomic numbers from 90 to
103. Of these, only the first three are naturally occurring: thorium, protactinium and
uranium. The rest are the transuranium elements which are artificially produced. They
are naturally radioactive elements existing in mixtures of isotopes. They are closely
related to the uranium nuclear fuel cycle, hence their environmental significance.
Also closely linked to the uranium nuclear fuel cycle are radioactive isotopes of other
elements, 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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Lanthanides, Actinides and Related Materials
phenomenon’, natural nuclear reactors can operate, generating fission products of
actinides within small regions and at elevated concentrations (West, 1976;
Greenwood and Earnshaw, 1993).
Actinides are electropositive and reactive, with most current knowledge
concentrated on the chemistry of uranium and, to a lesser extent, thorium. For the
first three elements of the actinides the most stable oxidation state is the one
involving all the valence electrons. Additional oxidation states are possible. The
common oxidation state is +6 for uranium and +4 for thorium.
Application of biosorption
The application of biosorption for the sequestering of the lanthanides, the actinides
and related elements was primarily motivated by environmental concerns over the
release to the environment and the subsequent fate of radioactive isotopes from the
uranium nuclear fuel power generation cycle. Therefore, interest has focused mostly
on uranium, thorium, radium and, to a lesser extent, other elements associated with
nuclear activities, such as cobalt and strontium. Interest in the application of
biosorption for rare earths sequestering is more recent and originated, primarily, with
industrial interest in scavenging and recovering rare earth metal values from aqueous
dilute process or waste streams.
Information on biosorption will be presented separately for elements of interest in
these groups.
Uranium biosorption
In examining the biosorptive uptake of uranium by microbial biomass, the
equilibrium and the rate of the process need to be defined. The equilibrium of
biosorption has been successfully described by the use of the Langmuir and
Freundlich relationships which show the equilibrium distribution of the biosorbed
element 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 any
mechanistic significance and should only be interpreted as mathematical tools for
describing the distribution of the element between the solid and the liquid phases in
biosorption. The effects of parameters such as the solution pH, the biomass growth
conditions and the solution ionic matrix on the microbial biomass biosorptive uptake
have been discussed in detail and have been presented in other publications by several
authors (Tsezos, 1985; Tsezos and McCready, 1989; Tsezos, 1990; Ehrlich and
Brierley, 1990). Therefore, the detailed discussion on the effects of the above
parameters on the biosorptive uptake of the metals of interest will not be discussed in
this chapter.
Most of the uranium biosorptive uptake studies have been conducted utilising
synthetic uranium solutions, i.e. single-element solutions. The corresponding
solution ionic matrices have, therefore, been kept simple, well defined and
controllable. Less work has been carried out and reported on industrial or complex
matrix solutions.
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Biosorbents for Metal lons
Several different microbial biomass types have been examined for their uranium
uptake capacity. Figure 5.7 shows typical reported uranium biosorption isotherms for
simple uranyl nitrate solutions at moderately acidic pH values (pH=4) (Tsezos and
Volesky, 1981).
The isotherms in Figure 5.7 demonstrate that the biosorptive uptake of uranium
can be significant (up to about 20% of the biomass dry weight). They also suggest
that the uranium biosorptive uptake can be efficient and ‘aggressive’ since selected
biomass types may demonstrate high uranium uptake capacities at low equilibrium
uranium solution concentrations. This is a very desirable characteristic for the
processes application potential of biosorption, as it secures significant biomass
uranium 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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Lanthanides, Actinides and Related Materials
Table 5.1 Reported uranium biosorption uptake capacities (at
various pH values)
summarises reported uranium biosorption uptake capacities by a variety of microbial
biomass types (Tsezos and Volesky, 1981; Tobin et al., 1994).
Similar order uranium uptake capacities have been reported for several biomass
types 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 et
al., 1994).
Very few kinetic experiments on the rate of uranium biosorption by microbial
biomass have been reported (Tobin et al., 1994). The results available have shown
that the intrinsic rate of uranium biosorption by R. arrhizus is a very rapid process
and will likely not be the rate limiting step in any engineering application of
biosorption (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 typical
intrinsic uranium biosorption rate curve for native R. arrhizus biomass and confirms
the above conclusion. The use of immobilised R. arrhizus microbial biomass,
however, results in a completely different kinetic behaviour as diffusional processes
superimpose on the intrinsic uranium biosorption rate resulting in substantially
slower kinetics. Figure 5.9 is a typical example of the rate of uptake of uranium by
immobilised R. arrhizus biomass from synthetic uranyl nitrate solutions. Comparison
of the curves in Figures 5.8 and 5.9 clearly shows the effects of diffusion on the
observed overall uranium biosorption rate when the biomass is immobilised into
particulate form (Ehrlich and Brierley, 1990; Tsezos and Volesky, 1981; Tsezos and
McCready, 1991; Tsezos and Deutschmann, 1992; Ryon et al., 1982).
The technical application potential of uranium biosorption is substantially
dependent on the recovery of the uranium which has been sequestered by the
microbial biomass as well as the potential for re-using the regenerated biomass in
multiple biosorption-desorption cycles. The recovery of the adsorbed uranium can be
achieved by the use of an appropriate elution solution capable of effectively stripping
the 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. A
systematic study on the elution of uranium which has been sequestered by microbial
biomass has been reported (Tsezos, 1984). The work has suggested that sodium
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Biosorbents for Metal lons
Figure 5.8 Uranium concentration in solution during biosorption by R. arrhizus at pH 4: kinetic
data
Figure 5.9 Comparison of experimental (?) and model-predicted (line) uranium solution
concentration profiles
100
Lanthanides, Actinides and Related Materials
bicarbonate solutions are the most appropriate eluents for uranium, as they
completely strip the biosorbed uranium while maintaining intact the microbial
biomass uranium biosorption characteristics. Mineral acids and sulphate-rich
solutions 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 used
for uranium elution on R. arrhizus biomass.
Implementation of optimised solid to liquid ratios in elution enables the generation
of highly concentrated uranium eluates with concentration factors of over 10 3
(Tsezos, 1984).
Engineering applications of uranium biosorption
Studies on the engineering application of biosorption for the recovery of uranium
from industrial process or waste solutions in batch form and at laboratory scale
continuous pilot installations have been reported. The solutions treated have been the
biological leachates of uranium-bearing pyritic ore from the Elliot Lake district of
Canada (Tsezos, 1990; Tsezos and McCready, 1991). The above leachates are
typically dilute, very complex solutions with a pH value in the range of 1–2 and
uranium concentrations in the range of 200–500 mg/l.
The continuous laboratory pilot testing of uranium biosorption as a process for the
removal/recovery of uranium from the above complex waste or process solutions has
confirmed that biosorption is a very selective process and that uranium can be
selectively sequestered by the microbial biomass out of the complex leachate solution
matrix. Figures 5.10 and 5.11 show, respectively, typical pilot plant performance data
for 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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Biosorbents for Metal lons
Figure 5.10 Typical uranium biosorption breakthrough curve
Figure 5.11 Typical uranium elution curve
102
Lanthanides, Actinides and Related Materials
Figure 5.12 Summary of uranium elution efficiency observed during the first 11 cycles of the
pilot plant operation
of one continuous pilot plant operation, suggesting the complete recovery of all
biosorbed uranium for each cycle (Macaskie, 1991; Tsezos, 1990; Tsezos and
McCready, 1991; Tsezos et al., 1996c).
Ionic competition effects
In the course of the continuous pilot plant testing of the biosorptive uranium recovery
from mine leach solutions by immobilised microbial biomass of R. arrhizus, a
gradual reduction of the uranium biosorptive uptake capacity of the biomass has been
reported (Tsezos and McCready, 1991; Tsezos et al., 1996c). These results are
summarised 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 immobilised
biomass quantity gradually declined. The phenomenon was investigated via the use
of experimental techniques involving electron microscopy, microprobe analysis and
equilibrium studies. The results of this work have suggested an interesting
mechanism of interference between uranium and aluminium co-existing within the
same solution during their biosorption by the microbial biomass of R. arrhizus. The
interference mechanism operates via a shift in the contact solution pH, caused by
the microbial biomass. This shift is more prominent in the immediate region of the
microbial cell and brings the contact solution within the cell wall chitin network
close to neutral solution pH values. Aluminium is an element which hydrolyses
extensively at near-neutral pH. It generates a complex range of low-solubility
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Biosorbents for Metal lons
Table 5.3 Loading/elution cycling results
hydrolysis products. Aluminium is sequestered by the microbial biomass as shown
in the typical aluminium biosorption breakthrough curve (Figure 5.13) which has
been observed and reported on in the course of the operation of a continuous
uranium 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 of
metastable amorphous aluminium hydrolysis species within the cell wall. This
precipitate gradually fills the voids of the chitin cell wall network and limits the ability
of the fungal cell to biosorb uranium by primarily affecting the second of the three
processes active in the uranium uptake mechanism (Tsezos, 1984; Georgousis, 1990).
Figure 5.13 Typical Al breakthrough curve
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Lanthanides, Actinides and Related Materials
Figure 5.14 Laboratory scale immobilised R. arrhizus biomass pilot plant treating uranium mine
wastewaters
This mechanism of interference is a typical example of what we can call the ‘steric
hindrance’ type of competition among elements in biosorption. Our systematic work
on the subject of ionic competition in biosorption has suggested the existence of a
second type of mechanism of interference in biosorption among metals co-existing in
complex 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 by
different mechanisms. Major classes of microbial biopolymers, such as proteins,
nucleic acids and polysaccharides, provide sites on which metal ions may bind. The
ligand groups available include negatively charged groups, such as carboxylate,
thiolate, or phosphate and groups such as amines, which often coordinate to the metal
through lone pairs of electrons. The metal ionic species should exhibit a preference
for the ligand binding sites of the biomass based on their chemical coordination
characteristics. Different ionic species of the same element can potentially exhibit
preference for different binding sites.
If the preference of one metal ion for a ligand is similar to that of another metal
ion, 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, the
biosorptive uptake capacity for the targeted metal can be lower than that
corresponding to single metal solutions of the targeted element. If, however, the metal
ions species exhibit preferences for different biomass binding sites, their
simultaneous presence in solution may not significantly affect their individual uptake
capacities by the microbial biomass used. In order to understand such competition
effects, it has often been suggested that the differentiation of metals’ expected
105
Biosorbents for Metal lons
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 reported
for 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 the
examination of combinations of metals representative of each class (A, B,
borderline). The biosorption results obtained from solutions containing each pair of
metals have been compared to the corresponding single metal biosorption results.
These results have shown that elements belonging to either the hard or soft class
exhibit binding competition effects among members of their own class. Borderline
elements were affected by the presence of either hard or soft elements. Pearson’s
reasoning appears to be a useful tool in interpreting aspects of the ‘binding
competition’ mechanism, but needs to be assisted by a detailed examination of metal
solution (hydrolysis behaviour, stereochemical) and biomass characteristics.
Thorium biosorption
The interest in the biosorption of thorium, as evidenced by the number of papers
published 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 an
environmental point of view, the biosorption of thorium is of interest (Tsezos and
Volesky, 1981).
In general, thorium appears to be sequestered well by microbial biomass (Tsezos
and Volesky, 1981; Tobin et al., 1994). The locus of thorium biosorption in the case
of 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 is
localised within the cell well chitin network while thorium is localised on the external
surface of the cell wall. This difference in the biosorptive loci enables the
simultaneous biosorption of uranium and thorium from the same solution by the same
biomass without immediate competition effects. Reported results on the operation of
a biosorption pilot plant utilising immobilised R. arrhizus biomass and treating acidic
mine waters from an uranium mine in Canada have shown both uranium and thorium
to 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 singleelement 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. Systematic
studies on the elution of thorium are not available. Table 5.4 and Figure 5.15
summarise representative information available on the biosorptive uptake of thorium
by 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
106
Lanthanides, Actinides and Related Materials
Table 5.4 Reported thorium biosorption uptake capacities (at
various pH values)
the uranium-thorium radioactive decay series. Radium-226 is of particular
environmental interest because it has a long half life and generates radon, a gaseous
radioactive daughter product (Ryon et al., 1982; Tsezos, 1985).
Most of the work reported on radium sequestering refers to several types of
inorganic adsorbents such as ion exchange resins or zeolites (Greenwood and
Earnshaw, 1993). Limited information is available on the biosorption of radium.
Early work by the Czech Atomic Energy Commission reported radium biosorptive
uptake by Penicillium chrysogenum to the order of 10 3 pCi/1 of wet biomass
(Stamberg et al., 1975). In another publication, municipal sludge originating from
two Canadian wastewater treatment plants was reported to have sequestered radium
up to 1024 pCi/kg (Durham and Joshi, 1979).
In a systematic evaluation of radium biosorption, selected samples of waste
microbial biomass, used in industrial fermentation processes and wastewater
biological treatment plants, were studied for their radium biosorption ability from
aqueous solutions. Equilibrium biosorption isotherms were used to quantify the
radium uptake capacity of the various types of biomass, which were also compared to
two types of activated carbon. Solution pH was shown to affect the observed uptake
significantly. In general, the biomass types which showed appreciable sorption
capacity exhibited maximum uptake between pH 7 and 10. The uptake was reduced
considerably at pH 4, and little or no uptake was observed at pH 2. Radium
biosorptive uptake capacities of the order of 4.5×10 4 nCi/g at pH 7 and at an
equilibrium radium concentration of 1000 pCi/1 were determined for a mixed culture,
while the biomass of Penicillium chrysogenum adsorbed 5×104 nCi/g radium under
the same conditions.
Figure 5.16 shows typical examples of linearised radium biosorption isotherms for
the biomass of Rhizopus arrhizus, demonstrating the effect of solution pH on the
observed radium biosorptive uptake (Tsezos and Keller, 1983; Tsezos et. al., 1986c).
Competitive radium biosorption equilibrium uptake studies have also been reported
for Penicillium chrysogenum and a mixed culture from a municipal wastewater
treatment installation (Tsezos et. al., 1986c). The IIA group of elements was reported
to be the most effective radium cationic competitors. Iron was also reported to act as
107
Biosorbents for Metal lons
Figure 5.15 Comparison of thorium uptake capacities for selected sorbent materials
a competing element. Fine FeO(OH) precipitates formed at near-neutral pH values
have been reported to coat the surface of the microbial biomass cells, limiting access
of radium to the biomass biosorption sites. A similar phenomenon has been reported
for the case of uranium biosorption (Tsezos et al., 1986a).
The potential of eluting the biosorbed radium by washing the loaded microbial
biomass 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 most
efficient radium eluants. The rate of radium elution is reported to be very rapid, with
complete elution achieved within one or two minutes (Tsezos et al., 1986b).
108
Lanthanides, Actinides and Related Materials
Figure 5.16 Linearised radium-226 adsorption isotherms by inactive biomass of Rhizopus
arrhizus
The radium re-adsorption capacity of the microbial biomass following elution was
reported to be reduced substantially as the acidic elements damaged the microbial cell
architecture (Tsezos et al., 1986b).
Immobilised microbial biomass has been used in a laboratory-scale continuous
pilot plant for the treatment of radium-bearing waste waters from the Elliot Lake
district 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 the
immobilised biomass were reported, suggesting that biosorption can be an efficient
process for the selective extraction of radium from the waste streams. The subsequent
elution of radium in a concentrated form and the re-use of the biomass in a limited
number of cycles have been reported as possible (Tsezos et al., 1986b). Table 5.5
summarises the reported re-use potential of the immobilised biomass particles, where
a mixed culture of predominantly bacterial organisms from a municipal wastewater
treatment plant was used.
The work reported on radium biosorption has suggested that the biosorptive
sequestering of radium could be a reasonable alternative to the Ba-Ra sulphate
precipitation technology as it does not produce, as a by-product, large volumes of
radioactive sludge and it is affected less than ion exchange resins by the presence of IIA
elements.
Closing comments
The information presented in this chapter summarises some of the work and the
experience accumulated over the past 20 years on the biosorption of members of the
lanthanides, actinides and related elements. One could potentially include more
109
Biosorbents for Metal lons
Table 5.5 Radium biosorption immobilised biomass re-use
potential
elements such as Sr or Co, which are related to nuclear applications, or include
daughter products of the radioactive decay series of some of the elements discussed
above. However, these are outside the scope of the present chapter.
It is important to note that 20 years ago the mechanistic understanding of
biosorption was quite nebulous, and biosorption was mostly an interesting
phenomenon related mainly to microorganisms. Since then a substantial volume of
systematic work has been added. The engineering applications potential of the
phenomenon is being investigated, and numerous scientists and engineers are
working on the subject. The differentiation of the ‘biosorptive’ versus the
‘bioaccumulatory’ process has also been a positive step in the direction of the better
understanding of the underlying mechanisms in biosorptive phenomena. The
specificity of biosorption makes it an excellent candidate technology for industrial
applications where large volume, low concentration, complex ionic matrix waste or
process solutions need to be treated for the purpose of sequestering targeted elements.
Tertiary or polishing treatment applications are therefore good candidate
application areas. Finally, it should be noted that the phenomenon of biosorption
exists not only for inorganic ionic species but also for organic molecules, and this
observation opens up the opportunity to study the interactions between biosorption
and biodegradation for organic molecules of interest (Tsezos and Wang, 1991).
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